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Biochemical Dechlorination of Hexachloro-1,3-butadiene
Donny Lawrence James
A thesis presented for the degree of Doctor of Philosophy in Environmental Biotechnology
August 2009 Division of Science and Engineering
School of Biological Science and Biotechnology Murdoch University, Western Australia
Project supported by Environmental Biotechnology Co-operative Research Centre (EBCRC)
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I hereby declare that the thesis is my own account of my research and
contains as its main content work that has not been previously been
submitted for a degree at any university
Donny Lawrence James
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Table of Contents Project Summary…………………………………...……………………………..…....1 List of Abbreviations…………………………………………...……………………...3 Chapter 1 Literature Review…………………………..………………………………...6 General Introduction………………….………………………...……………………27 Aims of Thesis………………………….…………………………………………..…28 Chapter 2 Cyanocobalamin Enables Activated Sludge to Dechlorinate Hexachloro-1,3-butadiene to Non-Chlorinated Gases…………………………………………………….30 Chapter 3 Enrichment of Microorganisms Specific to Cyanocobalamin Reduction..…51 Chapter 4 Cyanocobalamin Enables Thermophilic Bacteria and Methanogens from Anaerobic Digested Effluent to Dechlorinate Hexachloro-1,3-butadiene to Non-Chlorinated Gases……....……………………………………………………...………...72 Chapter 5 Bacterially Produced Mediators Enhance The Dechlorination of Hexachloro-1,3-butadiene to Non-Chlorinated Gases & Investigation into why Dechlorination Stalls……………………………………………………………………….....................106 Chapter 6 The Use of Redox Potential to Monitor HCBD Dechlorination…….…….138 Chapter 7 Conclusions and Outlook…………….……………………………………157 Addendum…...……….…………………………………………………….…………171 References……………………………………………………..……………………..173 Curriculum Vitae……………………………..…….……..…..……………………..193 Acknowledgements………………………….………….…...…………………...…195
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Project Summary
Hexachloro-1,3-butadiene (HCBD) is a toxic aliphatic chlorinated hydrocarbon which is
widely used as a fungicide, herbicide and heat transformer fluid. HCBD is resistant to
microbial degradation and, therefore, persists in aquatic and soil environments
worldwide. In this thesis, the ability of non-specific bacteria from various sources to
dechlorinate HCBD in the presence of either acetate or lactate (as an electron donor) and
cyanocobalamin (as an electron shuttle) under different conditions was investigated.
Cultivating specific populations to reduce cyanocobalamin as a method to increase
HCBD dechlorination rate was investigated. Also, the factors responsible for HCBD
dechlorination and the stalling of dechlorination were studied. Lastly, redox potential
measurement during the microbial reductive dechlorination of HCBD for online detection
of ongoing dechlorination was evaluated.
Findings from the Project
Non-specific bacteria from activated sludge, anaerobic digested effluent from
municipal waste, piggery waste and sheep rumen content are able to dechlorinate
HCBD in the presence of cyanocobalamin to chlorine-free C4 gases in a
biochemical reaction.
Dechlorination was equated to the formation of completely dechlorinated end-
products from HCBD dechlorination.
Methanogens were found to be involved in HCBD dechlorination.
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Mediators rather than specific bacteria were responsible for the fast dechlorination
rates. Results suggest that activated sludge may release synthesized mediators into
the supernatant to enable enhanced HCBD dechlorination.
HCBD dechlorination can be monitored using oxidation reduction potential
(ORP). ORP has an effect on HCBD dechlorination rate.
Scientific Significance/Novelty
The most significant finding from this research is that it demonstrates chlorine-free end-
products in contrast with other studies in literature (Booker and Pavlosthasis, 2000;
Bosma et al., 1994) where dechlorination was equated with disappearance of HCBD into
bacterial biomass and the detection of partially dechlorinated gases such as
trichlorobutadiene. It also shows that, in contrast to literature where specific bacteria (i.e.,
pure strains/cultures) were commonly used for the dechlorination of polychlorinated
hydrocarbons, results from this thesis show that non-specific bacteria were able to
dechlorinate HCBD in the presence of cyanocobalamin at rates sufficiently high to be
considered for bioremediation projects. Moreover, results demonstrate that ORP can be
used to monitor HCBD dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
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List of Abbreviations
ACNQ 2-amino-3-carboxy-1,4-naphtoquinone
ADE Anaerobic Digested Effluent
Ag/AgCl Silver/Silver Chloride
AOX Organic Halogen Compounds
BD 1,3-butadiene
BES 2-bromo-ethane sulfonate
BTEX Benzene, Toluene, Ethylbenzene and Xylene
C4 gases Chlorine-free gases
CBD 1-chloro-butadiene
CF Chloroform
Co Cobalt
2-CP 2-chlorophenol
CC Cyanocobalamin
CPW Car Park Waste
CT Carbon Tetrachloride
DCBD Dichloro-1,3-butadiene
DCE cis-and trans-1,2-dichloroethene
2,4-DCP 2,4-dichlorophenol
1,2-D 1,2-dichloropropane
DPW Digested Pig Waste
DSMZ German Collection of Microorganisms and
Cell Cultures
EAg/AgCl Redox Potential (Silver/silver chloride
reference electrode)
EBCRC Environmental Biotechnology Co-Operative
Research Centre
GC Gas Chromatograph
GC-MS Gas Chromatograph-Mass Spectrometry
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H2O2 Hydrogen Peroxide
HCB Hexachlorobenzene
HCBD Hexachloro-1,3-butadiene
HRT Hydraulic Retention Time
KCl Potassium Chloride
KOH Potassium Hydroxide
K3Fe(CN)6 Potassium Ferricyanide
mV mill volts
MSD Mass Selective Detector
OCS Octachlorostyrene
ORP Oxidation Reduction Potential
Oxd Oxidation
PCB Polychlorinated Biphenyl
PCBD Pentachloro-1,3-butadiene
PCE Tetrachloroethene
PCP Pentachlorophenol
PHB Poly-ß-hydroxybutyrate
PTFE polytetrafluoroethylene
Red Reduction
SHE Standard Hydrogen Electrode
SRC Sheep Rumen Content
TCA Trichloroacetic Acid
1,1,2-TCE 1,1,2-trichloroethane
TCBD Trichloro-1,3-butadiene
TCButyne Trichloro-1-buten-3yne
TCE Trichloroethene
TCP 2,4,6-Trichlorophenol
TeCA 1,1,2,2-tetrachloroethane
TetraCBD Tetrachloro-1,3-butadiene
TRFLP Terminal Restriction Fragment Length
Polymorphism
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TSS Total Suspended Solids
VC Vinyl Chloride
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Chapter 1
LITERATURE REVIEW
Reductive Dechlorination of
Chlorinated Hydrocarbons in Anaerobic
Environments1
1 This chapter has been submitted to Soil and Sediment Contamination - An International Journal.
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1 Scope of Literature Review
Information regarding reductive dechlorination of chlorinated hydrocarbons in anaerobic
environments by anaerobes is highlighted in the first part. The processes involved in the
dechlorination of chlorinated hydrocarbons including rates and extent to which
dechlorination been observed in both pure and mixed have been highlighted in the second
part. The factors that are both essential and affect biodegradation of chlorinated
hydrocarbons are highlighted in the third part while some remediation technologies that
currently exist for the treatment of contaminated soil are discussed in the fourth part.
2 Introduction
Chlorinated hydrocarbons belong to either the aliphatic or aromatic group. In other
words, they are able to exist as either straight chain or polycyclic structures. Many
chlorinated hydrocarbons are produced as by-products during chemical synthesis, such as
dioxins, polychlorinated biphenyls (PCB), pentachlorophenol (PCP), tetrachloroethene
(PCE) and the fuel constituents namely, benzene, toluene, ethylbenzene and xylene
(BTEX) (Van Pée and Unversucht, 2003).
Chlorinated hydrocarbons are widely used in industrial applications as herbicides,
fungicides, heat transfer fluid, pharmaceuticals, flame retardants and solvents for
removing dirt and oils from clothes, engines and electronic parts (McCarthy, 1997; Van
Pée and Unversucht, 2003; Verschueren, 1996). A list of estimated annual production of
chlorinated hydrocarbons and major applications are listed in Table 1.1.
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Table 1.1 Estimated annual industrial production of chlorinated hydrocarbons and major applications. (Adapted from Fetzner, 1998). Chlorinated Production (x 103 tonnes) in: Year of Major use Hydrocarbon Western Europe United States Japan World estimate Chloromethanes Monochloromethane 230 250 50 1983 Production of silicones, tetramethyllead, 390 1992 methylcellulose; other methylation reactions Dichloromethane 210 270 35 1983 Degreasing agent; paint remover; pressure 162 1992 mediator in aerosols; extraction technology Trichloromethane 90 190 45 1983 Production of monochlorodifluoromethane
229 1991 (for the production of tetrafluoroethene, which is used for the manufacture of Hostaflon and Teflon); extractant for pharmaceutical products
Tetrachloromethane 250 250 75 1983 Production of trichloromonofluoromethane 143 1991 and dichlorodifluoromethane; solvent Chloroethanes
Monochloroethane 300 1984 Production of tetramethyllead; production of 67 1990 ethylcellulose; ethylating agent for fine
chemical production; solvent for extraction processes
1,1-Dichloroethane 200-250 1985 Feedstock for the production of 1,1,1-trichloroethane 1,2-Dichloroethane 8000 7000 2500 1985 Production of vinyl chloride; production of
7230 1992 chlorinated solvents such as 1,1,1- trichloroethane and tri- and tetrachloroethene; synthesis of ethylenediamines
1,1,1-Trichloroethane 150 300 1984 Dry cleaning; vapour degreasing; solvent for 327 1992 adhesives and metal cutting fluids; textile
processing 1,1,2-Trichloroethane 40 200-220 1984 Intermediate for the production of 1,1,1-
trichloroethane and 1,1-dichloroethene Chloroethenes Monochloroethene 5000 4000 13 600 1985 Production of poly (vinyl chloride) (PVC);
(vinyl chloride) 6000 1992 production of chlorinated solvents (primarily 1,1,1-trichloroethane)
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Chlorinated Production (x 103 tonnes) in: Year of Major use Hydrocarbon Western Europe United States Japan World estimate 1,1-Dichloroethene 150-200 1986 Basic material for poly (vinylidene chloride)
(vinylidene chloride) and its copolymers; production of 1,1,1- trichloroethane
Trichloroethene (TCE) 200 110 80 1984 Solvent for vapour degreasing in the metal industry and for dry cleaning; extraction solvent; solvent in formulations for rubbers, elastomers, and industrial paints
Tetrachloroethene (PCE) 220 600-700 1985 Solvent for dry cleaning, metal degreasing, 110 1992 textile finishing, dyeing, extraction
processes; intermediate for the production of trichloroacetic acid and some fluorocarbons
2-Chloro-1,3-butadiene 648 1983 Starting monomer for polychloroprene (chloroprene) rubber Chlorinated paraffins 350 1986 Plasticizers in PVC; flameproofing agents in
rubber, textiles, plastics; water-repellent and rot-preventive agents; elastic sealing compounds; paints and varnishes; metal-working agents (cutting oils); leather finishing
Nucleus-chlorinated aromatic hydrocarbons Monochlorobenzene 130 34 1981 Production of nitrophenol, nitroanisole,
chloroaniline, and phenylenediamine for the manufacture of dyes, crop protection products, pharmaceuticals, and rubber chemicals
1,2-Dichlorobenzene 23 9 1981 Production of 1,2-dichloro-4-nitrobenzene for the production of dyes and pesticides; production of disinfectants, deodorants
1,4-Dichlorobenzene 33 16 1981 Production of disinfectants, room deodorants; moth control agent; production of insecticides; production of 2,5-dichlrornitrobenzene for the manufacture of dyes; production of polyphenylenesulfide-based plastics
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Chlorinated Production (x 103 tonnes) in: Year of Major use Hydrocarbon Western Europe United States Japan World estimate Chlorinated toluenes 30 1983 Hydrolysis to cresol (monochlorotoluenes);
solvents for dyes; precursors for dyes, pharmaceuticals, pesticides, preservatives, and disinfectants
Chlorophenols 38-40 34-40 100 1986 Preparation of agricultural chemicals (herbicides, insecticides, fungicides), pharmaceuticals, biocides, and anthraquinone dyes
Chlorophenoxy- 200 1982 Herbicides alkanoic acids Side-chain chlorinated aromatic hydrocarbons
Chloromethylbenzene 80 160 1984 Production of plasticizers, benzyl alcohol, (benzylchloride) phenyl acetic acid, quaternary ammonium
salts, benzyl esters, triphenylmethane dyes, dibenzyl disulfide, benzylphenol, benzylamines
Dichloromethylbenzene 15 30 1984 Production of benzaldehyde (benzalchloride) Trichloromethylbenzene 30 60 1984 Production of benzoylchloride; production
(benzotrichloride) of pesticides, UV stabilizers, and dyes
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Due to their wide spread use, chlorinated hydrocarbons are commonly encountered
environmental pollutants, in estuarine sediments found throughout North America and
Western Europe (Durham and Oliver, 1983; Li et al., 1976; Rostad et al., 1989). These
chlorinated hydrocarbons enter the environment through production, commercial
application, and waste (Chaudhry and Chapalamadugu, 1991). The presence of these
chlorinated hydrocarbons in estuarine sediments leads to the bioaccumulation in fat
tissues of aquatic animals. This is due to high octanol: water coefficients exhibited by
those chlorinated hydrocarbons (Qiu and Davis, 2004).
Chlorinated hydrocarbons are recalcitrant, in that they are able to persist in natural
environments for long periods of time without microbial degradation (Alexander, 1985).
They also exhibit toxic and carcinogenic effects on humans. Due to this toxicity to
humans and other potential environmental hazards, the cleanup or destruction of these
chlorinated hydrocarbons from polluted sites is required (Berededsamuel et al., 1996).
Due to the high costs associated with large-scale remediation, bioremediation using
microorganisms serves as a cheaper alternative.
Anaerobic dechlorination is seen as an important mechanism in the bioremediation
(biodegradation) of chlorinated hydrocarbons resistant to aerobic degradation. Due to the
oxidised state conferred by the highly electronegative halogen substituents, the carbon
backbone of chlorinated hydrocarbons cannot be attacked by oxygen (Wohlfahrt and
Diekert, 1997). Therefore, biodegradation in the form of reductive dechlorination has
been shown to occur rather than oxidative dechlorination (Beurskens et al., 1995;
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Fathepure et al., 1988; Middeldorp et al., 1997; Pavlosthasis et al., 2003; Quensen 3rd et
al., 1988; Sahm et al. 1986; Vogel et al., 1987).
Anaerobic conditions naturally prevail in most contaminated groundwater and soils
(Zhang and Bennett, 2005) and in several instances aerobic degradation did not occur
without anaerobic dechlorination (Master et al., 2002). Suflita et al. (1982) noted that
there was a high affinity of anaerobes for chlorinated hydrocarbons. Lowe et al. (1993)
noted that this high affinity makes it possible to remove trace levels of chlorinated
hydrocarbons using anaerobes. Aerobic processes also require expensive oxygen delivery
systems (Baker and Herson, 1994). Thus, it appears that reductive dechlorination by
anaerobes were better suited for the removal of chlorinated hydrocarbons.
2 Reductive Dechlorination by Anaerobic Bacteria
In anaerobic environments, dechlorination occurs reductively (Suflita et al., 1982). This
process involves the removal of a halogen substituent (chlorine atom) from a chlorinated
molecule and the concurrent addition of a hydrogen atom (Mohn and Tiedje, 1992).
Several reports on the reductive dechlorination of chlorinated hydrocarbons by anaerobes
have been published (Dolfing and Beurskens, 1995; El Fantroussi et al., 1998; Fetzner
and Lingens, 1994; Holliger and Schraa, 1994; Kazumi et al., 1995; Kuhn and Suflita,
1989; Mohn and Tiedje, 1992). These reports also highlight the bioremediation potential
of contaminated sites using fermentative, sulfidogenic, methanogenic, homoacetogenic
and iron-reducing anaerobic consortia.
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There are three ways in which chlorinated hydrocarbons can be dechlorinated to chlorine-
free products by anaerobes. Firstly, anaerobes are able to utilize chlorinated hydrocarbons
as electron acceptors in a process known as halorespiration (Holliger and Schumacher,
1994). In halorespiration, anaerobes are able to metabolize energy through the
dechlorination of chlorinated hydrocarbons (McCarthy, 1997; Schmidt et al., 2000;
Schuhmacher et al., 1997; Wohlfahrt and Diekert, 1997). Secondly, anaerobes are able to
utilize chlorinated hydrocarbons as the sole source of carbon and energy (El Fantroussi et
al., 1998; Messmer et al., 1993). Thirdly, chlorinated hydrocarbons are dechlorinated via
enzymes or co-factors in a process known as co-metabolism. In this type of reaction,
there is no apparent energy generated for the benefit of the anaerobe involved (Holliger
and Schraa, 1994). Van Eekert and Schraa (2001) described these mechanisms in detail in
their review. Husain and Husain (2008) reviewed the use of several enzymes, from a
range of sources, for a range of applications including the removal of aromatic
compounds in the presence of redox mediators (Table 1.2).
2.1 Reductive Dechlorination by Pure Strains and Enrichment
Cultures
Dechlorination of chlorinated hydrocarbons has been reported by both pure strains and
mixed consortia. The rates of dechlorination by various pure strains and
enrichment/mixed consortia have been highlighted in Table 1.3 and 1.4. From Table 1.3
and 1.4, complete dechlorination of several chlorinated hydrocarbons was reported by
mixed consortia at high rates.
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Table 1.2 List of enzymes and their potential applications for the treatment of organic waste. (Adapted from Husain and Husain, 2008). Enzyme Source Applications References Alkylsulfatase Pseudomonas C12B Surfactant degradation Thomas and White, 1991 Azoreductase Pseudomonas sp. Decolorization of dyes Husain, 2006 Chitinase Serratia marcescens Bioconversion of shellfish waste Cosio et al., 1982 Chloro-peroxidase Caldariomyces fumago Oxidation of phenolic compounds Aitken et al., 1994 Cyanidase Alcaligenes denitrificans Cyanide decomposition Basheer et al., 1992 Haemoglobin Blood Removal of phenols and aromatic amines Chapsal et al., 1986 Laccase Several fungi, e.g.,Trametes versicolor, Removal of phenols, decolorization of Kraft Duran and Esposito, 2000; Duran
Fomas annosus bleaching effluents, binding of phenols and et al., 2002; Christian et al., aromatic amines with humus 2005; Husain, 2006
Lignin peroxidase Phanerochaete chrysosporium Removal of phenols and aromatic Christian et al., 2005; Husain, compounds, decolorization of kraft 2006 bleaching effluents
Lipase Various sources Improved sludge dewatering Thomas et al., 1993; Jeganathan et al., 2006
Lysozyme Bacterial Improved sludge dewatering Duran and Esposito, 2000; Manganese peroxidase Phanerochaete chrysosporium Oxidation of phenols and aromatic dyes Christian et al., 2005; Husain,
2006 Microperoxidase-11 Horse heart Horseradish Decolorization of dyes Husain, 2006 Peroxidase roots, tomato, white radish, turnip, bitter gourd Akhtar et al., 2005a, 2005b;
Oxidation of phenols, aromatic amines and dyes, Akhtar and Husain, 2006, Husain, decolorization of kraft bleaching effluents 2006; Kulshrestha and Husain,
2007; Matto and Husain 2007 Phosphatase Citrobacter sp. Removal of heavy metals Thomas et al., 1993 Proteases Bacterial, e.g., Bacillus subtilis, Solubilization of fish and meat remains Karam and Nicell, 1997
Pseudomonas marinoglutinosa Tyrosinase Mushroom Removal of phenols, aromatic amines Duran and Esposito, 2000; Duran
et al., 2002 Polyphenol oxidases Solanum melongena, Solanum Reactive and other dyes, dye effluents Khan and Husain, 2007
tuberosum Organophosphorus Bacterial and recombinant Organophosphorus compounds Shimazu et al., 2001; Mensee et Hydrolase al., 2005; Lei et al., 2005
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Toluene oxygenases Bacterial and recombinant Hydrocarbons Yeager et al., 2004; Johnson et
al., 2006 Parathione hydrolase Pseudomonas, Flavobacterium, Hydrolysis of organophosphate pesticides Caldwell and Raushel, 1991
Streptomyces sp.
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Table 1.3 Dechlorination of some chlorinated hydrocarbons by mixed consortia.
Electron Acceptor Electron Donor
Mediator Used
Complete Rate
Rate of Chloride released
Source Reference
Yes / No µmoles/ L culture/day (µmoles/ g
biomass/day)
µmoles/ L culture/day
Hexachloro-1,3-butadiene (HCBD) (400 nM)
None No HCBD → TCBD
3 x 10-5
1.8 x 10-4
Enrichment cultures from
Rhine River sediment Bosma et al., 1994
HCBD Methanol Lactate
None No HCBD → TCBD
0.3 (1.5)
1.8 Methanogenic enriched from contaminated estuarine
sediment (Bayou d’Inde -Lake Charles, Louisiana, USA)
Booker and Pavlostathis, 2000
Tetrachloroethylene (PCE) (10 µM) Lactate None Yes PCE → Ethane
88.8 (-)
355.2 Anaerobic sediment from the Rhine river and anaerobic
granular sludge (3:1)
De Bruin et al., 1992
3-chlorobenzoate (750 µM) Acetate Formate
None Yes
(54) - Methanogenic anaerobic granular sludge with
Desulfomonile tiejei, a benzoate degrader, and an H2-
utilizing methanogen
Ahring et al., 1992
2-chlorophenol (2-CP) (0.1% vol/vol.) Yeast Extract Peptone
None Yes 2-CP → CH4 +
CO2
0.18 0.18 Sewage sludge Dietrich and Winter, 1990
1,2-dichloropropane (1,2-D) Hydrogen None Yes 1,2-D → Propene
(7.2) - Enrichment cultures from Red Cedar Creek sediment
Löffler et al., 1997
Phenanthrene Acetate Yeast extract
Glucose
None Yes 24.23 - Aerobic mixed culture Yuan et al., 2001
2,4,6-Trichlorophenol (TCP) Glucose None No TCP →
3,4-dichlorophenol
(DCP)
1.01 1.01 TCP and DCP-adapted microbial communities
Chang et al., 1995
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Table 1.4 Dechlorination of some chlorinated hydrocarbons by pure strains.
Electron Acceptor Electron Donor Mediator Used
Complete Rate
Rate of Chloride released
Pure Strain Reference
Yes / No µmoles/ L culture/day (µmoles/ g
biomass/day)
µmoles/ L culture/day
Organism
PCE Methanol None No PCE → TCE
0.84 3.36 Methanosarcina sp. Fathepure et al., 1987
PCE (10 µM) 3-Chlorobenzoate Pyruvate
Rumen fluid
None No PCE → TCE
2.34 9.36 Dechlorinating bacterium DCB-1
Fathepure et al., 1987
Carbon tetrachloride (CT) (470 µM)
Fructose Hydroxoc-obalamin (OH-Cbl) (10 µM)
Yes CT → CO
188 (~ 99)
(Biomass
conc. - 2 g/L)
752 Acetobacterium woodii Hashsham and Freedman, 1999
2,4,6- trichlorophenol (TCP)
Pyruvate None No TCP → 2,4-
DiChlorophen-ol (DCP)
- 0.34 Dechlorinating bacterium DCB-2
Madsen and Licht, 1992
Hexachlorobenzene (HCB) (0.2 M)
Acetate None No HCB →
Pentachloro-benzene
0.75 0.75 Dehalococcoides sp. strain CBDB1
Jayachandran et al., 2003
Trichloroacetic Acid (TCA) (1 mM)
Acetate None No TCA →
Dichloroacetic acid
3.2 3.2 Trichlorobacter thiogenes De Wever et al., 2000
3,4,5,6-tetrachloro-2-methoxyphenol (Tetrachloroguaiacol) (10 µM)
PCP None Yes 0.05 0.05 Rhodococcus chlorophenolicus PCP-I
Häggblom et al., 1988
Benzene
Nitrate Phosphate
Sulfate
None Yes 3.84 - Pseudmonas sp. D8 strain Chang et al., 1997
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3 Factors Affecting Dechlorination of Chlorinated
Hydrocarbons
Generally, chlorinated hydrocarbons will undergo microbial degradation if
microorganisms are able to use the products as a substrate or intermediate for their
metabolic pathways. However, there are several factors essential for the dechlorination of
chlorinated hydrocarbons. Oxidative/reductive potential, temperature, solvent polarity
(solubility of chlorinated hydrocarbons), and the choice of electron donors and electron
mediators are some of these factors. These factors may influence or limit the rate or
extent a particular contaminant biodegrades. Thus, these factors play an important role to
both the growth of cultures and increasing dechlorination rates.
3.1 Oxidation/Reduction Potential
The oxidation/reduction (redox) potential is a measure of electron activity that indicates
the relative ability of a solution to accept or transfer electrons (Wilson et al., 1997).
Several studies have shown that low redox potential is required for dechlorination
(Beunink and Rehm, 1988; Middeldorp et al., 1997; Miller et al., 1997; Neumann et al.,
1996). Chlorinated hydrocarbons may migrate through sediments into deeper aquifers.
Even though redox potentials in contaminated sites are variable, low redox potentials
found in subsurface environments make it conducive for reductive dechlorination (Bose
and Sharma, 2002).
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Redox potential has been used in dechlorination studies. Redox potential has been used to
predict the dominant anaerobic degradation pathway of hexachlorobenzene in microbial
systems (Dolfing and Harrison, 1993).
Chlorinated hydrocarbons are usually sequentially reduced. Farwell et al. (1975)
observed that the reduction potential becomes more negative as the number of chlorine
atoms decrease. Kargina et al. (1997) also observed that the removal of the last chlorine
occurs at extreme negative potentials.
3.2 Temperature
The effect of temperature on dechlorination rates and on dechlorination pathways using
anaerobic cultures have been studied (De Bruin et al., 1992; Kengen et al., 1999; Kohring
et al., 1989; Wiegel and Wu, 2006; Wu et al., 1996; Wu et al., 1997; Zhuang and
Pavlosthasis, 1995).
Kohring et al., (1989) observed that the dechlorination rate of 2,4-dichlorophenol (2,4-
DCP) increased exponentially between 15 °C and 30 °C while Wu et al. (1997) noted that
temperature influenced the timing and the relative predominance of parallel pathways of
dechlorination. Wu et al. (1997) observed that para dechlorination of 2,3,4,6-
tetrachlorobiphenyl was dominant at 18 °C and 36 °C while, at all other temperatures
tested, ortho dechlorination dominated.
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An adaptation phase was required at lower temperatures. De Bruin et al. (1992) noted
that reductive dechlorination of PCE required an adaptation phase at 10 °C before similar
dechlorination was obtained as at 20 °C.
3.3 Solvent Polarity
Polarity may in part explain the limitation with which a particular solvent is soluble in
any other given solvent. Solubility of chlorinated hydrocarbons affects their
bioavailability and subsequent microbial dechlorination rate (Brusseau et al., 2001).
Dielectric constants are numerical values assigned to solvents to indicate their polarity.
The dielectric constants of various solvents and mixtures have been listed in Table 1.5.
Table 1.5 Dielectric constants of various solvents.
Solvent Dielectric constant, ε, @ 20 °C
Boiling Point (°C)
Hexachloro-1,3-butadiene (HCBD)
2.6 210 - 220 °C
1-Butanol 17.8 117.73 °C (390.9 K) Iso-propanol (propan-2-ol) 19.9 82.3 °C (355 K) Acetone 20.7 56.3 °C (329.4 K) Ethanol 24.6 78.4 °C (351.6 K) Methanol 32.9 64.7 °C (337.8 K) Glycerol 46 290 °C (554°F) Water 80 100 °C (373.15 K) (212 ºF)
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The overall dielectric constants of solvent mixtures can be calculated using equation 1.1.
The dielectric constants of several solvents from Table 1.5 (mixed 50:50) are listed in
Table 1.6.
εmix = εws fws + εss fss Equation (1.1)
where ε and f are the dielectric constant and volume fraction, respectively; and subscripts
mix, ws, and ss represent values for the mixture, weaker solvent, and stronger solvent,
respectively (Seedher and Bhatia, 2003).
Table 1.6 Dielectric constants of various solvent mixtures (50:50).
1-Butanol Ethanol Glycerol Acetone Iso-propanol (propan-2-ol)
Methanol Water
1-Butanol 21.2 31.9 19.25 18.85 25.35 48.9 Ethanol 21.2 35.3 22.65 22.25 28.75 52.3 Glycerol 31.9 35.3 33.35 32.95 39.45 63 Acetone 19.25 22.65 33.35 20.3 26.8 50.35 Iso-propanol (1-propanol)
18.85 22.25 32.95 20.3 26.4 49.95
Methanol 25.35 28.75 39.45 26.8 26.4 56.45 Water 48.9 52.3 63 50.35 49.95 56.45
It can be seen from Table 1.6 that the polarity of a non-polar solvent does not change
considerably when mixed with a polar solvent (such as water). Thus, the choice of a more
polar solvent mixed with water may keep the solution polar at the same time increase the
solubility of contaminants.
3.4 Electron Donors
Glucose, pyruvate, butyrate, succinate, acetate, formate, lactate, hydrogen, methanol and
ethanol have been observed to serve as electron donors for reductive dechlorination of
22
various chlorinated hydrocarbons (Carr and Hughes, 1998; DiStefano et al., 1992;
Dolfing, 1990; Gerritse et al., 1996; Holliger et al., 1993).
In general, the addition of an electron donor increased dechlorination rates due to the
increase in cell numbers (Doong et al., 1996; Zhuang and Pavlosthasis, 1995). Results
from several studies suggested that the dechlorinating capability of microorganisms to
dechlorinate chlorinated hydrocarbons was successful in the presence of properly selected
electron donors (Ballapragada et al., 1997; Carr and Hughes, 1998; Doong et al., 1996;
Gibson and Sewell, 1992; He et al., 2002; Holliger et al., 1992a). Similarly,
dechlorination was not observed when electron donors were either not supplied or
inappropriate (Ballapragada et al., 1997).
Electron donor concentrations were also reported to play a part in dechlorination. Zhuang
and Pavlosthasis (1995) noted that the higher the concentration of acetate added to their
PCE-dechlorinating culture, the higher the dechlorination rate.
3.5 Electron Mediators
Mediators are compounds that speed up the rate of reaction by shuttling electrons from
the biological oxidation of electron donors to the electron-acceptors (Husain and Husain,
2008). They are reduced and oxidised as a result of electron shuttling. Some examples of
mediators used are humic substances, anthraquinone-2, 6-disulfonate (AQDS) (Field and
Cervantes, 2005), indigo carmine (Nicholson and John, 2005) and cyanocobalamin
(Hashsham and Freedman, 1999). They have been applied in ferric iron reduction
23
(Lovely et al., 2004), microbial dechlorination (Van der Zee et al., 2001), azo dye
reduction (Dos Santos et al., 2004) and microbial fuel cells (Hernandez and Newman,
2001; Rabaey et al., 2005).
One such mediator, cyanocobalamin, has been used in dechlorination of chlorinated
hydrocarbons (Guerrero-Barajas and Field, 2005; Hashsham and Freedman, 1997; Kim
and Carraway, 2002). Cyanocobalamin was shown to enhance the rate of carbon
tetrachloride degradation both by specific anaerobic bacteria as shown with pure cultures
Acetobacterium woodii (Hashsham and Freedman, 1999) and anaerobic microbial
enrichments (Hashsham and Freedman, 1997). Assaf-Anid et al. (1992) noted that
cyanocobalamin reductively dechlorinated hexachlorobenzene to pentachlorobenzene and
hydrogen from water was shown to be the source of proton for replacement of chlorine
atom (Assaf-Anid et al., 1992).
In reducing conditions, the active transition metal in cyanocobalamin, cobalt(I) reduced
from the cobalt(III) binds to the chlorinated hydrocarbon and eliminates one chlorine
atom (Shey and Van der Donk, 2000; Wohlfarth and Diekert, 1997).. The regeneration of
cobalt (III) during this process allows the dechlorination step to be repeated (Kataky and
Wylie, 2001). An in-depth study of the electron transfer mechanism of cyanocobalamin
catalysed dechlorination of PCE was undertaken by Shey and Van der Donk (2000).
24
4 Remediation Technologies
Currently, several remediation technologies exist for the treatment of soil contaminated
with a range of contaminants. These technologies include solidification, asphalt batching,
encapsulation, vitrification, bioventing, phtytoremediation, thermal desorption, biopiles,
land farming, aeration, soil washing, soil flushing, soil vapour extraction and bioslurry
(Khan et al., 2004).
4.1 Contaminant Immobilisation Options
Solidification, asphalt batching and encapsulation are all a means of contaminant
containment and fixation and have been reported to have limited effectiveness against
organic contaminants (Mitchell and Potter, 1999). These technologies do not detoxify
contaminants but rather can be used to limit contaminant migration away (immobilise)
from heavily contaminated sites. These technologies do not involve the degradation of
contaminants.
4.2 In-situ Treatment Options
Vitrification, bioventing and pyhtoremediation are some examples of in-situ remediation
technologies. Vitrification is an in-situ physical decontamination method that does not
involve the use of biological agents. While this technology is an effective remediation
technology, it requires an extensive set-up of specialized equipment. Bioventing involves
pumping air into an unsaturated zone to stimulate the in-situ degradation of contaminants.
Bioventing is not appropriate to the remediation of sites where reductive dechlorination is
required. Phytoremediation takes advantage of plants to accumulate contaminants present
25
in soil. This technology, commonly used in heavy metal-contaminated sites, can be used
to stabilize, extract or transform contaminants. Small scale experiments have proven that
different contaminants can be transformed using phytoremediation (Nedunuri et al.,
2000). However, further studies are required to ascertain phytoremediation as an effective
and reliable remediation technology.
4.3 Treatment Options
4.3.1 Ex-situ
Aeration, soil washing, soil flushing, soil vapour extraction and the bioslurry are some
examples of in-situ remediation technologies.
Thermal desorption involves the ex-situ treatment of excavated contaminated soil. This
technology, while effective, is expensive and requires the use of large amounts of energy
and equipment (up to approximately US$400/ton) (FRTR, 1999). Landfarming involves
the excavation and spreading of contaminated soil on a treatment site while biopiles
involves the piling of excavated soil. Both these technologies aim to stimulate ex-situ
aerobic microbial activity. Like bioventing, both these technologies involve the
introduction of oxygen which would prevent the reductive dechlorination reaction.
Aeration, soil washing, soil flushing and soil vapour extraction involve the removal of
contaminant from soil into either an aqueous, solvent or gas phase. These technologies do
not involve either the destruction or detoxification of the contaminant and are simply a
means of extracting contaminants absorbed onto soil into an external medium. These
26
technologies serve as precursors for further treatment using incineration or
bioremediation.
4.3.2 In-situ
The main advantage of in-situ treatment compared to ex-situ treatment is that there are no
risks associated with excavation, namely volatilization or flushing. Therefore,
remediation in-situ by improving the conditions and/or the degradation potential in the
contaminated soil layer is preferred (Romantschuk et al., 2000). However, without a
method to retain/immobilise cyanocobalamin in the contaminated zone, it is highly
unlikely that any HCBD dechlorination would occur. Riser-Roberts (1998) and Reddy et
al. (1999) note that, ultimately, the successful treatment of a contaminated site depends
on the contaminant, site characteristics and regulatory requirements cost and time
constraints.
27
General Introduction
This research project was funded by Orica Pty Ltd in partnership with Environmental
Biotechnology Co-Operative Research Centre (EBCRC) to develop bioremediation
technologies that can be applied to polluted sites.
At a specific site located at the Botany Industrial Park (16-20 Beauchamp Road,
Matraville, Sydney 2036), 45 000 m3 of sandy soil was contaminated with chlorinated
hydrocarbons from solvent plants. This sandy contaminated soil, labeled Car Park Waste
(CPW), was encapsulated in a hyperlon liner and covered with a bitumen car park. The
CPW was primarily contaminated with HCBD at a concentration of 3225 mg/kg. Other
contaminants, namely octachlorostyrene (OCS), hexachlorobenzene (HCB) and
tetrachloroethene (PCE) were present in 2 - fold lower concentrations. The primary focus
was the removal of HCBD from the CPW.
The reductive dechlorination of HCBD using biomass was the primary subject of this
study. Microbial reductive dechlorination of HCBD would represent substantial cost
savings compared to the expensive thermal desorption method.
28
Aims of Thesis
Dechlorination rates are higher and the extent of dechlorination is more complete when
mixed microbial cultures are used as reported in literature (Table 1.3 and 1.4.). No
anaerobic pure culture, specific enzyme able to dechlorinate HCBD or the dechlorination
mechanism was reported at the commencement of the project.
The overall aim of the thesis was to develop a biochemical method to dechlorinate HCBD
using the most appropriate current knowledge from literature. It was not to identify or
isolate a specific bacterium or an enzyme to dechlorinate HCBD due to sterility concerns
when applied on-site.
The specific aims of this study were
1. To develop a method to reductively dechlorinate HCBD using biomass (Chapter
2).
2. Assess the extent and rates of biotransformation of HCBD incubated with the
various biological inoculants (e.g. Activated Sludge, Anaerobic Digester Sludge,
Digested Piggery Waste, Sheep Rumen Content and Anaerobic Digested Effluent)
in the presence of an electron mediator (cyanocobalamin) via Gas
Chromatography/ Mass Spectrometry analysis (Chapter 2 and 4).
29
3. Manipulate physical factors to enhance the rate of dechlorination (Chapter 2, 4
and 5).
4. Set-up, monitor and computer control a bioreactor able to enrich bacteria that use
oxidized cyanocobalamin as electron acceptor and acetate as the electron donor
(Chapter 3).
5. Set-up an on-line monitored bioreactor with bacteria able to reductively
dechlorinate HCBD (Chapter 6).
30
Chapter 2
Cyanocobalamin Enables Activated Sludge to
Dechlorinate
Hexachloro-1,3-butadiene to Non-Chlorinated Gases†
† A major part of this chapter was published in Bioremediation Journal 12(4), 177 - 184.
31
1 Introduction
Hexachloro-1,3-butadiene (HCBD) (Fig. 2.1) is a toxic, aliphatic chlorinated
hydrocarbon. It is carcinogenic, mutagenic and fetotoxic (Anonymous, 1992). It is
produced as a by-product from the production of tetrachloroethene, trichloroethene and
carbon tetrachloride (Booker and Pavlostathis, 2000). It is a pollutant in sediment
samples throughout North America and Western Europe (Durham et al., 1983; Li et al.,
1976; Rostad et al., 1989) where it was used as a chlorine recovery solvent in the
production and processing of rubber. It was also used as fungicide, herbicide and heat
transformer fluid (Verschueren, 1996). HCBD is hydrophobic with low water solubility
(3.20 mg/L at 25 ºC). Due to its high octanol-water partition coefficient (log Kow = 4.78)
(Mackay et al., 1993), HCBD tends to accumulate in the lipids of aquatic organisms and
is resistant to microbial degradation. This explains its persistence in the environment
(Murray and Beck, 1989; Pereira et al., 1988). Due to the highly oxidised state of the
carbon atoms in HCBD, and to the highly electronegative halogen substituents,
biodegradation in the form of reductive dechlorination is more likely to occur than the
more traditional biodegradation via oxidative processes (Pavlostathis et al., 2002).
Figure 2.1 Molecular structure of Hexachloro-1,3-butadiene (HCBD).
Cl
Cl
Cl
Cl
Cl
Cl
32
In general, the capacity of bacteria to reductively dechlorinate other polychlorinated
compounds (e.g., polychlorinated biphenyls, halogenated alkanes and alkenes) has been
described (Dolfing and Beurskens, 1995; El Fantroussi et al., 1998). Low et al. (2007)
noted that no HCBD dechlorination was observed by pure cultures of bacteria
(Dehalococcoides, Dehalobacter and Desulfitobacterium).
Cyanocobalamin has been shown to enhance the rate of carbon tetrachloride degradation
both by specific anaerobic bacteria as shown with pure cultures Acetobacterium woodii
(Hashsham and Freedman, 1999) and anaerobic microbial enrichments (Hashsham and
Freedman, 1997).
This study aimed to investigate the potential of non-specific bacteria, from activated
sludge, to dechlorinate HCBD given either acetate or lactate as an electron donor and
with cyanocobalamin as an electron shuttle. It also quantifies dechlorination rates, and the
effect of environmental conditions on the efficiency of cyanocobalamin mediated,
bacterial reductive dechlorination of HCBD.
2 Experimental Procedures
2.1 Medium Composition
Return activated sludge (100 mg/mL dry weight) from the local wastewater treatment
plant (Water Corporation - SBR reactor at Woodman Point, Western Australia) was
obtained for use as inoculum.
33
The composition of artificial wastewater basal medium used was based on DSMZ 334
medium (German Collection of Microorganisms and Cell Cultures (DSMZ), 1983). The
basal medium contained (per litre): 1.0 g NH4Cl, 0.3 g KH2PO4, 0.6 g NaCl, 0.1 g
MgCl2.2H2O, 0.08 g CaCl2.2H2O, 3.5 g KHCO3, 1.0 mg resazurin, 10.0 mL vitamin
solution and 5.0 mL trace element solution.
The vitamin solution was based on DSM 141 medium and contained (per litre): 2.0 mg
biotin, 2.0 mg folic acid, 10.0 mg pyridoxine hydrochloride, 5.0 mg thiamin
hydrochloride, 5.0 mg riboflavin, 5.0 mg nicotinic acid, 5.0 mg DL-calcium pantothenate,
0.1 mg cyanocobalamin, 5.0 mg p-aminobenzoate and 5.0 mg lipoic acid.
The trace element solution was based on DSM 318 medium (per litre): 12.8 g
nitrilotriacetic acid, 1.35 g FeCl3.6H2O, 0.1 g MnCl2.4H2O, 0.024 g CoCl2.6H2O, 0.1 g
CaCl2.2H2O, 0.1 g ZnCl2, 0.025 g CuCl2.2H2O, 0.01 g H3BO3, 0.024 g Na2MoO4.4H2O,
1.0 g NaCl, 0.12 g NiCl2.6H2O, 4.0 mg Na2SeO3.5H2O, 4.0 mg Na2WO4.2H2O. The trace
element solution was adjusted to pH 6.5 with 1 M KOH.
2.2 Dechlorination Experiments
Acetate (40 mM) and lactate (40 mM) were added as electron donors and
cyanocobalamin (0.05 to 0.8 mM) (Sigma catalog No. 68-19-9) was supplied as the
electron shuttle. Unless otherwise specified, HCBD (Sigma catalog No. 112-19-4) was
added as the electron acceptor at a concentration of 1 mM. The electron source was added
in excess relative to the electron transfer mediator and the pollutant as expected in natural
34
environments (Schwarzenbach et al., 1990). In one experiment, ethanol was added at
concentrations of 1 % to 8 % (vol./vol.). In the experiment testing the effect of agitation
(Fig. 2.11), cultures were agitated at 180 revolutions per minute in a 40-litre WiseBath®
water bath kept at 55 °C. In the experiment testing the effect of solvents, acetone, 1-
butanol, propan-2-ol or glycerol were added at concentrations of 0 mM, 160 mM, 320
mM, 640 mM and 1.28 M. Thirty mL of activated sludge was incubated with electron
donor, shuttle and acceptor, and topped up with artificial wastewater basal medium to a
final volume of 60 mL in 100 mL Wheaton glass serum bottles (Sigma catalog No. Z11,
400-6). Serum bottles were sealed with rubber stoppers (Bellco catalog No. BEL 2048-
11800) and the headspace flushed with N2:CO2 (80:20) gas. The same gas was used to
purge all solutions to remove oxygen. Cultures were incubated either at 37 °C or 55 °C.
Triplicate treatments were set-up for all experiments.
2.3 Sampling and Analyses
Hydrocarbons and chlorinated hydrocarbons were analysed by headspace sub-sampling at
each sampling time (Day 0, 5, 8, 14, 20, 27, 35 and 40). Three hundred µL of the culture
vessel headspace was removed via a gas-tight syringe (Hamilton, 500 µL) and injected
onto a Hewlett Packard 5890 series II Gas Chromatograph (GC) equipped with a
split/splitless inlet operated in splitless mode and a Hewlett Packard 5972 mass selective
detector (MSD). Chlorinated hydrocarbon separation was achieved on a DB-17MS
column (30 m x 0.25 mm (internal diameter) x 0.25 µm film thickness) (J&W scientific,
Folsom, CA, USA) using helium as the carrier gas at a flow rate of 1 mL/min. The
column was subjected to the following temperature program: 50 oC for 1 min and then
35
increased by 15 oC /min to a final temperature of 300 oC and then held for 5 min.
Chlorine-free hydrocarbons were separated on a GASPRO (GS) column (60 m x 0.32
mm internal diameter) (J&W scientific, Folsom, CA, USA) with the same temperature
program as specified for chlorinated hydrocarbons. The MSD was operated in scan mode
across the range of 49 - 300 amu in both cases. Spectral scans were compared to scans
from Fattore et al. (1996). The detection limit was 0.02 nmoles (4.08 µM).
Quantitation of chlorine-free hydrocarbons namely 1,3-butadiene, 1-buten-3-yne and 1,3-
butadiyne (collectively labelled C4 gases) was achieved by using a calibration curve
derived from analysis of standard gas samples of 1,3-butadiene in nitrogen over the
concentration range of 0.1 - 1.6 mmol/L. Due to the lack of availability of 1,3-butadiyne
and 1-buten-3-yne as pure substances, an assumption was made that the MSD response to
these compounds in full scan mode will be similar to that of 1,3-butadiene.
3 Results and Discussion
3.1 Effect of Cyanocobalamin on HCBD Dechlorination
When freshly collected activated sludge was incubated in the presence of HCBD, the
analysis of filtered samples of test vials suggested that HCBD disappeared over a
relatively short incubation time of approximately 4 - 6 weeks (data not shown). However,
solvent extraction of the biomass revealed that HCBD was absorbed into the biomass.
Hence, rather than monitoring HCBD disappearance, the appearance of the dechlorinated
endproducts was monitored in the gaseous phase using the GC/MS. Lower rates of
36
HCBD dechlorination were obtained with similar experiments using anaerobic digester
sludge instead of activated sludge (data not shown).
It was observed by other authors (Booker and Pavlostathis, 2000; Bosma et al., 1994;
Low et al., 2007) that the microbial dechlorination of HCBD led to the formation of
partly dechlorinated compounds such as pentachlorobutadiene, tetrachlorobutadiene,
trichlorobutadiene and dichlorobutadiene. In contrast to observations made from other
studies, in our study, completely dechlorinated endproducts were also formed, namely, 1-
buten-3-yne, 1,3-butadiene and 1,3-butadiyne (Fig. 2.2 - 2.4). The sum of the three
completely dechlorinated C4 gases was used to quantify HCBD dechlorination.
Figure 2.2 Mass Spectrum of 1,3-butadiyne.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
HH
37
Figure 2.3 Mass Spectrum of 1,3-butadiene. Figure 2.4 Mass Spectrum of 1-buten-3-yne.
The initial accumulation of these C4 gases demonstrated a rate of HCBD dechlorination
of 0.25 µmoles/L culture/day (0.02 µmoles/g biomass/day) which was about 10 times
faster than previously described biologically driven partial dechlorination in anaerobic
sediments (Bosma et al., 1994) and comparable to the HCBD disappearance rates found
for cyanocobalamin supplemented methanogenic enrichment cultures (Booker and
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
H
H
H
H
H
H
H
H
H
H
H
H
H H
H
H
H H
H
H
38
Pavlostathis, 2000). Dechlorination did not occur in the absence of cyanocobalamin,
when activated sludge was autoclaved or when the activated sludge was inhibited by
sodium azide.
The most significant difference in the dechlorination reaction described here, compared
to the literature, is the fact that chlorine-free C4 gases were formed as the major
endproducts (Fig. 2.5). In previously described dechlorination of HCBD, only partly
dechlorinated by-products such as trichlorobutadiene were detected (Booker and
Pavlostathis, 2000; Bosma et al., 1994).
Figure 2.5 Gas chromatogram of gases detected from headspace of activated sludge cultures (100 mg/mL) incubated with acetate (40 mM), cyanocobalamin (0.4 mM), and HCBD (10 mM) taken on Day 40.
If completely dechlorinated endproducts are formed, this indicates that the reduction
reaction is rather non-specific (unlike enzyme catalysed reactions) as not only HCBD but
also its partly dechlorinated byproducts have been reacting. This apparently non-specific,
cyanocobalamin mediated dechlorination was also obtained when biomass was replaced
10 15 20
1,3 butadiyne
1,3 butadiene1-buten-3-yne
Hexachloro-1,3-butadiene
Trichloro-1,3-butadiene
Trichloroethene
Tetrachloroethene
Dichloro-1,3-butadiene
Chloro-1,3- butadiene
Dichloroethene
10 15 20
1,3 butadiyne
1,3 butadiene1-buten-3-yne
Hexachloro-1,3-butadiene
Trichloro-1,3-butadiene
Trichloroethene
Tetrachloroethene
Dichloro-1,3-butadiene
Chloro-1,3- butadiene
Dichloroethene
Time (min)
Area Counts
100000
50000
39
with a chemical reducing agent. In incubations where biomass was replaced with a
chemical reducing agent, elemental zinc, along with cyanocobalamin, HCBD was also
dechlorinated to C4 gases. This biomass-free dechlorination demonstrates the non-
specific nature of HCBD dechlorination in the presence of cyanocobalamin. Should
specific enzymes that catalyse each step of the dechlorination reaction be required, then
partly dechlorinated products would be expected to accumulate rather than C4 gases.
Bosma et al. (1994) also found that the completely dechlorinated product, 1-buten-3-yne,
was formed from HCBD dechlorination by Titanium (III) citrate in the presence of
hydroxocobalamin. Figure 2.6 shows possible HCBD degradation pathways adapted from
Bosma et al. (1994).
40
Figure 2.6 HCBD dechlorination pathways. The left-hand flow shows sequential dechlorination while the right-hand flow shows dechlorination from dihalo elimination (Adapted from Bosma et al., 1994). Dihalo elimination is defined as the removal of two chlorine atoms from adjacent carbon atoms with the formation of an additional bond between the carbon atoms (Mohn and Tiedje, 1992).
41
In general, it was found that during biological dechlorination C4 gases accumulated after
an initial lag phase. A possible explanation could be that, initially, there were interfering
substances or alternative electron acceptors, with a more positive redox potential, that
could accept electrons in place of oxidised cyanocobalamin. Furthermore, it was also
found that reactions stalled after approximately 20 days.
There was significant variation in rates observed with different batches of activated
sludge. Our results intend to demonstrate trends rather than absolute rates obtained with
different samples of activated sludge.
3.2 Effect of Cyanocobalamin Concentration on HCBD Dechlorination
There is a high cost associated with the use of cyanocobalamin (approximately A$400/ 5
grams). For large-scale bioremediation purpose, the high cost may rule out the use of
cyanocobalamin. In the interest of cost savings, several concentrations of
cyanocobalamin lower than 1 mM were tested for their effect on HCBD dechlorination
rates. Cyanocobalamin concentrations used were 0, 0.05, 0.1, 0.2, 0.4 and 0.8 mM.
Results show that the highest dechlorination rate was observed at 0.4 mM
cyanocobalamin (Fig. 2.7). At a concentration of 0.8 mM cyanocobalamin, there was an
inhibitory effect on HCBD dechlorination. Using 0.4 mM cyanocobalamin instead of 1
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
42
mM represents a 2.5 - fold cost savings. Again, no dechlorination was observed in the
absence of cyanocobalamin.
0
2
4
6
8
10
12
0 0.05 0.1 0.2 0.4 0.8
Concentration of cyanocobalamin (mM)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cul
ture
)
Figure 2.7 Effect of cyanocobalamin concentration on the concentration of C4 gases (µmoles/L culture) from HCBD dechlorination by anaerobically incubated activated sludge cultures (100 mg/mL TSS) in the presence of acetate (40 mM), and HCBD (1 mM) at 37 °C on Day 35. Cyanocobalamin concentrations used were 0, 0.05, 0.1, 0.2, 0.4, and 0.8 mM.
A number of other parameters were manipulated with the purpose of increasing HCBD
solubility and dechlorination rates.
3.3 Effect of Temperature on HCBD Dechlorination
Temperature is not only known to affect biological reaction rates (2 - fold increase for
every 10 °C rise) but also the solubility of the target substance. An increase in
4 4
8
12
16
20
24
43
temperature from 37 °C to 55 °C caused a 3 - 4 fold increase in dechlorination rates (Fig.
2.8). The higher dechlorination rate* at 55 °C compared to 37 °C was unexpected as the
activated sludge bacteria used had developed under mesophilic conditions. It is surprising
that mesophilic bacteria could survive in thermophilic conditions. However, Marchant et
al. (2002) observed the existence of highly thermophilic bacteria in cool soil
environments.
0
5
10
15
20
25
30
35
40
45
0 10 20 30 40Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 2.8 Effect of temperature on the concentration of C4 gases (µmoles/L culture)* from HCBD dechlorination by anaerobically incubated activated sludge cultures (100 mg/mL TSS) in the presence of acetate (40 mM), cyanocobalamin (0.4 mM), and HCBD (1 mM). Temperatures were set at 22 ºC (●), 37 ºC (▲), 45 ºC (■), and 55 ºC (○).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
44
3.4 Effect of Ethanol Concentration on HCBD Dechlorination
The slow reaction rates have been identified as one of the main drawbacks of HCBD
dechlorination. One likely reason is the poor availability of HCBD to the biochemical
reduction process due to its low solubility in water (3.20 mg/L at 25 ºC). It was thought
that one way to increase HCBD solubility and its dechlorination rates was by adding
ethanol to the reaction mixtures. An added advantage is that the ethanol can be used as a
source of electrons in the biotic reductive dechlorination. The presence of even low
concentrations of ethanol increased HCBD solubility until the HCBD became fully
miscible at levels of ethanol in excess of 15 % (Fig. 2.9).
0
1
2
3
4
5
6
7
0 20 40 60 80 100
Ethanol in water (%)
HC
BD
(mm
oles
/L)
0
100
200
300
400
500
HC
BD
(mm
oles
/L)
Figure 2.9 Effect of increasing amounts of ethanol in water on the solubility of HCBD (mmoles/L). 0 - 50 % (▲ - refers to the y-axis on the left) and 60 - 85 % (■ - refers to the y-axis on the right).
45
The highest rate of HCBD dechlorination when compared to the control sample was
observed with an addition of 2 % (vol. /vol.) ethanol (Fig. 2.10). While higher levels
(e.g., 8 % (vol. /vol.)) of ethanol enhanced HCBD solubility (Fig. 2.9), it is also known to
inhibit bacterial metabolism and hence interfered with the dechlorination process (Fig.
2.10).
0
2
4
6
8
10
12
14
0 1 2 4 8
Concentraion of Ethanol (% vol./vol.)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 2.10 Effect of ethanol concentration on the concentration of C4 gases (µmoles/L culture)* from HCBD dechlorination by anaerobically incubated activated sludge cultures (100 mg/mL TSS) in the presence of acetate (40 mM), cyanocobalamin (0.4 mM), and HCBD (1 mM) at 55 °C on Day 35. Ethanol concentrations used were 0 %, 1 %, 2 %, 4 %, and 8 % (vol./vol.).
Along with ethanol, different concentrations of acetone, 1-butanol, propan-2-ol and
glycerol were also tested for their ability to increase the solubility and hence the rate of
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
4
8
12
16
20
24
28
46
HCBD dechlorination. The highest rates obtained were still from cultures incubated with
ethanol (2 % vol./vol.), which were 10 - fold higher than the rates* observed with all
other incubations (data not shown).
3.5 Effect of Biomass Concentration on HCBD Dechlorination
Previous results revealed that the cyanocobalamin mediated HCBD dechlorination by
activated sludge was optimal in the presence of 2 % (vol./vol.) ethanol and elevated
temperature (up to 55 °C). However, the impact of activated sludge biomass
concentration was unclear. In general, microbially catalysed reaction rates increase
proportionally to the biomass concentration. To clarify the relationship between HCBD
dechlorination rate and biomass concentration, different activated sludge concentrations
were tested. As expected, increased biomass levels enabled a faster dechlorination
reaction. The initial dechlorination rates* increased proportionally with the biomass up to
100 g/L of biomass. This linear trend flattened off between 100 and 200 g/L indicating
factors other than the biomass concentration became limiting for the reaction rate (Fig.
2.11). Such other factors could include the diminishing mass transfer due to increasing
viscosity caused by the higher biomass levels. Also, if the reaction was limited by the
availability of oxidized or reduced mediators, a further increase in biomass concentration
could not enable faster rates.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
47
0
1
2
3
4
5
0 50 100 150 200 250
Biomass Concentration (g/L)
Rat
e of
C4
gase
s ac
cum
ulat
ed o
ver 8
day
s .
(µm
oles
/L c
ultu
re/d
ay)
Figure 2.11 Effect of biomass concentration (g TSS/L culture) on the rates of C4 gases accumulated (µmoles/L culture/day) from HCBD dechlorination by anaerobically incubated activated sludge cultures in the presence of acetate (40 mM), cyanocobalamin (0.4 mM), ethanol (2 % vol./vol.), and HCBD (1 mM) at 55 °C.
3.6 Effect of Agitation on HCBD Dechlorination
Agitation enables the mass transfer of compounds in solution. It was believed that mixing
would increase bacterial contact with cyanocobalamin and HCBD in solution, and
thereby increase HCBD dechlorination. Both HCBD and bacteria are essentially non-
soluble in water. By providing mixing, the mass transfer of soluble species from bacterial
cells to HCBD would be expected to be increased.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
48
Agitation increased the HCBD dechlorination rate (Fig. 2.12). A 4 - fold increase in both
the rate and total concentration of HCBD dechlorination* was observed in agitated
cultures compared to stationary cultures. Again, HCBD dechlorination only occurred in
the presence of cyanocobalamin.
0
5
10
15
20
25
30
35
40
45
0 10 20 30 40
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cul
ture
)
Figure 2.12 Effect of agitation on the concentration of C4 gas (µmoles/L culture)* from HCBD dechlorination by anaerobically incubated activated sludge cultures (100 mg/mL TSS) in the presence of acetate (40 mM), cyanocobalamin (0.4 mM), ethanol (2 % vol./vol.), and HCBD (1 mM) at 55 °C. Cultures were either incubated with agitation (▲) or no agitation (■).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
49
A stimulating effect of agitation on the dechlorination on other chlorinated solvents has
been shown (Khleifat, 2006). In contrast, other studies have either shown a decrease in
dechlorination rates or no dechlorination with agitation (Chang et al., 1998; Juteau et al.,
1995; Yuan et al., 1999). On the whole, the effect of agitation on microbial
dechlorination rates seems to be complex as an increased substrate supply can also mean
increased inhibition if the substrate or metabolites are toxic.
4 Conclusion
This chapter shows that the dechlorination of HCBD to C4 gases is possible with
anaerobically incubated activated sludge and cyanocobalamin as the electron shuttle. The
number of cyanocobalamin reducing bacteria seems to be one of the limiting factors in
the dechlorination rate. The highest overall rate achieved was approximately 4 µmoles/L
culture/day with an agitated culture incubated with ethanol (2 % (vol./vol.)) at 55 °C.
Further investigations into why reactions start after an initial lag phase and stall after
approximately 20 days would be beneficial for the implementation of bioremediation.
Overall, results imply that cyanocobalamin acts as an electron shuttle between bacteria
and HCBD (Fig. 2.13). This concept, proposing cyanocobalamin to be the electron carrier
(mediator) between the bacterial substrate oxidation and the reduction of HCBD, is
supported by the observations that a) chemically reduced cyanocobalamin can cause
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
50
reductive dechlorination (Kim and Carraway, 2002) and b) cyanocobalamin stimulates
the microbial reductive dechlorination of carbon tetrachloride (Guerrero-Barajas and
Field, 2005; Hashsham and Freedman, 1999).
Figure 2.13 Schematic representation of the production of chlorine-free C4 gases from HCBD dechlorination by activated sludge in the presence of an electron donor (acetate) and cyanocobalamin (CC).
Organic electron donor
(e.g. lactate)
Bacterium
CC oxd
CC red
HCBD
Chlorine free C4 gases
Cl
Cl
Cl
Cl
Cl
Cl
H H
H
H
HH
H
H
H
H
H
H
1,3-butadiyne
1,3-butadiene
1-buten-3-yne
Organic electron donor
(e.g. lactate)
Bacterium
CC oxd
CC red
HCBD
Chlorine free C4 gases
Cl
Cl
Cl
Cl
Cl
Cl
H H
H
H
HH
H
H
H
H
H
H
1,3-butadiyne
1,3-butadiene
1-buten-3-yne
(Acetate)
51
Chapter 3
Enrichment of Microorganisms Specific to
Cyanocobalamin Reduction‡
‡ This chapter has been submitted to Journal of Biotechnology.
52
1 Introduction
It was found in a previous study (Chapter 2) that biological Hexachloro-1,3-butadiene
(HCBD) dechlorination occurred only in the presence of cyanocobalamin. In addition, the
HCBD dechlorination rate was related to biomass concentration (Chapter 2). The fact
that biomass concentration increased the rate of HCBD dechlorination demonstrated the
need for suitable cyanocobalamin reducing bacteria if the process was to be applied for
field trials of contaminated sites.
In tests where biomass and its electron donor, acetate were replaced with the chemical
reducing agent, zinc, along with cyanocobalamin, HCBD was dechlorinated to C4 gases
(Chapter 2). Since the interaction between reduced cyanocobalamin and HCBD is a
chemical process that occurs instantaneously, the key step in enhancing the rate of HCBD
dechlorination would be to increase the rate of biological cyanocobalamin reduction. In
the case of adequate substrate supply, the biological cyanocobalamin reduction rate
depends largely on the concentration of bacteria capable of using cyanocobalamin as the
terminal electron acceptor. However, the growth of those bacteria that can enable
reductive dechlorination by keeping the cyanocobalamin in a reduced state is difficult if it
had to be coupled to the dechlorination reaction, as this reaction is extremely slow and it
would take years to produce sufficient biomass for field applications.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
53
It is the goal of this chapter to develop and test a method that specifically selects and
enriches cyanocobalamin-reducing bacteria from activated sludge bacteria, without
requiring the presence of HCBD. A reactor was operated with the aim to enrich bacteria
that use oxidized cyanocobalamin as electron acceptor and acetate as the electron donor.
Biologically reduced cyanocobalamin exiting the reactor was re-oxidised with oxygen
and re-circulated back into the reactor. It was postulated that provided the electron donor
and oxidised cyanocobalamin were not limiting, over time, the continuous reduction
activity by activated sludge bacteria would result in a consortium of cyanocobalamin-
reducing bacteria.
Currently, information on the biological reduction of a mediator such as cyanocobalamin
followed by its re-oxidation is lacking. However, several studies involving microbial fuel
cells have shown that biologically reduced mediators can be subsequently re-oxidised
using electrodes for sustained electricity production (Bullen et al., 2006; He and
Angenent, 2006; Lovley, 2006; Wilkinson et al., 2006).
2 Experimental Procedures
2.1 Medium Composition
Return activated sludge (100 mg/mL dry weight) from the local wastewater treatment
plant (Water Corporation - SBR reactor at Woodman Point, Western Australia) was
obtained for use as inoculum.
54
The composition of artificial wastewater basal medium used was based on DSMZ 334
medium (German Collection of Microorganisms and Cell Cultures (DSMZ, 1983)). The
basal medium contained (per litre): 1.0 g NH4Cl, 0.3 g KH2PO4, 0.6 g NaCl, 0.1 g
MgCl2.2H2O, 0.08 g CaCl2.2H2O, 3.5 g KHCO3, 1.0 mg resazurin, 10.0 mL vitamin
solution and 5.0 mL trace element solution.
The vitamin solution was based on DSM 141 medium and contained (per litre): 2.0 mg
biotin, 2.0 mg folic acid, 10.0 mg pyridoxine hydrochloride, 5.0 mg thiamin
hydrochloride, 5.0 mg riboflavin, 5.0 mg nicotinic acid, 5.0 mg DL-calcium pantothenate,
0.1 mg cyanocobalamin, 5.0 mg p-aminobenzoate and 5.0 mg lipoic acid.
The trace element solution was based on DSM 318 medium (per litre): 12.8 g
nitrilotriacetic acid, 1.35 g FeCl3.6H2O, 0.1 g MnCl2.4H2O, 0.024 g CoCl2.6H2O, 0.1 g
CaCl2.2H2O, 0.1 g ZnCl2, 0.025 g CuCl2.2H2O, 0.01 g H3BO3, 0.024 g Na2MoO4.4H2O,
1.0 g NaCl, 0.12 g NiCl2.6H2O, 4.0 mg Na2SeO3.5H2O, 4.0 mg Na2WO4.2H2O. The trace
element solution was adjusted to pH 6.5 with 1 M KOH.
2.2 Incubation Conditions
2.2.1 Cyanocobalamin Reducing Bioreactor Set Up
A two-stage bioreactor consisting of an anaerobic chamber (360 mL) and an aerobic
recirculation loop was constructed for the proliferation of cyanocobalamin-reducing
bacteria (Fig. 3.1). The anaerobic chamber was responsible for the reduction of
55
cyanocobalamin while recirculation through silicon tubing allowed the maximal
oxidation of the reduced cyanocobalamin by permeated oxygen. Oxygen is known to
readily diffuse through the wall of silicon tubing.
56
Figure 3.1 Schematic diagram of cyanocobalamin-reducing bioreactor set-up. This set-up prevented dissolved oxygen of entering into the cyanocobalamin reducing bioreactor.
Dissolved Oxygen Probe
Air
Cyanocobalamin-Reducing Bioreactor
Sampling port (Outlet)
Redox probe
Cyanocobalamin-Oxidation (Silicon) Loop
Sampling Port (Inlet)
Oxidizing Chamber
Gas Release
Recirculation Pump
Water
Air Stone
Air Pump
CC
-red
ucin
g bi
orea
ctor
Redox probe
57
The overall dimensions of the silicon tubing used were length (700 mm), external
diameter (10 mm), internal diameter (8 mm) and wall thickness (2 mm). In addition,
following oxygen permeation, air sparging was used to supply maximal oxygen to re-
oxidise reduced cyanocobalamin in an oxidizing chamber. Dissolved oxygen and redox
potential (EAg/AgCl) were continually monitored in the recirculating liquid using a
dissolved oxygen probe and a redox probe respectively (Fig. 3.1). The hydraulic retention
time (HRT) was 3 minutes.
The EAg/AgCl measurements were indicative of how reduced or oxidized the recirculating
liquid was at both the entry and exit of the bioreactor. Dissolved oxygen readings
indicated if any oxygen penetrated the system. Dissolved oxygen supply is
counterproductive because irrelevant bacterial consortia could develop. Therefore, when
dissolved oxygen was detected, the aeration was terminated until the dissolved oxygen
reading dropped to 0 mg/mL.
Using the oxygen diffusion rate into the loop, the total cyanocobalamin concentration and
the EAg/AgCl, the bacterial cyanocobalamin reduction rate and the oxidation rate of reduced
cyanocobalamin were obtained and on-line monitored.
2.2.2 Cyanocobalamin Reduction Experiment
Cyanocobalamin (Sigma catalog No. 68-19-9) (1 mM) was added to a 60 mL conical
flask with acetate (40 mM) and synthetic wastewater basal media to a final volume of 50
mL. A magnetic stirrer was used to stir the liquid volumes to facilitate mass transfer. A
58
stirrer (Thermolyne Cimanec 2, model no. SP 46020-26) set at 180 revolutions per
minute was used. Headspace was degassed using a N2:CO2 (80:20) mix. Redox probes
fitted into stoppers were used to plug conical flasks.
Wet weights of biomass were weighed using a weighing scale (Sartorius Basic, model no.
BA 4100S). Biomass wet weights from both activated sludge and cyanocobalamin-
reducing bacteria were normalised to 3.0 grams. Biomass was added to the conical flasks
after approximately 20 minutes from the start of the experiment.
Biomass extracted from the cyanocobalamin-reducing reactor after 8 weeks was labelled
cyanocobalamin-reducing bacteria. The same vessel was used at different times for
incubations with cyanocobalamin-reducing bacteria and activated sludge in the presence
of acetate (40 mM) and cyanocobalamin (1 mM). Redox potential (EAg/AgCl)
measurements were taken with the same redox probe (to reduce fluctuations) and
recorded online.
2.2.3 HCBD Dechlorination Experiment
Cyanocobalamin-reducing bacteria were incubated with acetate (40 mM),
cyanocobalamin (1 mM), HCBD (Sigma catalog No. 112-19-4) (1 mM) and synthetic
wastewater to a final volume of 60 mL in 100 mL Wheaton glass serum bottles (Sigma
catalog No. Z11, 400-6). Serum bottles were sealed with rubber stoppers (Bellco catalog
No. BEL 2048-11800) and the headspace flushed with N2:CO2 (80:20) gas. The same gas
was used to purge all solutions to remove oxygen. Activated sludge incubated with
59
acetate (40 mM) in the presence of cyanocobalamin (1 mM) and HCBD (1 mM) was used
as the control. Approximately 5.0 grams of biomass wet weight from both activated
sludge and cyanocobalamin-reducing bacteria were used. All cultures set-up were
incubated at both 37 °C and 55 °C. The headspaces were then analyzed over 50 days for
HCBD dechlorination by-products.
2.3 Sampling and Analyses
Detection methods were similar to those described in Chapter 2 (section 2.3).
2.3.1 Acetate Analysis
Acetate analysis was performed according to Cheng et al. (2008).
2.3.2 Calibration of Redox Electrodes
Ag/AgCl redox reference electrodes (ionode® intermediate junction - IJ 64) were used in
all experiments. Calibration was performed using ZoBell’s solution [3.2 mM potassium
ferrocyanide (K4Fe(CN)6·3H2O) and 2.8 mM potassium ferricyanide ((K3Fe(CN)6) in 0.1
M potassium chloride (KCl)) (ionode® Redox Electrode Manual)]. All redox potentials
(referred to as EAg/AgCl) were referenced to Ag/AgCl electrolyte (-0.199 V vs. Standard
Hydrogen Electrode (SHE)).
2.3.3 Calculations
Redox potentials were recorded online via LabView® (National Instruments) every 10
seconds. The recorded redox potentials were averaged using a running average of 10
60
values to achieve a smooth plot of EAg/AgCl vs. time. The rate of redox potential change
was then obtained from the gradient of the averaged records.
3 Results and Discussion
3.1 Oxidation of Reduced Cyanocobalamin by Dissolved Oxygen
To quantify the rate of oxidation by oxygen of reduced cyanocobalamin, cyanocobalamin
(chemically reduced using palladium and hydrogen) was aerated with a constant oxygen
input. The specific oxygen mass transfer coefficient could be obtained from the oxygen
build-up in the reactor when reduced cyanocobalamin was depleted (Fig. 3.2).
Figure 3.2 Effect of oxidation (using oxygen) on reduced cyanocobalamin monitored using EAg/AgCl (♦) and dissolved oxygen (■) readings.
By plotting the oxygen transfer rates versus the oxygen concentration, the standard or
maximal oxygen transfer rate of the reactor (for dissolved oxygen = 0) was determined to
be 5 mM/h. Figure 3.2 shows that in spite of this oxygen supply rate, dissolved oxygen
could not be measured, but was instead used by the reduced cyanocobalamin resulting in
E Ag/
AgC
l (V)
Diss
olve
d O
xyge
n (C
L) (m
g/L)
10
10
6
8
2
4
-2
0
0
200100
-100-200
-500-400-300
6 25Time (min)
44
E Ag/
AgC
l (V)
Diss
olve
d O
xyge
n (C
L) (m
g/L)
10
10
6
8
2
4
-2
0
0
200100
-100-200
-500-400-300
6 25Time (min)
44
61
a gradual increase of redox potential from about -500 to -100 mV. At this final redox
potential, oxygen was no further used for cyanocobalamin oxidation, resulting in a
characteristic increase of oxygen eventually reaching saturation level at 8 mg/L. This
result indicated that reduced cyanocobalamin could be readily regenerated by aeration.
3.2 Enrichment of Cyanocobalamin-Reducing Bacteria from Activated
Sludge
Cobalt in cyanocobalamin exists in 3 different oxidation states, 3+, 2+ and +. EAg/AgCl
measurements can be used as an indicator of the oxidation state of cobalt in
cyanocobalamin. Using the Nernst equation (Equation 3.1), the EAg/AgCl was converted
into a ratio of oxidised (Co3+) and reduced cobalt (Co2+ and Co+) present in the system at
any given EAg/AgCl (Fig. 3.3).
0
0.2
0.4
0.6
0.8
1
1.2
-1 -0.8 -0.6 -0.4 -0.2 0 0.2
EAg/AgCl (V)
Cob
alt c
once
ntra
tion
(mM
).
Figure 3.3 The proportions of Co3+ (♦), Co2+ (▲) and Co+ (■) at different EAg/AgCl (V). Cobalt conversions were calculated using the Nernst Equation.
62
From knowing the total concentration of cyanocobalamin, these ratios allowed to
calculate the concentration (mM) of each of the above species of cyanocobalamin.
E = E0 - RT/nF ln (Red/Oxd) Equation (3.1)
where,
E0 (SHE) of cyanocobalamin (Co3+/Co2+) = 200 mV
E0 (SHE) of cyanocobalamin (Co2+/Co+) = -600 mV
(Guerrero-Barajas and Field, 2006; Lexa and Saveant, 1983)
E0 (EAg/AgCl) of cyanocobalamin (Co3+/Co2+) = 0 mV
E0 (EAg/AgCl) of cyanocobalamin (Co2+/Co+) = -800 mV
n = number of electrons
RT/F = 0.0615 (at 37 °C)
R = 8.31451 J/(mol.K)
T = 2713.16 + °C (K)
F = 96485.3 (C/mol)
In order to enrich cyanocobalamin reducing bacteria, an environment needs to be created
that provides a constant supply of suitable electron donor and oxidized cyanocobalamin
as the sole electron acceptor. To establish whether a mixed culture of activated sludge
bacteria could link acetate oxidation with the reduction of cyanocobalamin, the effect of
acetate addition on the reduction of oxidized cyanocobalamin was tested.
63
The addition of a small spike of acetate (100 µM) to the starved activated sludge culture
enabled bacterial reduction of oxygen and of cyanocobalamin demonstrating that the
cyanocobalamin reduction was acetate-dependent (Fig. 3.4).
-0.150
-0.100
-0.050
0.000
0.050
0.100
0.150
0.200
0.250
0.300
0 50 100 150 200 250
Time (min)
Eh (V
)(SH
E)
Figure 3.4 EAg/AgCl response obtained at the bioreactor outlet after addition of acetate (100 µM) at Time 100 min.
Approximately 800 µM of cyanocobalamin was reduced within 5 minutes (0.16
mM/min). In this time, the bacterial reductive processes were faster than the oxidative
processes. The reduced cyanocobalamin was re-oxidised after 30 minutes by oxygen
entry through the silicon tubing and using aeration. This indicated that now the oxidative
processes caused by oxygen entry into the system, were faster than the reductive
processes (all the while dissolved oxygen was controlled at set-point zero). This result
also demonstrated that the reactor described allows both bacterial reduction and the
oxygen driven oxidation of reduced cyanocobalamin. The reactor could be suitable for
EAg/AgCl (V)
Acetate (100 µM) added
64
enriching specific cyanocobalamin-reducing bacteria as long as oxygen concentration is
controlled such that it is used exclusively for the oxidation of reduced cyanocobalamin
and remains unavailable to the bacteria.
3.3 Steady State of Reduced Cyanocobalamin During Oxygen
Controlled Acetate Oxidation
By making use of the known standard oxygen transfer rate via the silicon tubing and
aeration prior to the entry into the reactor, a steady state experiment was carried out. The
acetate served the bacteria as the electron donor for cyanocobalamin reduction while the
oxygen entry allowed cyanocobalamin re-oxidation. In this experiment, the oxygen
supply was controlled at a set-point of zero preventing oxygen from becoming available
to the bacteria (Fig. 3.1). A mass balance was constructed using the following equations
(Equations 3.2 and 2.3).
CH3COOH + 4H2O + Co3+ → 8Co2+ + 2HCO3- + 10H+ Equation (3.2)
8Co2+ + 2O2 + 8H+ → 8Co3+ + 4H2O Equation (3.3)
Seven mM of acetate was found to be oxidised over approximately 7 hours (Fig. 3.5).
Since the standard or maximal oxygen transfer rate of the reactor (for dissolved oxygen =
0) was determined to be 5 mM/h, over the duration of acetate oxidation, the oxygen
transferred equates to 35 mM. However, since the oxygenation was supplied for
approximately 50 % of the duration of the experiment (to maintain the dissolved oxygen
65
at set-point zero), approximately 17 mM of oxygen was used for acetate oxidation. The
mass balance roughly reflects the theoretical ratio of 1:2 oxygen to acetate reacted
(Equations 3.2 and 3.3).
0
1
2
3
4
5
6
7
8
0 100 200 300 400 500 600
Time (min)
Con
cent
ratio
n of
Ace
tate
(mM
)
Figure 3.5 Acetate (7 mM) degraded in the cyanocobalamin reducing bioreactor.
The concentration of cyanocobalamin reduced was calculated using the difference (0.1
mM) between the average cyanocobalamin concentration at the inlet and outlet of the
reactor (Fig. 6). This was calculated from the redox potential measurements and Nernst
equation (Equation 1). By considering the constant liquid flow through the reactor (7.2
L/h) the reaction time (7 h) the average cyanocobalamin conversion was calculated to be
60 mM resulting in a cyanocobalamin/oxygen ratio of about 3.5 which is somewhat lower
than the ratio expected from equation 2, but indicates that cyanocobalamin was the key
electron acceptor to the bacteria.
66
Figure 3.6 Co2+ accumulated through cyanocobalamin reduction in the outlet (●) and inlet (■) of the reactor. Acetate (7 mM) was added at Time 0.
Figure 3.5 also shows that acetate was used in approximately 400 minutes while the time
taken for the reduction of cyanocobalamin in the reactor continued to 650 minutes (Fig.
3.6). Storage using poly-ß-hydroxybutyrate (PHB) was suspected to be the cause of this
observation. PHBs, intracellular storage material found to accumulate when nutrients
decrease, are used as an internal reserve of energy when starved of nutrients (Page and
Knosp, 1991; Singleton, 2004). Activated sludge bacteria fed with acetate have also been
known to store PHBs within a few hours (Pandolfi et al., 2007). PHBs synthesized within
a few hours (as acetate depleted at 400 min) could also explain the source of energy
necessary for continued cyanocobalamin reduction even after acetate was degraded (from
400 min to 650 min) (Fig. 3.6).
0
0.2
0.4
0.6
0.8
1
1.2
0 200 400 600 800 1000
Time (min)
Co2+
con
cent
ratio
n (m
M)
67
3.4 Effect of Cyanocobalamin-Reducing Bacteria on Cyanocobalamin
Reduction
From previous tests, it was deduced that the rate of cyanocobalamin reduction may be the
rate limiting step in HCBD dechlorination (Chapter 2). It was for this reason that
cyanocobalamin-reducing bacteria were built from activated sludge bacteria over 8 weeks
with the intention of increasing cyanocobalamin reduction rates. After 8 weeks of
enrichment, it was tested whether cyanocobalamin reduction was faster compared to the
original activated sludge.
Results show that compared to the unadapted activated sludge culture the
cyanocobalamin-reducing enrichment showed a shorter lag time for cyanocobalamin
reduction but no significant difference in the maximum rate (Fig. 3.7). This was an
unexpected result as maximum rates of cyanocobalamin reduction using cyanocobalamin-
reducing bacteria were expected to exceed that of activated sludge bacteria. On the other
hand, the result suggests that cyanocobalamin reduction is a generic feature of aerobic
mixed bacterial consortia such as in activated sludge.
68
0
0.2
0.4
0.6
0.8
1
1.2
0 20 40 60 80 100Time (min)
Co2+
conc
entr
atio
n (m
M)
Figure 3.7 Effect of cyanocobalamin-reducing bacteria (■) and activated sludge bacteria (●) on Co2+ accumulated. Biomass were added at 18 minutes. Biomass concentration was 5 g wet weight in both cases.
The reason for the extended lag phase observed in the activated sludge is unclear.
However, it could involve bacterial adaptation to the new media through the induction of
catalytically active enzymes. Quorum sensing may also be involved by which bacteria
communicate and coordinate behavior via signaling molecules (McFall-Ngai, 1999;
Singh et al., 2000). In the developed biofilm (cyanocobalamin-reducing enrichment),
where a higher concentration of specific cells or similar bacteria exists, quorum sensing
may be easily facilitated in comparison to activated sludge where different and numerous
bacteria exist.
Biomass added
69
3.5 Effect of Cyanocobalamin-Reducing Bacteria on HCBD
Dechlorination
This experiment tested if the shorter lag phase observed in cyanocobalamin reduction
using cyanocobalamin-reducing bacteria results in faster HCBD dechlorination compared
to activated sludge. Microbial cyanocobalamin reduction rates and the subsequent effect
on dechlorination of HCBD from augmented bacteria have not been reported elsewhere
thus far.
Cyanocobalamin-reducing bacteria dechlorinated HCBD approximately 2 - 4 - fold
faster than activated sludge bacteria within the first 5 days (Fig. 3.8) at 55 °C. When the
same comparison was carried out at 37 °C, this trend of cyanocobalamin-reducing
bacteria being more active in HCBD dechlorination could not be confirmed.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
70
0
10
20
30
40
50
60
70
80
0 10 20 30 40 50 60
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 3.8 Effect of cyanocobalamin-reducing bacteria (▲) and activated sludge (■) on HCBD dechlorination (as measured by C4 gases production). Biomass concentration was 5 g wet weight in both cases.
It is believed that enrichment of cyanocobalamin-reducing bacteria over a longer period
of time may enable higher dechlorination rates. Tests using activated sludge incubated on
agar plates with cyanocobalamin, as the sole electron acceptor, in an anaerobic chamber
to isolate cyanocobalamin-reducing bacteria were attempted but did not succeed to obtain
pure cultures of cyanocobalamin-reducing bacteria.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
71
5 Conclusion
Overall, the results support findings reported in Chapter 2 that the enrichment of specific
cyanocobalamin mediated dechlorination cultures is not likely to significantly improved
reductive dechlorination when compared with original activated sludge cultures.
In addition, in this chapter, it has been shown that
reduced cyanocobalamin could be readily re-oxidised by aeration using oxygen,
enabling in principle the enrichment of specific cyanocobalamin-reducing
bacteria.
re-oxidised cyanocobalamin can be subsequently supplied to activated sludge
bacteria for continued cyanocobalamin reduction as a means of enriching
cyanocobalamin-reducing bacteria (as long as dissolved oxygen did not enter the
reactor).
acetate, cyanocobalamin and oxygen concentrations match theoretical ratios.
cyanocobalamin-reducing bacteria reduced cyanocobalamin at similar rates but
with a shorter lag phase, compared to activated sludge bacteria. However,
cyanocobalamin-reducing bacteria were able to reduce HCBD to up to 4 - fold
higher rates compared to activated sludge bacteria.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
72
Chapter 4
Cyanocobalamin Enables Thermophilic Bacteria and
Methanogens from Anaerobic Digested Effluent to
Dechlorinate
Hexachloro-1,3-butadiene to Non-Chlorinated Gases
73
1 Introduction
In chapter 2, the reductive dechlorination of HCBD by activated sludge bacteria
incubated in the presence of cyanocobalamin was demonstrated. The reaction was
entirely dependent on cyanocobalamin, which was proposed to act as an electron shuttle
(mediator, carrier) between the bacteria and HCBD. HCBD dechlorination rates were also
shown to be 3 - 4 - fold higher when activated sludge bacteria were incubated under
thermophilic conditions (55 °C). Higher rates of dechlorination would be beneficial for
the implementation of bioremediation in sites with a high level of HCBD contamination.
The use of thermophilic conditions for the biological dechlorination of chlorinated
hydrocarbons has previously been reported (Ahring et al., 1992; Allard et al., 1992;
Benabdallah et al., 2007; Kengen et al., 1999; Larsen et al., 1991; Maloney et al., 1997;
Truex et al., 2007). Thermophilic (60 °C to 65 °C) anaerobic dechlorination of PCE
(Kengen et al., 1999) and polychlorinated biphenyls (PCB) (Benabdallah et al. 2007; Wu
et al., 1996) using an anaerobic enrichment culture have also been reported. Maloney et
al. (1997) observed thermophilic anaerobic biodegradation of chlorobenzoates at 75 °C.
Benabdallah et al. (2007) observed that the total PCB removal efficiency was in the range
of 59.4 - 83.5 % under thermophilic conditions and 33.0 - 58.0 % under mesophilic
condition. In addition, they found that adsorbed organic halogen compounds (AOX)
removal efficiency was approximately 40.4 - 50.3 % for thermophilic conditions and 30.2
- 43.2 % for mesophilic conditions.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
74
1.1 Methanogens and Dechlorination
Dechlorination has been identified under both methanogenic conditions and by pure
methanogenic strains. The reductive dechlorination of chlorinated aliphatic and aromatic
compounds has been observed in a wide variety of methanogenic environments. Booker
and Pavlosthasis (2000) noted the partial dechlorination of HCBD in methanogenic
enrichments of contaminated estuarine sediments. Jones et al. (2006) identified
acetoclastic methanogens from the class Methanomicrobia, in mixed microbial consortia,
to be involved in dechlorination. The contaminants 1,1,2,2-tetrachloroethane (TeCA),
trichloroethene (TCE), cis and trans 1,2-dichloroethene (DCE), 1,1,2-trichloroethane
(1,1,2-TCA), 1,2-dichloroethane, and vinyl chloride (VC) were dechlorinated to the non-
chlorinated hydrocarbons ethane and ethane (Jones et al., 2006).
The abundance and scarcity of methanogens was reported to either enhance or inhibit
dechlorination activity respectively. Hashsham and Freedman (1999) observed the
enhancement of TCE dechlorination by methanogenic cultures. Moreover, in highly
enriched Dehalococcoides cultures containing the methanogenic bacteria,
Methanosarcina sp., 7 - fold higher rates of VC dechlorination was obtained in
comparison to 2-bromo-ehane sulfonate (BES) amended cultures (Heimann et al., 2006).
In addition, Kim and Rhee (1998) observed that the inhibition of methanogens, using
BES, reduced the rate and extent of dechlorination of Aroclor® 1248 without affecting the
size of dechlorinating populations. Furthermore, methane production was noted along
with PCB dechlorination in sediment slurries (Morris et al., 1992; Nies et al., 1990; Ye et
al., 1992).
75
It was unclear if methanogens were directly involved in dechlorination or if they aided
dechlorination via the maintenance of electron flow. Nonetheless, it is apparent that
methanogens or methanogen-associated bacteria are involved in dechlorination.
Several mechanisms involved in dechlorination by methanogens have been reported.
Methanogens possess the methyl reductase enzyme complex, which catalyses the final
step in methane formation (DiMarco et al., 1990; Sparling and Daniels, 1987). This
enzyme contains a unique cofactor, coenzyme M (CoM) (2-mercaptoethanesulfonate),
found only in methanogens (Löffler et al., 1997). Holliger et al. (1992b) reported that the
CoM plays an important role in reductive dechlorination.
One other possible mechanism is the participation of transition metal cofactors (e.g. F430)
in dechlorination. Anaerobes rich in these cofactors, particularly methanogens, may
catalyse reductive dechlorination in anaerobic environments (Tandoi et al., 1994). This
reductive dechlorination occurs as a form of co-metabolism in which reduced forms of
the cofactors catalyzed the reductions (Zinder and Gossett, 1995).
Aceticlastic methanogens of the genus Methanosarcina may greatly increase rates of
chloroethene degradation by the interspecies transfer of H2 to dehalorespiring microbes
(Heimann et al., 2006). The dechlorination of chlorinated hydrocarbons by aceticlastic
methanogens is important because these bacteria are common in nature. This
transformation may also be vital in treating contaminated soils and waters (Fathepure and
Boyd, 1988).
76
With the goal of further increasing microbial reductive dechlorination of HCBD, the
current study uses thermophilic bacteria from anaerobic digested effluent (ADE) from
municipal waste and methanogens cultured from ADE as the catalyst for microbial
reductive dechlorination, again in the presence of cyanocobalamin as the electron shuttle
under thermophilic conditions.
2 Experimental Procedures
2.1 Medium Composition
2.1.1 Anaerobic Digested Effluent
The thermophilic anaerobic inoculum and liquid was obtained from a laboratory scale
anaerobic digestion process that combined composting and thermophilic anaerobic
digestion in a single closed vessel under batch conditions (Walker et al., 2006). The
anaerobic samples (ADE) (2 g biomass/L) taken from the process were stored in Schott®
bottles with the headspace flushed with N2:CO2 (80:20) mix. The pH was 7.8, COD 30 g
/L, ammonia 80 mM, acetate 10 mM, propionate 20 mM and butyrate 5 mM. The
predominant methanogens in the anaerobic liquid have been identified using Terminal
Restriction Fragment Length Polymorphism (T-RFLP) as Methanoculleus species and
Methanosarcina thermophila. Other archaeal groups were also found, but at lower levels.
The eubacterial communities in the laboratory scale reactor were dominated by
Prevotella sp. and Thermodesulfobacterium (up to 16 % of total eubacterial
communities).
77
2.1.2 Digested Piggery Waste
The anaerobic digestion process utilized thermophilic anaerobic digestion of mid organic
strength piggery waste in a closed reactor under batch conditions with a hydraulic
retention time of 2 days. Digested piggery waste (DPW) (1 g biomass/L) was obtained
from South Australian Research and Development Institute (SARDI). DPW was
anaerobic liquid drained from the reactor after the second day and was stored in a Schott®
bottles with the headspaces flushed with N2:CO2 (80:20) mix.
2.1.3 Sheep Rumen Content
Sheep rumen content (SRC) (5 g biomass/L) was obtained from Commonwealth
Scientific and Research Organization (CSIRO) based in Floreat, Western Australia. SRC
was extracted from a live sheep via a rubber stopper attached to its rumen.
2.1.4 Cultivation of Methanogens
Two strains of the hydrogenotrophic methanogens Methanoculleus thermophilus and
Methanobacterium Thermoautothrophicum used were isolated from ADE. Both strains of
methanogens were grown as pure cultures in 30 mL reduced basal medium in a 100 mL
serum bottles sealed with rubber stoppers. The headspace of the serum bottle with
Methanoculleus thermophilus was flushed with H2 which was provided as an electron
source. In addition, an overpressure of CO2 was supplied to make up a final headspace
mix of H2:CO2 (80:20). Formate (80 mM) (Sigma catalog no. 141-53-7) was supplied as
the electron donor for Methanobacterium Thermoautothrophicum.
78
An experiment was set up as shown in Table 4.1 to test if the pure methanogens could
dechlorinate HCBD and to test the effect of the anaerobic liquid on dechlorination rates.
Table 4.1 Dechlorination experiments with pure strains of Methanothermoculeus and Methanothermobacter sp.
Trial Sterile reduced basal medium
(mL)
Sterile anaerobic liquor (mL)
Pure methanogenic culture (mL)
Control 1 5.5 - -
Control 2 4.5 1.0 -
Test 1 1.5 - 4
Test 2 0.5 1.0 4
Both the pure cultures, Methanoculleus thermophilus and Methanobacterium
Thermoautothrophicum, were pre-grown until turbid and transferred into sterile 16 mL
Hungate fitted with butyl rubber septa. Cyanocobalamin was supplied at a concentration
of 0.1 mM. The headspaces of all tubes were flushed with H2 as an electron source with
an overpressure of 3 mL CO2 to make up a final headspace mix of H2:CO2 (80:20) mix.
After flushing, 1 mM HCBD was added. Incubation conditions were similar to those
described in this chapter in section 2.2.1. The headspace of each tube was analysed every
24 hours.
2.1.5 Basal Medium
The composition of the basal medium used was based on DSM 334 medium (DSMZ,
1983) (Chapter 2, section 2.1). The medium was dispensed into screw cap 16 mL
Hungate tubes fitted with butyl rubber septa. Headspaces were then flushed with N2:CO2
(80:20) mix and autoclaved at 126 C for 20 min. Prior to inoculation, the medium was
79
reduced with sterile stock solutions of Na2S and cysteine-HCl to a final concentration of
0.3 g/L.
2.1.5 Sterile Anaerobic Liquor
Sterile anaerobic liquor (SAL) was prepared by centrifuging ADE in a Sorvall®
ultracentrifuge for 5 minutes at 10,000 g to pellet all biomass and particulate matter. The
supernatant was sterilised by filtering through a Scheicher & Schuell® filter (0.2 m)
fitted to a needle passing through the septum into a sterile 16 mL Hungate tube
previously flushed with N2:CO2 (80:20) mix.
HCBD stock (10 % (vol./vol.)) solution was prepared by dissolving neat HCBD (97 %) in
methanol (100 %) for use in experiments with methanogenic cultures.
2.2 Dechlorination Experiments
HCBD dechlorination experiments were similar to those described in Chapter 2 (section
2.2). The following mediators 2-anthraquinone disulfonic acid (AQDS) (Aldrich catalog
no. 131-08-8), Cobalt (II) chloride 6-hydrate (Aldrich catalog no.7791-13-1), Cysteine
(Aldrich catalog no.52-90-4), Humic Acids (Aldrich catalog no. 68131-04-4), Indigo
Carmine (Aldrich catalog no. 860-22-0, Jacobsen’s Catalyst (Aldrich catalog no. 47-460-
6), Methylene Blue (Aldrich catalog no. 7220-79-3), Neutral Red (Aldrich catalog no.
553-24-2), Quinhydrone (Aldrich catalog no.106-34-3), Resazurin (Aldrich catalog no.
62758-13-8), and Safranin O (Aldrich catalog no. 477-73-6), were added at both 0.1 and
1 mM in separate test vials.
80
2.3 Sampling and Analyses
Detection methods were similar to those described in Chapter 2 (section 2.3).
3 Results and Discussion
3.1 Effect of Thermophilic Bacteria on HCBD Dechlorination
It was shown that HCBD dechlorination rates* were 4 - fold higher when activated sludge
bacteria were incubated under thermophilic rather than mesophilic conditions in the
presence of cyanocobalamin (Chapter 2; James et al., 2008). In this experiment, the effect
of thermophilic bacteria from ADE on HCBD dechlorination in the presence of
cyanocobalamin was tested.
Thermophilic bacteria from ADE were able to dechlorinate HCBD completely to non-
chlorinated endproducts, as evidenced by monitoring the completely dechlorinated C4
gases using GC-MS in the presence of cyanocobalamin (Fig. 4.1). No dechlorination was
observed in the absence of cyanocobalamin. These findings are in line with findings
described previously (James et al., 2008). It was immediately clear that using
thermophilic bacteria allowed more effective dechlorination than previously reported.
The thermophilic bacteria produced about 120 µmoles of C4 gases per litre culture
within one day. From the measurements, it can be estimated that the minimum rate of
dechlorination was faster than 120 µmoles/L culture/day (Fig. 4.1). This rate was about Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
81
30 - fold higher than previously reported using activated sludge bacteria (Chapter 2;
James et al., 2008) under otherwise identical conditions.
0
40
80
120
160
0 0.5 1 1.5 2 2.5
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.1 Dechlorination by thermophilic bacteria from ADE in the presence (●) and absence (■) of cyanocobalamin (0.1 mM) as measured by the concentration of C4 gases from HCBD dechlorination.
Compared to activated sludge cultures which typically had a lag phase of 7 days, the
thermophilic ADE cultures dechlorinated within the first day. This suggests that not
specific microflora needed to develop but that the bacteria present in the anaerobic
thermophilic consortium were able to dechlorinate instantly, presumable via non-specific
reactions involving the reduction of cyanocobalamin.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
82
Similar to activated sludge cultures, dechlorination did not continue but stalled after
about 30 to 120 µM of completely dechlorinated products/ L culture were formed. As
the cultures in the current study dechlorinated at higher rates, this stalling was reached
within a few days.
3.2 Effect of Cyanocobalamin Concentration on HCBD Dechlorination
There is a high cost associated with the use of cyanocobalamin. For large-scale
bioremediation purpose, the high cost may rule out the use of cyanocobalamin. In the
interest of cost savings, cyanocobalamin concentrations lower than 1 mM were tested for
their effect on C4 gases production from HCBD dechlorination.
The rate of HCBD dechlorination* was not significantly slowed by lowering the
cyanocobalamin concentration from 1 mM down to 0.05 mM (Fig. 4.2). The minimum
concentration of cyanocobalamin required here, were 8 - fold lower than with activated
sludge cultures previously described (James et al., 2008). This represents an 8 - fold cost
saving per application. In all subsequent experiments, 0.1 mM cyanocobalamin
concentration was used, to ensure that cyanocobalamin availability was not limiting the
reaction.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
83
0
40
80
120
160
200
0 0.5 1 1.5 2 2.5
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.2 Effect of cyanocobalamin concentration on the concentration of C4 gases from HCBD dechlorination. Cyanocobalamin concentrations used were 0.01 (▲), 0.05 (●), 0.1 (■) and 1 (∆) mM.
3.3 Effect of Ethanol Concentration on HCBD Dechlorination
Ethanol was used with activated sludge cultures to increase the solubility of HCBD, a key
factor in its bioavailability and subsequent biological dechlorination. The effect of
ethanol on thermophilic bacteria was not clear. Ethanol concentrations of 0.5 % to 2 %
(vol./vol.) were tested for their effect on C4 gases production from HCBD dechlorination.
In contrast to activated sludge cultures, ethanol did not stimulate HCBD dechlorination in
ADE cultures. In fact, the highest concentration of C4 gases* produced was in the culture
without ethanol (Fig. 4.3).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
84
0
50
100
150
200
250
300
0 1 2 3 4Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.3 Effect of ethanol concentration on the concentration C4 gases from HCBD dechlorination. Ethanol concentrations used were 0 (●), 0.05 (■) and 2 (▲) % (vol./vol.).
3.4 Effect of ADE Biomass Age on HCBD Dechlorination
Thus far, fresh cultures of ADE were used as inocula for dechlorination reactions. If fresh
ADE cultures were indeed necessary for effective dechlorination, it would be difficult in
practice to produce suitable cultures when required. In order to test if fresh and active
anaerobic populations were required for effective dechlorination, ADE biomass of
different ages were incubated with HCBD.
ADE biomass age did not influence HCBD dechlorination. Dechlorination rates and
concentrations of C4 gases* produced were similar in cultures of all ages of biomass (Fig. Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
85
4.4). Batches of anaerobic liquor could thus be prepared and stored for future use without
compromising reaction rates. The extent to which the respective microbes could be
concentrated into slurry or freeze dried was not tested in this study.
0
20
40
60
80
100
120
140
0 0.2 0.4 0.6 0.8 1
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.4 Effect of ADE biomass age on the concentration of C4 gases* from HCBD dechlorination. Biomass ages used were 1 day (▲), 2 months (■) and 1.5 years (●).
Because of more frequent sampling during the first day of incubation this experiment also
revealed that after a few hours of lag phase, the dechlorination occurred mostly as a boost
over about 5 to 8 hours and then quickly slowed down when about 100 µM was
dechlorinated (Fig. 4.4).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
86
3.5 Effect of ADE Solid Fraction on HCBD Dechlorination
Since the anaerobic digestion process, involved in the production of ADE utilized the
organic fraction of municipal solid waste, the majority of microbes involved in anaerobic
digestion of the solids were expected to be adsorbed to the solid waste material. This
experiment tested if the presence of these solids in dechlorination reactions stimulated
HCBD dechlorination. ADE solid compost (5 grams) was added to 1 serum bottle while
deoxygenated de-ionised water was added to the treatment without solids to equal levels.
The addition of solids from the ADE process did not stimulate HCBD dechlorination
(Fig. 4.5). It is possible that HCBD was adsorbed onto the solid surface and not
accessible to the microbes for dechlorination i.e., decreased bioavailability. This may also
explain why soil contaminated with HCBD is more difficult to be biologically
dechlorinated (Fig. 4.8).
87
0
20
40
60
80
100
120
140
0 1 2 3 4
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.5 Dechlorination by thermophilic bacteria from ADE in the presence (▲) and absence (■) of solids as measured by the concentration of C4 gases from HCBD dechlorination.
3.6 Effect of Thermophilic Bacteria from Other Anaerobic Digestion
Processes on HCBD Dechlorination
3.6.1 Effect of Digested Piggery Waste on HCBD Dechlorination
So far, faster dechlorination has been obtained using ADE cultures than activated sludge
cultures. If other thermophilic bacterial consortia could dechlorinate just as well as ADE
cultures, it would enable the use of those other cultures should ADE digestion become
unavailable. A test using different thermophilic bacterial consortia would also answer if
the fast dechlorination was specific to ADE or if it was a more general feature of
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
88
thermophilic bacterial consortia. In this experiment, the effluent of a thermophilic reactor
treating piggery waste (DPW) was tested for HCBD dechlorination in both the presence
and absence of cyanocobalamin.
Thermophilic bacteria from DPW dechlorinated HCBD in the presence of
cyanocobalamin (Fig. 4.6). However, the rate of C4 gases produced was approximately 5
- fold less (per gram biomass) than that observed using ADE. Again, as in all other
successfully dechlorinating cultures, the presence of cyanocobalamin was essential.
0
10
20
30
40
0 2 4 6 8 10
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.6 Dechlorination by thermophilic bacteria from DPW in the presence (■) and absence (▲) of cyanocobalamin (0.1 mM) as measured by the concentration of C4 gases* from HCBD dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
89
3.6.2 Effect of Sheep Rumen Content on HCBD Dechlorination
In all previous cultures tested, there was a dependence on cyanocobalamin to enable
HCBD dechlorination. Rather than adding cyanocobalamin as an ingredient external to
the cells, it seemed possible that those bacteria that are known to produce
cyanocobalamin, such as rumen microbial consortia (Bigger et al., 1976; Caldwell et al.,
1973) may reduce the dependence for the external addition of cyanocobalamin. Thus, the
inclusion of bacteria able to produce cyanocobalamin themselves may be able to replace
external cyanocobalamin addition.
If successful, sheep rumen content (SRC) would offer an inexpensive source of biomass
for a possible dechlorination reactor. The added advantage of the presence of shuttles in
the rumen may also serve to enhance HBCD dechlorination rates. In this experiment, the
effect of sheep rumen content on HCBD dechlorination both in the presence and absence
of cyanocobalamin were tested.
At 55 °C bacteria from SRC were capable of HCBD dechlorination, but only in the
presence of cyanocobalamin at 55 °C. However, dechlorination by SRC bacteria was
about 16 times slower (per gram biomass) than observed for ADE (Fig. 4.7). No HCBD
dechlorination was observed when sheep rumen was incubated at 37 °C in both the
presence and absence of cyanocobalamin (data not shown).
90
0
10
20
30
40
50
60
70
0 5 10 15 20 25 30
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cul
ture
)
Figure 4.7 Dechlorination by bacteria from sheep rumen content in the presence (▲) and absence (■) of cyanocobalamin (0.1 mM) as measured by the concentration of C4 gases from HCBD dechlorination.
The fact that no HCBD dechlorination was observed in the absence of cyanocobalamin
showed either that insufficient cyanocobalamin was produced by microorganisms in SRC
or that the cyanocobalamin produced by microorganisms in the SRC could not act as a
proper shuttle as it would be present inside the cell for pathways such as the methyl-
malonyl pathway of propionate fermentation.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
91
The tests carried out above show that HCBD dechlorination was not specific to ADE
cultures. Thermophilic bacterial consortia from DPW and SRC could also dechlorinate
HCBD. Like ADE cultures, it was observed that both the DPW and SRC cultures
required cyanocobalamin and stalled after between 35 and 60 µM of dechlorinated
endproducts were formed, which took 7 and 18 days respectively.
3.7 Dechlorination of Soil Contaminated with HCBD
While the thermophilic conditions necessary for the reductive dechlorination described in
this study are not suited for in-situ treatment of soils, it is perceivable that an ex-situ
treatment or a “pump and treat” system could be used for the bioremediation of
contaminated sites. In this experiment, the potential of the ADE cultures to dechlorinate
soil was evaluated.
HCBD contaminated soil (10 g) was added to a serum bottle and incubated with 30 mL of
ADE culture in the presence of cyanocobalamin. Evidence of HCBD dechlorination
inside the soil was found (Fig. 4.8). The maximum rate obtained was about 16 µmoles/L
culture/day* which was about 10 - fold lower than in cultures to which pure HCBD was
added. Approximately 2 % of the total HCDB in the soil was dechlorinated to C4 gases.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
92
0
10
20
30
40
50
60
0 10 20 30 40Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.8 Dechlorination from HCBD contaminated soil by thermophilic bacteria from ADE in the presence of cyanocobalamin (0.1 mM) as measured by the concentration of C4 gases from HCBD dechlorination. Headspace was flushed on Day 9, 20 and 25.
HCBD sequestration into the nanopores of sediments could be an explanation for the
slower rates observed with contaminated soil cultures. Guerin and Boyd (1992) observed
that the fate of hydrophobic contaminants in soils was complicated by the competing
processes of contaminant sorption and biodegradation. Abramowicz et al. (1993) also
noted that PCB dechlorination rates were affected by PCB contaminated sediment
concentrations.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
93
Regardless of the slow rate of dechlorination, this is the first study to demonstrate that
HCBD in contaminated soil could be dechlorinated to chlorine-free C4 gases. In addition,
it was observed that headspace removal re-initiated C4 gases production (Fig. 4.8). This
result suggests that C4 gases could be built up to inhibitory levels. Removal of inhibitors
would be necessary to sustain dechlorination. In column experiments, where
dechlorination products are continuously removed, production inhibition would
potentially be reduced.
3.8 Effect of Diluting Accumulated Dechlorination Byproducts in the
Culture Headspace
From the previous experiment, it was determined that the build-up of C4 gases could be
one reason for the early termination of the sustained production of C4 gases. The removal
of headspace then re-initiated the production of C4 gases. In other words, the removal of
headspace content with the C4 gases at inhibitory levels and the subsequent introduction
of C4 gases-free headspace resumed dechlorination.
One way of avoiding the build-up of inhibitory levels of gaseous products is by dilution
such as by providing a larger gas space. This would allow for a greater amount of C4
gases to be produced before reaching inhibitory levels. Thus, the effect of different ADE
culture to headspace ratios on HCBD dechlorination was tested (Fig. 4.9). The dilution of
gaseous products by increasing headspace volumes enhanced HCBD dechlorination (Fig.
4.9). The culture with the smallest headspace to liquid ratio showed virtually no
94
dechlorination, supporting the idea that gaseous endproducts were responsible for the
stalling of the reaction.
0
100
200
300
400
500
0 1 2 3 4 5 6Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.9 Effect of different ADE culture to headspace ratios on the concentration of C4 gases* from HCBD dechlorination. Culture to headspace ratios used were 4:1, 1:2, 1:4, 1:9 and 1:20.
By diluting the gaseous products with increasing headspace volumes, one would expect
twice as much dechlorination for each doubling in headspace volume. However, the total
amount of C4 gases produced* was not proportional to the size of the headspace used.
Considering that 1 mM HCBD was added, the dechlorination was almost 50 % by day 5*
in the culture with the largest headspace. At that level of dechlorination*, it may be
possible that HCBD depletion may play a role in slowing HCBD dechlorination rates.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
1:20
1:4
1:2
4:1
1:9
95
3.9 Effect of Various Mediators on HCBD Dechlorination
Due to the high cost of cyanocobalamin when applied to large-scale operations, a number
of low-cost electron mediators were tested for their effect on HCBD dechlorination. No
dechlorination was observed with any of the mediators listed in Table 4.2.
Cyanocobalamin is still the most effective mediator for HCBD dechlorination.
Table 4.2 List of various mediators tested for C4 gases concentration from HCBD dechlorination.
2-anthraquinone disulfonic acid (AQDS) Cobalt (II) chloride 6-hydrate Cysteine Humic Acids Indigo Carmine Jacobsen’s Catalyst Methylene Blue Neutral red Quinhydrone Resazurin Safranin O
3.10 Effect of 2-bromo-ethane sulfonate, a Methanogen Inhibitor, on
HCBD Dechlorination
The reduction of carbon dioxide (CO2) in methanogenesis behaves as an electron sink.
This electron sink diverts away electrons that could otherwise be used in dechlorination.
Hashsham and Freedman (1999) used BES to enhance dechlorination by avoiding the
electron flow via methanogenesis. BES is a specific inhibitor of methanogenesis. BES is
a competitive inhibitor of the methyl reductase enzyme that catalyses the final step in
methanogenesis in methanogens. In this experiment, the effect of BES on HCBD
dechlorination was tested.
96
In our experiments, the BES addition did not stimulate but completely inhibited HCBD
dechlorination (Fig. 4.10). BES also inhibited HCBD dechlorination by activated sludge
cultures (data not shown). This result demonstrates that either methanogens in activated
sludge played a role in HCBD dechlorination or that there was selective debromination of
BES by both activated sludge and ADE.
0
50
100
150
200
250
0 2 4 6 8
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 4.10 Dechlorination by thermophilic bacteria from ADE incubated with cyanocobalamin (0.1 mM) in the presence (■) and absence (▲) of BES (10 mM) as measured by the concentration of C4 gases from HCBD dechlorination.
The possibility/prospect of methanogenic involvement in HCBD dechlorination was
further investigated using two strains of methanogens isolated from ADE.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
97
3.11 Dechlorination of HCBD by Methanogenic Cultures Purified from
ADE
ADE contains high population of methanogens. Methanogenic species active in the ADE
digestion process include Methanoculleus sp., Methanobacterium sp. and
Methanosarcina sp.. Methanosarcina sp. is an acetoclastic methanogen and is usually the
major methanogen found in anaerobic sewerage sludge (Petersen and Ahring, 2006).
Methanobacterium sp. and Methanoculleus sp. are hydrogenotrophic methanogens.
Unlike most anaerobic waste treatment systems, Methanoculleus sp. was found to be the
predominant methanogen in the ADE digestion system and is suspended in high number
in the anaerobic liquor. Significant HCBD dechlorination rates achieved using ADE may
be attributed to the activity of the mixed methanogenic population when an appropriate
electron source is provided. In this experiment, the two hydrogenotrophic methanogens,
Methanoculleus thermophilus and Methanobacterium Thermoautothrophicum isolated
from ADE were tested for their ability to dechlorinate HCBD.
Dechlorination was not detected in both controls 1 and 2, eliminating the possibility that
the dechlorination may be due to chemical reaction between the basal media, SAL,
HCBD and cyanocobalamin. A brief comparison between the dechlorination activity of
Methanothermoculeus and Methanothermobacter sp. is shown in Table 4.3.
98
Table 4.3 Dechlorination by Methanoculleus thermophilus and Methanobacterium Thermoautothrophicum isolated from ADE.
HCBD dechlorination
Greater dechlorination rates with SAL
Onset of dechlorination
Cell viability after
dechlorination Methanoculleus
thermophilus Yes No < 24 hours Not viable
Methanobacterium
Thermoautothrophicum Yes Yes >24 hours Viable
3.11A HCBD Dechlorination by Methanoculleus thermophilus
HCBD dechlorination was observed in both Methanoculleus thermophilus and
Methanobacterium Thermoautothrophicum cultures supplemented with cyanocobalamin.
HCBD dechlorination by Methanoculleus thermophilus occurred within 24 hours.
However, no viable Methanoculleus cells were observed under the light microscope after
24 hours of dechlorination and no methane was produced.
The Methanoculleus thermophilus trial supplemented with sterile SAL did not show a
better dechlorination rate. The average HCBD dechlorination rate of Methanoculleus
thermophilus was 0.06 µmoles/day and the cumulative C4 gases production was 15.3
µmoles/L culture*. No further dechlorination activity by Methanoculleus thermophilus
was detected after 24 hours.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
99
Although HCBD dechlorination by Methanoculleus thermophilus occurred within 24
hours, the dechlorination activity could not be sustained due to the complete cell death
within 24 hours. The cause of cell death could possibly be due to the toxicity of HCBD or
other HCBD dechlorination products to Methanoculleus thermophilus.
3.11B HCBD Dechlorination by Methanobacterium
Thermoautothrophicum
A 24-hour delay was observed on the onset of HCBD dechlorination by
Methanobacterium Thermoautothrophicum. In contrast to Methanoculleus thermophilus,
cells were viable after HCBD dechlorination. Methane was produced by all
Methanobacterium Thermoautothrophicum cultures within 24 hours.
The culture supplemented with SAL achieved greater cumulative C4 gases production
after 48 hours compared to the culture without SAL (Fig. 4.11). As C4 gases production*
increased rapidly from Day 2 to 3, methane production was not apparent in the culture
supplemented with SAL (Fig. 4.12).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
100
0
5
10
15
20
25
30
35
40
45
0 1 2 3 4
Time (Days)
Cum
ulat
ive
C4
gase
s pr
oduc
tion
(µm
oles
/L c
ultu
re)
Figure 4.11 Dechlorination by Methanothermobacter sp. isolated from ADE in the presence (■) and absence (▲) of SAL as measured by the concentration of C4 gases from HCBD dechlorination.
0
0.0005
0.001
0.0015
0.002
0.0025
0.003
0.0035
0.004
0.0045
0.005
0 1 2 3 4
Time (Days)
Cum
ulat
ive
met
hane
pro
duct
ion
(L/L
cu
lture
)
Figure 4.12 Methanogenesis by Methanothermobacter sp. isolated from ADE in the presence (■) and absence (▲) of SAL as measured by the cumulative methane production.
101
Conversely, in the absence of SAL, methane production increased rapidly from Day 2 to
3 whereas C4 gases production was almost stagnant. The highest HCBD dechlorination
rate of Methanobacterium Thermoautothrophicum in the presence and absence of SAL were
0.116 µmoles/day and 0.028 µmoles/day* respectively. After Day 3, the cumulative C4
gases produced in the Methanobacterium Thermoautothrophicum culture supplemented
with SAL was 38.9 µmoles/L culture* compared to 8.3 µmoles/L* in a culture where SAL
was absent.
HCBD dechlorination rate using a culture of Methanobacterium Thermoautothrophicum
increased by approximately 4 to 5 – fold* in the presence of SAL. SAL is rich in
chemical mediators that may have catalysed HCBD dechlorination in a similar way to
cyanocobalamin.
It is also possible that the presence of certain chemical mediators in SAL enabled
Methanobacterium Thermoautothrophicum to harness more energy from HCBD
dechlorination compared to methane production. As shown in Figure 4.11 and 4.12, when
C4 gases production* increased rapidly in the presence of SAL, methane production was
almost zero. Conversely, the absence of SAL favoured methane production by
Methanobacterium Thermoautothrophicum than HCBD dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
102
4 Discussion
Although relatively high rates of HCBD dechlorination were reported in this chapter,
HCBD dechlorination was not sustainable and stalled after less than 0.2 mM of HCBD
were dechlorinated*. This trend has also been observed with other cultures using
activated sludge bacteria (James et al., 2008). The fact that ADE cultures stalled earlier,
suggests that it is not a time-dependent effect of the substrate that caused the termination
of the reaction, but more likely the accumulation of metabolites of the reaction. The
finding that headspace degassing seems to re-initiate HCBD dechlorination (Fig. 4.8) is
further proof that endproduct inhibition is a likely cause of dechlorination stalling.
We have observed that inhibition sets in once C4 gases production reached between 100 -
150 µmoles/L culture (Fig. 4.1 - 4.5). Alvarez-Cohen and McCarty (1991a) noted that in
a methanotrophic culture, chloroform (CF) dechlorination inhibited trichloroethylene
(TCE) dechlorination due to toxicity from CF dechlorination products. They also noted in
another study that toxic effect by TCE and its dechlorination products caused the decline
in the activity of the dechlorinating culture (Alvarez-Cohen and McCarty, 1991b). In yet
another study, it was found that once dechlorination stalled, the number of dechlorinating
microroganisms began to decrease (Kim and Rhee, 1997). Although in the current study,
a single contaminant was used, it is possible that the products from its dechlorination
impart toxicity to the dechlorinating culture, thereby decreasing dechlorinating
microorganisms, after the initial dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
103
The accumulation of intermediates and the lack of availability of specific microorganisms
able to dechlorinate such intermediates have been associated with stalling dechlorination
reactions. Vinyl chloride (VC), one such intermediate accumulated from the biological
dechlorination of tetrachloroethene (PCE), is responsible for stalling (at 22 µM) and it
was thought that this apparent stall was due to a lack of required microorganisms (Sung et
al., 2006). Sung et al. (2006) then observed that the ability of Dehalococcoides sp. strain
BAV1 to couple growth to the degradation of the VC subsequently lead to efficient
dechlorination without stalling.
In general, the biological dechlorination rates reported in this chapter were higher than
that reported for other contaminants (Kengen et al., 1999; Larsen et al., 1991). Larsen et
al. (1991) observed anaerobic reductive dechlorination of pentachlorophenol (PCP) using
digested manure, digested sludge and inocula from natural ecosystems as incocula under
thermophilic (50 °C) conditions. The highest rate of dechlorination observed in their
study was 7.5 µmoles/L culture/day. The rates reported in this chapter were a minimum
25 - fold higher than those reported by Larsen et al. (1991).
In the cyanocobalamin dependent dechlorination of HCBD by activated sludge bacteria, a
lag phase (5 days) was observed before noticeable dechlorination stated (James et al.,
2008). Such lag phases were also observed in the dechlorination of other contaminants
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
104
(Bryant et al., 1991; Deweerd and Bedard, 1999; Fan and Scow, 1993). No such lag
phase was apparent in ADE cultures of the present study. The observed lag phase could
be explained by a relatively long acclimation period required for HCBD dechlorination in
activated sludge cultures. In their studies, Bryant et al. (1991) and Deweerd and Bedard
(1999) noted that the adaptation of sediment communities eliminated the lag phase before
dechlorination was observed along with an increase in the rate of contaminant
conversion. Fan and Scow (1993) observed the increase the lag phase with decreasing
temperature.
In one recent study, it was observed that HCBD was completely dechlorinated at a rate of
approximately 30 µmoles/L culture/day by Serratia marcescens without the use of any
mediators (Li et al., 2008). This study demonstrates that a pure strain may indeed be able
to dechlorinate HCBD.
Headspace degassing re-initiated dechlorination. However, the subsequent headspace
degassing was insufficient to prevent the stalling of dechlorination. This result seems to
suggest that factors other than end-product toxicity could have caused the dechlorination
to stall. One factor may be the depletion of nutrients required for dechlorination. Alvarez-
Cohen and McCarty (1991b) suggested that the depletion of microbial energy stores may
have affected sustained dechlorination. DiStefano et al. (1992) observed the nutritional
dependency of dechlorinating cultures to sustain dechlorination through the addition of
filtered supernatant from a methanol-fed microbial culture into PCE dechlorinating
culture.
105
5 Conclusion
In comparison to previously published rates of HCBD dechlorination (James et al., 2008),
a number of improvements are highlighted in this chapter. They are
No requirement for ethanol as an enhancer of the reaction.
Four - fold lower levels of cyanocobalamin required.
Cyanocobalamin is the most effective mediator for HCBD dechlorination.
Approximately 30 - fold faster dechlorination rates compared to activated sludge
cultures.
ADE biomass age did not influence HCBD dechlorination.
Dechlorination in soil contaminated with HCBD was possible with ADE in the
presence of cyanocobalamin at a rate of 16 µmoles/L culture/day.
Dilution of gaseous products in the headspace enabled some continued
dechlorination.
Methanogens are able to dechlorinate HCBD in the presence of cyanocobalamin.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
106
Chapter 5
Bacterially Produced Mediators Enhance The
Dechlorination of
Hexachloro-1,3-butadiene to Non-Chlorinated Gases
&
Investigations into Why Dechlorination Stalls
107
1 Introduction
1.1 Microbially Produced Mediators
Extracelluar electron transfer can occur in the presence of microbially produced
mediators. Bacteria can use redox-active organic small molecules, generated outside or
inside the cells, as mediators to shuttle electrons between reduced and oxidized
compounds (Hernandez and Newman, 2001). Microbially produced mediators have a
wide range of applications. These include assisting in power generation in microbial fuel
cells, the degradation of textile dyes and the reduction of metals.
Pyocyanin produced by Pseudomonas aeruginosa has been shown to aid in power
generation in microbial fuel cells (Rabaey K et al., 2004; Rabaey K et al., 2005). This
organism has also been known to produce a blue pigment, phenazine that can function as
an extracellular electron shuttle used for iron reduction. Shewanella oneidensis strain
MR-1 was shown to excrete a quinone-like molecule (menaquinone) that was also used
for iron reduction (Newman and Kolter, 2000). In addition, in low iron environments,
bacteria can make their own chelators; called siderophores which are small molecules
produced and sent out by bacteria to scavenge iron to aid growth (Morel and Hering,
1993).
The molecule, pyridine-2,6-thiocarboxylate, isolated from cell-free supernatants of iron-
limited cultures of a Pseudomonas stutzeri strain, was able to shuttle electrons to carbon
tetrachloride and dechlorinate it (Lee et al., 1999).
108
Extracellular electron transfer using shuttling compounds may also generate energy for
microbial cell growth and/or maintenance. Exchanges of shuttling compounds may
syntrophically link diverse organisms in nature (Hernandez and Newman, 2001). The
excretion of quinoid compound, 2-amino-3-carboxy-1,4-naphtoquinone (ACNQ) by
Propionibacterium freundenreichiiis is one example. The compound ACNQ has been
shown to stimulate growth by a shuttling mechanism in a beneficial population of
bacteria in the human gastrointestinal tract (Yamazaki et al., 1999). Pyocyanin may play
a role in energy metabolism under non-optimal growth conditions in Pseudomonas
aeruginosa (Whooley and McLoughlin, 1982).
In previous experiments, it has been shown that HCBD dechlorination rates incubated
with ADE were 16 - fold faster compared to activated sludge. An understanding of the
components and their function could consequently result in increased HCBD
dechlorination. In addition, in all cultures, HCBD dechlorination was observed to stall
after a number of days. Investigations into why dechlorination stalls and identifying how
to sustain dechlorination could prove beneficial.
The specific aims of this chapter were to investigate 1) the components in ADE and
activated sludge involved in HCBD dechlorination and 2) why dechlorination stalls.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
109
2 Experimental Procedures
2.1 Medium Composition
2.1.1 Biomass and Supernatant Dilutions
The ADE cultures were centrifuged at 10,000 g for 15 minutes. The pellet with biomass
was normalized to 1 gram wet weight. For the dilution series, different concentrations of
ADE medium (supernatant) were re-suspended in different volumes of de-ionised water.
ADE biomass concentrations (0.2, 0.1, 0.05 and 0.025 grams wet weight) and various
dilutions of supernatant (undiluted, diluted 2 - fold, diluted 4 - fold, diluted 8 - fold and
diluted 16 - fold) were all incubated in the presence of cyanocobalamin (0.01 mM).
2.1.2 ADE Supernatant Test
The ADE culture was centrifuged at 10,000 g for 15 minutes. The pellet with ADE
biomass was re-suspended in anaerobic medium (Treatment 1) to test how well the
biomass could dechlorinate HCBD. The supernatant was filtered (pore size, 0.25 µm) and
incubated with activated sludge biomass (Treatment 2) to test how well the medium,
which possibly contained suspended mediators, could influence HCBD dechlorination.
As control 1, ADE culture subjected to centrifugation was re-incubated with biomass to
test for the effect of centrifugation on C4 gases production. As control 2, Activated
Sludge bacteria was suspended in anaerobic medium to test how well the activated sludge
bacteria could dechlorinate HCBD without the aid of mediators suspended in ADE
medium. Biomass was normalized to 5 grams in all cultures. All cultures were incubated
with cyanocobalamin (0.1 mM).
110
2.1.3 Heating and Drying ADE Supernatant
The ADE culture was centrifuged at 10,000 g for 15 minutes. In the heating experiment,
the supernatant (30 mL) was removed and heated at 80 °C for 20 minutes. This heated
ADE supernatant was then cooled for 1 hour and 25 mL incubated with ADE biomass
(0.2 grams). The control used was ADE bacteria (0.2 grams) incubated with 25 mL ADE
supernatant (not heated).
In the drying experiment, the ADE supernatant (30 mL) was removed and heated at 80 °C
overnight. The dried powder from the heated ADE supernatant (0.26 grams) was then re-
dissolved in 30 mL of de-ionised water and incubated with ADE biomass (0.2 grams).
The controls used were ADE biomass incubated with ADE supernatant from the
centrifugation process and ADE biomass suspended in de-ionised water. All cultures
were incubated with cyanocobalamin (0.1 mM).
2.1.4 Extraction of Polyphenolics
Thirty mL of ADE supernatant was heated at 80 °C overnight and dried down to
approximately 0.26 grams powder. This powder was suspended in 4 mL of the following
solvents: Water, Ethanol, Acetone, Propan-2-ol, 1-Butanol, Acid (1 M), Base (1 M). The
reactions was agitated vigorously and left to stand in the fume hood overnight. Upon the
evaporation of the solvents the following morning, any residual matter in the test tubes
was re-suspended in an equivalent volume (4 mL) of water. Water was used to reduce
toxicity to bacteria during incubation. These solvent-extracted-water solutions were used
for incubation with activated sludge biomass.
111
Cultures of activated sludge culture were centrifuged at 10,000 g for 15 minutes. The
pellet with activated sludge biomass was normalized to 1 gram wet weight. The biomass
was re-suspended in de-ionised water (2 mL) (Negative Control) and in ADE supernatant
(2 mL) to test how well the biomass could dechlorinate HCBD. All other solvents were
incubated at equivalent volumes (2 mL) with activated sludge biomass as separate
cultures. All cultures were incubated with cyanocobalamin (0.1 mM).
2.1.5 Activated Sludge Biomass and Supernatant of Differing Ages
Cultures of activated sludge (Fresh and 6 - week old) were centrifuged at 10,000 g for 15
minutes. The pellets with activated sludge biomass were weighed (normalized to 1 gram)
and re-suspended in either fresh or 6 - week old supernatant from the centrifugation
process. Cultures were supplemented with cyanocobalamin (0.4 mM).
2.1.6 Extractions from Waste Substances
Pine tree bark, eucalyptus tree bark, eucalyptus tree leaf, banana peel, and chicken
manure were normalized to 20 grams and heated in 200 mL de-ionised water at 80 °C for
1 hour in individual Schott® bottles.
A 50 mL culture of ADE was centrifuged at 10,000 g for 15 minutes. The pellet with
ADE biomass normalized to 1 gram wet weight per reaction/extraction was used as
biomass source. The biomass was then re-suspended in activated sludge supernatant and
in ADE supernatant to serve as comparisons against all other extractions. Cultures were
incubated with cyanocobalamin (0.4 mM).
112
2.2 Dechlorination Experiments
Cultures were both topped up to final volumes of 30 mL in 100 mL serum bottles or 5
mL in 10 mL serum bottles with phosphate (5 mM, pH 7.0) and carbonate buffer (20
mM, pH 7.0). In one experiment, AQDS was added at a concentration of 1mM.
Headspaces were flushed using a N2:CO2 (80:20) mix. The temperature was set at 55 °C.
All other set-ups were similar to those described in Chapter (section 2.2).
2.3 Sampling and Analyses
Detection methods were similar to those described in Chapter 2 (section 2.3).
2.4 Studies using Cyclic Voltammetry
Cyclic Voltammetry (CV) was performed with a potentiostat (EG&G, Princeton Applied
Research, model 362 scanning electron potentiostat) interfaced to a personal computer. A
graphite rod working electrode (with contact surface area of 4.2 cm2 (21 mm length and 8
mm diameter), a platinum sheet counter electrode with contact surface area of 4.6 cm2 (21
mm length and 11 mm width), and an Ag/AgCl/saturated KCl reference electrode were
used in a 150 mL glass vessel. Before and after each measurement, the working electrode
was polished using a find abrasive paper, and were cleaned thoroughly with absolute
ethanol and deionised water. All three electrodes were inserted into the vessel without
any contact among them. 100 mL of various solution samples was carefully added into
the flask. Prior to each measurement, the solution inside the flask was continuously
flushed with pure nitrogen gas for 20 minutes to remove oxygen. CV was performed with
consortia in spent broth and with centrifuged consortia freshly resuspended in
113
physiological solution (i.e. 50 mM phosphate buffer at pH 7). Control CV experiments
with only physiological solutions were conducted to normalize the effect of suspension
medium on electrochemical activity of the measurement. To obtain a measurement
without components released into the solution, the bacterial cultures were centrifuged for
at 11,000 g 10 minutes, and then resuspended in an equal amount of physiological
solution (50 mM phosphate buffer), and flushed with nitrogen gas for 20 minutes. Both
the resuspended bacteria and the original supernatant were tested by using CV.
3 Results and Discussion
3.1 Effect of ADE Biomass and Supernatant Concentration on HCBD
Dechlorination
It was shown that HCBD dechlorination rates were enhanced with the use of ADE. A
better understanding of the responsible factor (ADE bacteria or ADE supernatant) in
HCBD dechlorination could allow further process optimization. In this experiment, ADE
biomass concentrations and supernatant concentrations were studied comparatively for
their effect on C4 gases production.
Diluting the ADE medium reduced HCBD dechlorination rates. The rate of C4 gases
production* was reduced by approximately 7 - fold over 5 days (Fig. 5.1) when the ADE
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
114
supernatant was diluted by 2 - fold.
0
20
40
60
80
100
120
140
160
180
0 2 4 6 8 10 12Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.1 Effect of ADE supernatant on the concentration of C4 gases* from HCBD dechlorination. Supernatant concentrations used were undiluted (▲), diluted 2 - fold (■), diluted 4 - fold (●), diluted 8 - fold (∆) and diluted 16 - fold (□).
Equivalent dilutions were used for biomass concentrations. The rates and concentrations
of C4 gases production* were comparable in all biomass dilutions tested. Hence, the
dilution of ADE biomass does not play a crucial role in influencing concentration and
rates of C4 gases production (Fig. 5.2). Mediators dissolved in ADE supernatant seem to
play a crucial role in HCBD dechlorination.
115
0
50
100
150
200
250
0 2 4 6 8 10 12
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.2 Effect of ADE biomass on the concentration of C4 gases from HCBD dechlorination. Biomass wet weight used were 0.2 (▲), 0.1 (■), 0.05 (●) and 0.025 (∆) grams.
3.2 Effect of ADE Supernatant on HCBD Dechlorination
It was shown from the previous result that diluting the ADE supernatant inhibited HCBD
dechlorination to a greater extent compared to diluting ADE bacteria. In order to study if
the higher rates observed with the ADE was indeed due to ADE supernatant, ADE
supernatant was incubated with activated sludge bacteria and ADE bacteria. This was
done to further assess the role of mediators dissolved in the ADE supernatant on HCBD
dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
116
Both ADE bacteria and supernatant seemed to stimulate dechlorination. However, the
effect from the ADE supernatant was stronger after Day 5. Dechlorination was greatest in
activated sludge cultures amended by ADE supernatant (Fig. 5.3).
0
50
100
150
200
250
300
0 2 4 6 8 10 12
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.3 Effect of bacteria or supernatant on the concentration of C4 gases from HCBD dechlorination.
This implies that the mediators in the supernatant rather than specific bacteria were
responsible for the fast reaction rates observed in the full ADE. It should be pointed out
that the suspected mediator in the ADE did not enable dechlorination by itself but only in
combination with cyanocobalamin. In combination with results shown above, the likely
reason is that ADE contains its own mediators. It is likely that a combination of different
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
Activated Sludge bacteria + ADE supernatant (Treatment 2)
ADE bacteria + ADE supernatant (Control 1)
ADE bacteria + Basal medium (Treatment 1)
Activated sludge bacteria + Basal medium (Control 2)
117
redox mediators with different redox potential midpoints is more effective than using one
single mediator.
3.3 Effect of Heated ADE Supernatant on HCBD Dechlorination
It has been shown that ADE mediators dissolved in the supernatant play a crucial role in
HCBD dechlorination. In order to understand the characteristics of the mediators
involved in HCBD dechlorination, as a first step, the heat stability of the mediators
involved in HCBD dechlorination was tested.
ADE mediators involved in HCBD dechlorination were heat stable. Results indicate that
heated ADE supernatant had no effect on HCBD dechlorination rates (Fig. 5.4)
compared to the control.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
118
020406080
100120140160180200
0 5 10 15 20
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cul
ture
)
Figure 5.4 Effect of heated ADE supernatant (▲) and control (■) on the concentration of C4 gases from HCBD dechlorination.
3.4 Effect of Drying ADE Supernatant on HCBD Dechlorination
Since it has been shown that ADE supernatant is heat stable, the effect of drying and re-
dissolving ADE supernatant on HCBD dechlorination was tested. This was done to
further explore the nature of the mediators in ADE supernatant involved in HCBD
dechlorination.
Results indicate that the ADE supernatant involved in HCBD dechlorination was active
after the drying and re-suspending process (Fig. 5.5). In addition, the rates of C4 gases
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
119
production were comparable to ADE supernatant not subjected to the drying process.
Thus, ADE supernatant could be dried down, possibly for easy storage.
0
20
40
60
80
100
120
140
160
180
200
0 5 10 15 20
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.5 Effect of dried ADE supernatant on the concentration of C4 gases* from HCBD dechlorination. Cultures used were ADE biomass in ADE supernatant (■), ADE biomass in ADE supernatant (dried and re-suspended in de-ionised water) (▲) and ADE biomass in de-ionised water (Control) (●).
3.5 Effect of AQDS (to Replace) ADE Mediators on HCBD
Dechlorination
Humic substances, organic materials found in terrestrial environments, were thought to be
a major component of ADE. This is because the starting materials used for the ADE
anaerobic digestion process were largely plant based material. Here, the humic analogue,
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
120
AQDS (compared to ADE medium) was tested for its effect on HCBD dechlorination, in
the presence of cyanocobalamin. Any HCBD dechlorination by AQDS in the presence of
cyanocobalamin would indicate that it was present in ADE medium and that AQDS was
responsible for HCBD dechlorination. No C4 gases were noticed in the incubation with
ADE in the presence of AQDS and cyanocobalamin (Fig. 5.6). AQDS is not the essential
mediator (in ADE) required for HCBD dechlorination.
0
5
10
15
20
25
30
35
40
45
50
0 2 4 6 8
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.6 Effect of AQDS and cyanocobalamin on the concentration of C4 gases from HCBD dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
ADE bacteria with AQDS (1 mM) and cyanocobalamin (0.1 mM)
ADE bacteria with DAR supernatant with cyanocobalamin (0.1 mM)
121
2.3 Determining the Electroactive Specie(s) Present in ADE Using
Cyclic Voltammetry
Cyclic voltammetry (CV) is an electrochemical technique that may be employed to study
bacterial use of redox shuttles to transfer electrons, or release the electrons via their
membrane complex directly to the electrode. CV offers information on the extent of the
redox mediation and the midpoint potentials (E0) of the electrochemically active
compounds in the system. An analysis of the supernatant of the centrifuged bacterial
culture is indicative of the presence of mobile, suspended redox shuttle(s).
The aim of the present study was to determine the number of mid point potential(s) of
electron shuttle(s) in the ADE.
Figure 5.7 Cyclic voltammograms of (A) supernatant and (B) re-suspension of ADE liquor collected from the reactor. (Re-suspension with 50 mM phosphate buffer at pH 7, 20±1oC; initial and final potentials were -700 and +700 mV, respectively).
-6
-4
-2
0
2
4
6
8
10
-800 -600 -400 -200 0 200 400 600 800
DiCOM_Supernatant_10mV/secDiCOM_Supernatant_20mV/secDiCOM_Supernatant_50mV/sec
-5
-4
-3
-2
-1
0
1
2
3
4
5
-800 -600 -400 -200 0 200 400 600 800
DiCOM_Resuspension_10mV/secDiCOM_Resuspension_20mV/secDiCOM_Resuspension_50mV/secPSB 10 mV/sec
Potential vs. EAg/AgCl (mV)
Cur
rent
(mA
) (A) (B)
122
Components that could be reversibly oxidized or reduced would show a peak on both the
upper and lower curves of the cyclic voltammogram. The potential at which a straight
line joining the upper and the lower peaks intersected with the x-axis (i.e. potential) at 0
mA, is considered as the E0 of the component. This was found to be approximately
0 mV (Fig. 5.7).
3.5 Effect of Polyphenolic Extractions of ADE Supernatant on HCBD
Dechlorination
Polyphenolics are compounds that are the most abundant secondary metabolites found in
plants. Thus, polyphenolics can be found in many foods (e.g., legumes (cereal, rice),
wheat, fruits and vegetables). Since ADE is the product of anaerobic digestion of
municipal waste (predominantly waste from households), it can be expected that ADE is
polyphenolics-rich.
It has been established, thus far, that the ADE supernatant contains mediators that are
heat-stable, and can be dried and re-dissolved without affecting HCBD dechlorination
rate. The extraction and subsequent identification of these mediators would be an
advantage in improving HCBD dechlorination. Therefore, the mediators in ADE
supernatant were extracted in a range of solvents. The extracted mediators’ effect was
then tested on HCBD dechlorination.
ADE mediators involved in HCBD dechlorination were extractable in all the solvents and
water extracted fractions of ADE, except acetone and propan-2-ol (Fig. 5.8). However, no
123
single extraction was comparable to ADE supernatant, in terms of rates and concentration
of C4 gases produced (Fig. 5.8). The rates and concentration of C4 gases* produced in
all solvent-extracted fractions were approximately half compared to ADE supernatant.
This is a sign that there are multiple solvent and water extractable mediators in the ADE
supernatant.
0
5
10
15
20
25
30
35
0 1 2 3 4 5 6 7
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/ L
cul
ture
)
Figure 5.8 Effect of different solvent extractions of ADE supernatant on the concentration of C4 gases * from HCBD dechlorination.
Procyanidins and catechins are two commonly occurring polyphenolics. Generally,
procyanidins are water soluble while catechins are lipid soluble (Shi et al., 2005). The
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
ADE
Base
Ethanol
1-butanol
Acid
Propan-2-ol
Negative control
Water
Acetone
124
extractions of these compounds are based on their solubility either in water or organic
solvents. Our results revealed that the polyphenolics extracted may be both procyanidins
and catechins, since all of the tested water and solvent (except acetone and propan-2-ol)
extracts dechlorinated HCBD.
ADE contains a mixture of waste substances as starting material. The isolation and
extraction of polyphenolics from the starting material and the subsequent testing for
HCBD dechlorination was believed help target the essential compound(s) that behave(s)
as mediator(s) in HCBD dechlorination.
3.6 Effect of Extractions from Waste Substances on HCBD
Dechlorination
It was shown that HCBD dechlorination rates were enhanced upon the addition of ADE
supernatant to activated sludge biomass (Fig. 5.3). Since, it has already been shown that
ADE mediators in the supernatant play a crucial role in HCBD dechlorination,
understanding the source and nature of the mediators in the ADE supernatant involved
would enable the improvement in dechlorination.
ADE is the product of anaerobic digestion of municipal waste, which predominantly
consists of waste from households (e.g. waste from fruit or vegetable peels, the garden,
etc.). It is possible that the majority of mediators in ADE supernatant responsible for
HCBD dechlorination could have been extracted from any one (or few) of the waste
substances. Extractions from various compounds that may be found in ADE supernatant
125
were tested for their effect on HCBD dechlorination. This was done to assess the role of
any of those extractions, from the waste substances tested, to act as potential mediator(s)
for HCBD dechlorination.
Results show that all the extractions tested (except pine bark and banana peel) contained
mediators able to enable HCBD dechlorination (Fig. 5.9). Dechlorination comparable to
ADE supernatant was observed in the eucalyptus tree bark and leaf. However, higher
dechlorination was observed in activated sludge supernatant (9 - fold) and in the chicken
manure extraction (5 - fold).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
126
0
10
20
30
40
50
60
70
80
90
100
0 5 10 15 20
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cul
ture
)
Figure 5.9 Effect of extractions from pine tree bark (▲), eucalyptus tree bark (●), eucalyptus tree leaf (○), banana peel (□), chicken manure (■), activated sludge supernatant (∆) and ADE supernatant (X) on the concentration of C4 gases* from HCBD dechlorination.
Since HCBD dechlorination was enabled in all the extractions from the waste substances
tested, except in pine bark and banana peel extract, mediators dissolved in ADE
supernatant could have evolved from a mixture of waste substances.
3.7 Effect of Activated Sludge Biomass and Supernatant on HCBD
Dechlorination
So far, it was observed that HCBD dechlorination occurred in activated sludge (dissolved
in its own supernatant) combined with cyanocobalamin (Chapter 2). Moreover, HCBD
dechlorination did not occur when activated sludge supernatant was removed (Fig. 5.3).
127
These results seem to suggest some level of dependence by activated sludge bacteria on
the supernatant (containing the dissolved mediators) for HCBD dechlorination reaction to
proceed.
HCBD dechlorination was observed in a test with anaerobically incubated activated
sludge after 2 days of incubation. In that test, a 6 - week old culture (activated sludge
biomass and supernatant) was used as inoculum. Comparatively, in all other previous
dechlorination tests, a lag phase of 7 - 14 days was observed. In those other previous
dechlorination studies, fresh culture (activated sludge biomass and supernatant) was used
(Chapter 2). Thus, anaerobically incubated activated sludge (over 6 weeks) used as
inoculum showed increased HCBD dechlorination and a reduced lag phase.
It was postulated that mediators produced by activated sludge bacteria (upon anaerobic
incubation) enabled the increased HCBD dechlorination as well as the reduced lag phase
observed. The aim of this experiment was to test how well HCBD dechlorinated in
individual cultures where fresh and 6 - week old activated sludge biomass were exposed
to batches of fresh and 6 - week old supernatants.
Results showed that 6 - week old activated sludge supernatant incubated with both fresh
and old activated sludge bacteria showed higher (between approximately 2 - 6 fold) total
HCBD dechlorination compared to fresh supernatant incubated with both fresh and old
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
128
activated sludge bacteria over 15 days (Fig. 5.10). Thus, the age of the supernatant
seemed to play a role in HCBD dechlorination.
0
20
40
60
80
100
120
140
0 5 10 15 20
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.10 Effect of activated sludge biomass and supernatant on the concentration of C4 gases from HCBD dechlorination. Fresh activated sludge with its own fresh supernatant (■), fresh activated sludge with 6 - week old supernatant (●), 6 - week old activated sludge bacteria with fresh supernatant (∆) and 6 - week old activated sludge with its own 6 - week old supernatant (▲).
HCBD dechlorination was observed after 2 days in a fresh batch of activated sludge
(biomass and supernatant). This is in contrast to previous observations where a lag phase
of 7 - 14 days was noticed. This could be due to a variation in activated sludge collected
at different times.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
129
It is possible that activated sludge bacteria, when incubated anaerobically over 6 weeks,
caused a change on the supernatant composition by releasing synthesized mediators into
the supernatant and that result in enhanced HCBD dechlorination.
3.8 Investigations into Sustaining HCBD Dechlorination
In all cultures tested so far, reactions stalled after initial dechlorination. The identification
and elimination of the key factor or factors that cause the stalling of reactions could lead
to sustained dechlorination. In this experiment, possible reasons into why HCBD
dechlorination reactions stalled were investigated. End-product inhibition, electron donor,
mediator and acceptor depletion were tested in different incubations.
HCBD dechlorinating cultures were amended with the following treatments after initial
HCBD dechlorination ceased between 1 - 3 days (Fig. 5.11 - 5.16).
1) Headspace flushing
2) HCBD addition
3) Acetate addition
4) Acetate and cyanocobalamin addition
5) Headspace flushing followed by second headspace flushing, acetate and
cyanocobalamin addition
Headspace flushing re-initiated HCBD dechlorination (Fig. 5.11). This observation was
similar to previous findings that showed headspace flushing re-initiated HCBD
130
dechlorination (Fig. 4.8). However, on repeated flushing, the rates and concentration of
C4 gases produced* were reduced.
0
20
40
60
0 2 4 6Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.11 Effect of repeated headspace flushing on the concentration of C4 gases from HCBD dechlorination. Headspace was flushed on Day 1, 1.9 and 2.5.
It is apparent from results that gases that tend to build up in the headspace upon initial
dechlorination inhibit further dechlorination. Upon removal of those gases, HCBD
dechlorination was revived, though, at lower rates and concentrations. Headspace
degassing removes HCBD along with partially dechlorinated intermediates in the gaseous
phase. Any of those dechlorinated gaseous intermediates could have possibly been
responsible for inhibiting further dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
Headspace flushed
131
It was found that the concentration of HCBD in solution was limiting in cultures (Fig.
5.12). In 24 hours, approximately 1 % of the 1 mM added at the start of the incubation
remains in the test vials. Hence, more HCBD (1 mM) was added to test vials (Fig. 5.13).
0
10
20
30
40
50
60
70
80
90
100
0 5 10 15 20 25 30
Time (Hours)
HC
BD
Ava
ilabl
e (%
)
Figure 5.12 Percent availability of HCBD in solution.
The addition of HCBD increased C4 gases production (Fig. 5.13). A possible postulation
is that microbial absorption or adsorption of HCBD leads to a lower level available for
dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
132
0
50
100
150
200
250
300
0 2 4 6 8 10
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.13 Effect of HCBD addition on the concentration of C4 gases from HCBD dechlorination.
Toxicity to bacteria by HCBD may not be a likely cause of the stall due to the finding
that dechlorination continues even after fresh addition of HCBD.
The addition of acetate did not increase C4 gases production* (Fig. 5.14). Therefore,
electron donor limitation was not a likely factor involved in stalling HCBD
dechlorination reactions.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
HCBD (1 mM) added
133
0
20
40
60
80
100
120
0 2 4 6Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.14 Effect of acetate on the concentration of C4 gases from HCBD dechlorination.
Cyanocobalamin degradation/consumption may be one reason why dechlorination stops.
In one study, the concentration of cyanocobalamin decreased from 1 mM to 0.6 mM after
approximately 5 days (data not shown). Therefore, cyanocobalamin was added to test for
its effect on sustaining HCBD dechlorination.
Acetate and cyanocobalamin addition increased HCBD dechlorination* (Fig. 5.15).
Since the addition of acetate alone did not increase C4 gases produced (Fig. 5.14), it can
be deduced that cyanocobalamin addition was most likely the reason for the increased C4
gases production.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
Acetate (40 mM) added
134
0
50
100
150
200
250
300
0 2 4 6
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.15 Effect of acetate and cyanocobalamin on the concentration of C4 gases from HCBD dechlorination.
The most likely reasons are that cyanocobalamin is consumed in bacterial cultures or that
the chemical structure of cyanocobalamin is altered upon bacterial incubation rendering
the role of electron shuttle ineffective.
It has been found earlier that headspace degassing re-initiated C4 gases production* (Fig.
5.16) and that cyanocobalamin addition increased C4 gases produced*. The collective
effect of the two separate amendments was believed to increase HCBD dechlorination
even further. Hence, the combined effect of headspace flushing, and acetate and
cyanocobalamin addition on HCBD dechlorination was tested. Headspace flushing, in
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
Acetate (40 mM) added + Cyanocobalamin (0.4 mM) added
135
combination with cyanocobalamin and acetate addition, increased HCBD dechlorination
(Fig. 5.16).
0
20
40
60
80
0 2 4 6Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 5.16 Effect of repeated headspace flushing and addition of acetate and cyanocobalamin on the concentration of C4 gases from HCBD dechlorination.
It has been shown that HCBD dechlorination rates were enhanced upon the addition of
ADE supernatant to activated sludge (Fig. 5.3). The ADE supernatant, containing
dissolved mediators, was responsible for the increased HCBD dechlorination rates. These
mediators, along with cyanocobalamin, essentially behaved as redox couples that
facilitated electron flow in a cascade (i.e., from the electron donor to the electron
acceptor).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
Headspace flushed + Acetate (40 mM) added + Cyanocobalamin (0.01 mM)
Headspace flushed
136
In all cultures tested so far, HCBD dechlorination stalled after a number of days. From
the results (Fig. 5.11 - 5.16), it appears that HCBD dechlorination was revived on 2
conditions. Firstly, when headspace was degassed, and secondly, when fresh HCBD and
cyanocobalamin were added. Although both headspace degassing and additions of fresh
HCBD and cyanocobalamin revived HCBD dechlorination, reactions stalled soon
thereafter. It appears that this may be an inherent and inevitable hurdle in HCBD
dechlorination.
A sacrificial control subjected to an extraction protocol was performed to account for the
completely dechlorinated products that may have dissolved in the aqueous phase or
biomass. This control was an activated sludge culture with a known concentration of
completely dechlorinated gases injected into the headspace. MS readings revealed that
over 7 days, the completely dechlorinated gases remained stable in the gas phase showing
none dissolved in the aqueous phase. In addition, upon degassing no new completely
dechlorinated gases were formed (data not shown).
The use of thermophilic conditions has been associated with high rates of dechlorination
upon the extra addition of chlorinated solvents after initial dechlorination. Larsen et al.
(1991) noted that extra additions of PCP after its initial degradation (over 8 months) were
degraded in a relatively short period of time (4 weeks). However, in our studies,
increased dechlorination rates were not observed upon subsequent addition of HCBD.
Even though HCBD can be dechlorinated in relatively short periods of time, there seems
to be a limit on the concentrations of HCBD dechlorinated to C4 gases.
137
4 Conclusion
In this chapter, the following points were demonstrated.
Mediators dissolved in ADE supernatant could have evolved from a mixture of
waste substances and play a crucial role in HCBD dechlorination. The midpoint
potential of those mediators was found to be 0 mV and unlikely humics.
These mediators rather than specific bacteria were responsible for the fast reaction
rates observed in the full ADE.
The mediators in ADE are heat stable, can be dried for easy storage without
affecting HCBD dechlorination rates and are multiple solvent- and water-
extractable.
It is plausible that when activated sludge is incubated anaerobically a change on
the supernatant composition resulted in enhanced HCBD dechlorination by the
release of synthesized mediators.
It appears that HCBD dechlorination could be sustained on 2 conditions. Firstly,
when headspace was degassed, and secondly, when fresh HCBD and
cyanocobalamin were added.
138
Chapter 6
The Use of Redox Potential to Monitor
HCBD Dechlorination4
4 This chapter was published in Journal of Biotechnology (142) 151 - 156.
139
1 Introduction
Hexachloro-1,3-butadiene (HCBD) is a toxic, aliphatic chlorinated hydrocarbon. It is
carcinogenic, mutagenic and fetotoxic, and is produced as a by-product from the
production of tetrachloroethene, trichloroethene and carbon tetrachloride (Booker and
Pavlostathis, 2000). It was also used as fungicide, herbicide and heat transformer fluid
(Verschueren, 1996). The world annual production was estimated to be 10,000 tonnes in
1982 (IPCS, 1994) and the calculated emission of HCBD in Europe for the year 2000 was
2.59 tonnes/year (Van der Honing, 2007).
Due to the highly oxidized state of the carbon atoms in HCBD and the highly
electronegative halogen substituents, biodegradation in the form of reductive
dechlorination is more likely to occur than the more traditional biodegradation via
oxidative processes (Pavlostathis et al., 2002). Only few studies are published on the
microbial dechlorination of HCBD and a slow dechlorination has been documented
resulting in the formation of partly dechlorinated products such as pentachlorobutadiene,
tetrachlorobutadiene, trichlorobutadiene and dichlorobutadiene (Bosma et al., 1994;
Booker and Pavlostathis, 2000). The dechlorination of HCBD to completely
dechlorinated endproducts (C4 gases) has recently been described as a process in which
the presence of cyanocobalamin was essential as an electron shuttle between mixed
microbial consortia and HCBD (James et al., 2008).
In the process of microbial reductive dechlorination, bacteria use chlorinated species as
the electron acceptor. Hence, in the presence of dechlorinating bacteria, the presence of a
140
chlorinated hydrocarbon represents oxidative power. Highly reduced anaerobic
environments, (indicated by a low redox potential) typical for methanogenesis, have been
found to be a requisite for the reductive dechlorination of halogenated compounds (i.e.,
the substitution of halogen atoms by hydrogen atoms) (Stuart et al., 1999). The process of
biochemical reductive dechlorination has been shown for a number of chlorinated
solvents.
The biochemical dechlorination of tetrachloroethylene (PCE) by various enzymes has
also been shown by bacterial transition metal coenzymes such as vitamin B12, coenzyme
F430, and hematin, as well as by corrinoid-containing enzymes (Ensley, 1991; Furukawa
et al., 2005; Burris et al., 1996; Gantzer and Wackett, 1991; Glod et al., 1997; Jablonski
and Ferry, 1992). These reactions have been reported to occur cometabolically or coupled
to energy generating reactions where PCE serves as an electron acceptor (Fathepure et
al., 1987; Gerritse et al., 1996; Holliger et al., 1993).
Several other studies have also shown that low redox potentials are required for
dechlorination (Arnold and Roberts, 2000; Masscheleyn et al., 1991; Olivas et al., 2002;
Pardue et al., 1988; Schumacher et al., 1997; Shimomura and Sanford, 2005; Stuart et al.,
1999). The highest rates of dechlorination were observed at lowest redox potentials tested
(EAg/AgCl = -348 mV) (Olivas et al., 2002).
Shulder (2006) noted that redox potential measurements can be an effective control
parameter for maintaining an oxidizing, or reducing environment. For example, redox
141
potential measurement were used to control the bioleaching of chalcopyrite (Third et al.,
2000) by maintaining conditions that were neither too reduced nor to oxidized for
optimum process conditions. While redox potential measurements have been used to
confirm that sufficiently reducing conditions were present to enable dechlorination, to our
knowledge they have not been described as a method for the monitoring of the reductive
dechlorination of substances such as HCBD.
HCBD and other chlorinated hydrocarbons are known to be toxic to dechlorinating
bacteria (Blum and Speece, 1991) and a stalling of the dechlorination reaction has been
documented (James et al., 2008). Hence, to avoid toxicity and stalling, these chlorinated
hydrocarbons need to be added ‘on- demand’. This requires the online detection of the
presence of chlorinated species.
The aim of this chapter was to evaluate the possibility of using redox potential
measurement during the microbial reductive dechlorination of a suitable chlorinated
hydrocarbon (HCBD) for online detection of ongoing dechlorination.
2 Experimental Procedures
2.1 Medium Composition
ADE digestion methods were similar to those described in Chapter 4 (section 2.1.1).
142
2.2 Reactor Set-Up
Hundred mL of anaerobic digested effluent was filled into a Schott® bottle and covered
with a rubber stopper. A Schott® bottle cap, with a 3 cm diameter hole drilled through its
top, was used to screw down the rubber stopper onto the Schott® bottle. A redox probe
(EAg/AgCl) was inserted through a hole made in the middle of the rubber stopper. The
Schott® bottle was immersed in a 500 mL water bath controlled at 55 °C (Fig. 6.1).
Figure 6.1 Reactor set-up, showing a stirred anaerobic batch reactor with a built-in, computer monitored redox probe in a stirred waterbath for temperature control at 55 oC.
2.3 Dechlorination Experiment
Biomass used were centrifuged at 10,000 g for 15 minutes and normalized to 5 grams in
all trials. Thermophilic bacteria from anaerobic digested effluent were used because
HCBD dechlorination rates were 16 - fold higher compared to activated sludge. All other
methods were similar to those described in Chapter 2 (section 2.2).
143
2.4 Culture Conditions for Redox Potential Measurements
The anaerobic effluent was supplemented with acetate (40 mM) and cyanocobalamin (0.1
mM) and agitated at 180 revolutions per minute. Agitation and temperature control were
provided via a stirrer (IKA® RCT BASIC). A stock concentration of HCBD (16 mM) in
ethanol was prepared. HCBD was dissolved in ethanol, in order to provide a more
uniform distribution of the poorly water soluble HCBD. The low concentration of ethanol
had been tested to not affect dechlorination rates in the presence of saturating
concentrations of acetate (Fig. 6.8). When required, 0.32, 0.64, 1.28, 2.56, 16 and 32 µM
were dispensed into the reactor.
Prior to measurements of the microbial reduction rate, the anaerobic digested effluent
needed to be oxidized by a suitable oxidant. Hydrogen peroxide (H2O2) (Sigma catalog
No. 7722-84-1) (1:20 diluted with de-ionised H2O) was chosen as the oxidant because a
previous study had shown that low concentrations of H2O2 did not interfere with the
microbial dechlorination of pentachlorophenol (Stuart et al., 1999).
To minimize the effect of other potential electron acceptors, such as nitrate or sulfate or
ferric, the anaerobic digested effluent had been kept anaerobic prior to experiments until
the redox potential had stabilised.
2.5 Calibration of Redox Electrodes
Ag/AgCl redox reference electrodes (ionode® intermediate junction - IJ 64) were used in
all experiments. Calibration was performed using ZoBell’s solution [3.2 mM potassium
144
ferrocyanide (K4Fe(CN)6·3H2O) and 2.8 mM potassium ferricyanide ((K3Fe(CN)6) in 0.1
M potassium chloride (KCl)) (ionode® Redox Electrode Manual)]. All redox potentials
(referred to as EAg/AgCl) were referenced to Ag/AgCl electrolyte (-0.199 V vs. Standard
Hydrogen Electrode (SHE)).
2.6 Calculations
Redox potentials were recorded online via LabView® (National Instruments) every 10
seconds. The recorded redox potentials were averaged using a running average of 10
values to achieve a smooth plot of EAg/AgCl vs. time. The rate of redox potential change
was then obtained from the gradient of the averaged records.
2.7 Sampling and Analyses
Detection methods were similar to those described in Chapter 2 (section 2.3).
2.8 Fuel Cell Setup
Dechlorination of HCBD was performed in a two-chamber fuel cell made of transparent
Perspex. The two chambers were physically separated by a cation exchange membrane
(CMI-7000, Membrane International Inc.). Conductive reticulated vitreous carbon (RVC)
blocks (ERG, Oakland, CA) with 80 pore per inch (ppi) and dimension (6.5 cm x 5.5 cm
x 1 cm) were used as the anode and the cathode. The electrodes were linked to a scanning
potentiostat (Model 362, Elmeasco Instruments Pty. Ltd.) with copper wires. The positive
pole of the potentiostat was connected to the cathode, while the negative pole was
145
connected to the anode. The reference pole of the potentiostat was connected to
Ag/AgCl/saturated potassium chloride placed in the anode chamber.
The anode and cathode chambers were filled with 500 mL of 100 mM phosphate buffer
(pH 7). The fuel cell was placed into a water bath maintained at 55 °C. Both chambers
were sealed and the anolyte and catholyte were constantly mixed with a magnetic stirrer.
The cathode chamber was supplemented with cyanocobalamin to a final concentration of
0.1 mM. A voltage in a range of 5 to 9 V was applied to the circuit using the potentiostat
(applied voltage was adjusted manually) to manipulate catholyte EAg/AgCl in a range of -
300 to -800 mV. HCBD was added to a final concentration of 1 mM when the EAg/AgCl of
the catholyte was reduced to -300 mV. A headspace sample was analysed for C4 gasses
after half an hour at the catholyte EAg/AgCl of -300 mV. The EAg/AgCl of the catholyte was
subsequently reduced to -580 mV, -720 mV and -800 mV, each EAg/AgCl was maintained
for half an hour before taking a headspace sample. A control experiment was conducted
by dechlorinating HCBD in the absence of cyanocobalamin.
3 Results and Discussion
3.1 Online Monitoring of HCBD Dechlorination
Highly reduced anaerobic environments are a pre-requisite for the reductive
dechlorination of HCBD. Based on our preliminary study, a typical EAg/AgCl suitable for
HCBD dechlorination was around -550 mV (Fig. 6.2). In view of the fact that HCBD is
an oxidising agent and dechlorination of HCBD is simply a transfer of electrons onto
146
HCBD, the dechlorination process should be able to be monitored by the change of redox
potential. A series of experiments were performed to investigate the relationship between
redox potential and HCBD dechlorination, and the possibility of using EAg/AgCl
measurements as a means of monitoring HCBD dechlorination. In this study, the
dechlorination was obtained using thermophilic microbes from anaerobic effluent
incubated in their natural supernatant in the presence of cyanocobalamin.
-600
-500
-400
-300
-200
-100
0
100
200
0 200 400 600 800 1000 1200Time (min)
E Ag/
AgC
l (m
V)
Figure 6.2 Effect of cyanocobalamin on EAg/AgCl.
When incubating the thermophilic anaerobic culture at 55 oC with acetate as the electron
donor, the redox potential decreased initially rapidly and more slowly towards the end
reaching a minimum of -530 mV within 12 hours (Fig. 6.2). However, dechlorination
No cyanocobalamin
With cyanocobalamin (0.1 mM)
147
could only be observed when cyanocobalamin was present (Fig. 6.3). Interestingly, the
presence of cyanocobalamin enabled the bacteria to reach a lower redox potential (-550
mV). As the microbial reductive dechlorination of HCBD requires the presence of
cyanocobalamin (James et al., 2008) (Fig. 6.3), it could be suggested that the key role of
cyanocobalamin in enabling HCBD dechlorination is in its role of lowering of the redox
potential. At that low redox potential, cyanocobalamin is itself reduced which then
transfers the electrons required for HCBD dechlorination.
0
40
80
120
160
0 0.5 1 1.5 2 2.5
Time (Days)
Con
cent
ratio
n of
C4
gase
s (µ
mol
es/L
cu
lture
)
Figure 6.3 HCBD dechlorination by thermophilic bacteria from anaerobic effluent in the presence (●) and absence (■) of cyanocobalamin (0.1 mM) as measured by the concentration of C4 gases.
The addition of HCBD caused a sudden increase in EAg/AgCl by about 100 mV (Fig. 6.4).
This increase signifies the presence of a suitable electron acceptor and its reduction (here
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
148
dechlorination) by making available an oxidizing half reaction (Equation 6.1). Within
approximately 5 hours, the redox potential decreased back to its original value (Fig. 6.4).
To confirm that the peak in redox potential observed was linked to the use of reducing
power in the microbial system, the dechlorination reaction was recorded by monitoring
the dechlorinated end-products (C4 gases)*.
C4Cl6 + 6H+ + 12e-→ C4H6 + 6Cl- Equation 6.1
0
10
20
30
40
50
60
4.8 5.8 6.8 7.8 8.8
Time (Days)
Cum
ulat
ive
conc
entra
tion
of C
4 ga
ses
(µm
oles
/L c
ultu
re)
-600
-550
-500
-450
-400
-350
-300
-250
EA
g/A
gCl (
mV
)
Figure 6.4 Effect of multiple HCBD additions (32 µM) on EAg/AgCl and on HCBD dechlorination (measured by following the cumulative concentration of C4 gases produced) (■).
Over the time EAg/AgCl increased and fell back to its original value, the formation of
dechlorination products indicated that the change in EAg/AgCl was linked to the
dechlorination of HCBD. This effect was reproducible and allowed an electrochemical Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
HCBD added (32 µM)
149
monitoring of the dechlorination reaction. Total concentration of C4 gases produced over
3 injections equated to approximately 50 % of HCBD added to the solution (Fig. 6.4).
Repeat experiments in the absence of cyanocobalamin showed that HCBD addition only
caused a marginal short term increase in redox potential (Fig. 6.5) and no detectable
endproduct or intermediates, which is in line with the original results that
cyanocobalamin is necessary for HCBD dechlorination to C4 gases (James et al., 2008).
-600
-500
-400
-300
-200
-100
04.8 5.8 6.8 7.8 8.8
Time (Days)
EA
g/A
gCl (
mV
)
Figure 6.5 Effect of HCBD additions (32 µM and 100 µM) on EAg/AgCl in the absence of cyanocobalamin.
The fact that the EAg/AgCl dropped back to its original value was interpreted as the
depletion of the oxidant (here HCBD). The fact that repeat injections of very low Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
HCBD added (32 µM)
HCBD added (100 µM)
150
concentrations of HCBD caused a peak in EAg/AgCl, over the duration of its reductive
dechlorination, suggests that the EAg/AgCl could be used as an online parameter for an
automated feeding regime, supplying new HCBD on demand and possibly avoiding the
build-up of inhibitory concentrations of HCBD.
In the absence of HCBD and the presence of acetate as the electron donor, the typical
EAg/AgCl of the anaerobic effluent was around -550 mV. This value is more negative than
what would be expected from the reduction potential of the acetate bicarbonate couple
(EAg/AgCl = -489 mV) (Kaden et al., 2002), and could be explained by the presence of
biologically synthesized organic species of a more negative redox potential such as a
sugar or perhaps pyruvic acid.
To assess the reduction capacity of this anaerobic culture, a small amount of H2O2 was
added, sufficient to increase the EAg/AgCl to -300 mV. Within 60 minutes, the microbial
communities in the anaerobic liquid reduced the EAg/AgCl from -300 mV to the original
EAg/AgCl of -550 mV (Fig. 6.6). The rate of change of the EAg/AgCl show that the reaction
rate was slowing down as the EAg/AgCl approached -560 mV. Overall, a polynomial
regression analysis best described effect of EAg/AgCl on the rate of its change (Fig. 6.7).
Interestingly, also a linear portion existed in the curve (Fig. 6.8) as would be expected
from reactions of first order. The intercept of about -560mV with the EAg/AgCl axis
indicated the lowest redox potential that could be reached.
151
-600
-550
-500
-450
-400
-350
-3000 50 100 150 200
Time (min)
EA
g/A
gCl (
mV)
Figure 6.6 Decrease in EAg/AgCl caused by acetate metabolizing anaerobic thermophilic bacteria obtained from the anaerobic effluent after the addition of H2O2.
y = 9E-05x2 + 0.1042x + 31.147R2 = 0.9929
-0.5
0
0.5
1
1.5
2
2.5
-600 -550 -500 -450 -400
EAg/AgCl (mV)
dEA
g/A
gCl/d
t (m
V/m
in)
Figure 6.7 Effect of rate of EAg/AgCl change at individual EAg/AgCl. (Polynomial regression.)
152
y = 0.0148x + 8.2212R2 = 0.9857
0
0.2
0.4
0.6
0.8
1
1.2
1.4
-600 -550 -500 -450 -400
EAg/AgCl (mV)
dEA
g/A
gCl/d
t (m
V/m
in)
Figure 6.8 Effect of rate of EAg/AgCl change at individual EAg/AgCl. (Linear regression.)
Under the assumption that a particular EAg/AgCl corresponds to a particular rate of
microbial reduction of oxidized species, one can derive that at higher EAg/AgCl values the
dechlorination proceeds at a faster rate. With more detailed studies, it should be possible
to read the rate of the real time dechlorination reaction from the EAg/AgCl obtained.
However, this was beyond the scope of this study.
With increasing concentrations of HCBD (0.32 µM to 16 µM) added, the EAg/AgCl of the
anaerobic liquid increased for longer times and to higher values (data not shown). From
the fact that the rate at which bacteria decreased the EAg/AgCl after oxidation with H2O2
(Fig. 6.6) and on the assumption that the redox buffer capacity (capacitance) of the
system was uniform over the redox potentials tested, it can be assumed that a more
153
positive redox potential implies a faster rate of dechlorination. Independent of this
assumption a larger peak area (as established by numerical integration) represents a
greater amount of HCBD dechlorination (Fig. 6.9).
0
1000
2000
3000
4000
5000
6000
7000
8000
9000
0 10 20 30 40
Concentration of HCBD (µM)
Pea
k A
rea
(mV∙
min
)
Figure 6.9 Peak areas of voltage * time observed as a function of the concentration of HCBD added.
This online monitoring of HCBD reductive dechlorination could potentially be used to
monitor the dechlorination of other chlorinated hydrocarbons and for improved process
control of bioremediation reactors. To our knowledge, the online detection and
monitoring of dechlorination using redox potential has not been reported in the literature.
154
3.2 Electrochemical Dechlorination of HCBD
So far, EAg/AgCl readings show that bacterial reduction of cyanocobalamin converted Co3+
to Co2+ which transferred 1 electron to enable HCBD dechlorination. Theoretically, when
Co3+ is converted to Co+ (instead of Co2+), 2 electrons are transferred (instead of 1) for
dechlorination. Thus, the effect of electrochemically increasing the proportion of Cobalt I
(Co+) on the rate of HCBD dechlorination was investigated here. The more the available
electrons from cyanocobalamin reduction, the higher the rate of dechlorination expected.
The HCBD dechlorination rate increased as the EAg/AgCl of the reaction medium was
reduced from -0.3 to -0.8 V (Fig. 6.10). The largest increase was observed from -0.6 V to
-0.8 V. This redox potential range coincides with the increase in proportion of Co+ as the
midpoint (E0) of Co2+/Co+ at 55 °C is -0.8 V. Hence, it is apparent that the greater the
proportion on Co+, the higher the HCBD dechlorination rate. HCBD dechlorination was
not detected in the absence of cyanocobalamin.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
155
0
40
80
120
160
200
-1 -0.8 -0.6 -0.4 -0.2 0Eh (mV vs. Ag/AgCl)
C4 gas production rate
(uM/day)
Figure 6.10 Effect of electrochemically reducing the EAg/AgCl (of the reaction medium) on the formation of C4 gases from HCBD dechlorination.
4 Conclusion
The change in EAg/AgCl can be linked to the dechlorination of HCBD. The peak in
redox potential can be linked to the formation of dechlorination products.
With increasing concentrations of HCBD (0.32 µM to 16 µM) added, the EAg/AgCl
of the anaerobic liquid increased for longer times and to higher values.
EAg/AgCl could be used as an online parameter for an automated feeding regime,
supplying new HCBD on demand and possibly avoiding the build-up of inhibitory
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
156
concentrations of HCBD, and could be used to monitor the dechlorination of other
chlorinated hydrocarbons.
EAg/AgCl affects the rate of HCBD dechlorination.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
157
Chapter 7
Conclusions and Outlook
158
The purpose of this section is to 1) summarise the significant findings and contributions
from the research presented in this thesis, 2) present possible applications of this research
and 3) identify limitations in the dechlorination of HCBD and some suggestions of future
work. The conclusions are described briefly below.
1 Significant Findings in This Research
Non-specific bacteria from activated sludge, ADE, DPW and SRC are able to
dechlorinate HCBD in the presence of cyanocobalamin to chlorine-free C4 gases.
A bacterial consortia specific to reducing cyanocobalamin can be built as a means
of increasing rates of HCBD dechlorination.
Methanogens, traditionally considered to compete with dehalorespiring organisms
for electron donors, were found to be involved in HCBD dechlorination.
Mediators rather than specific bacteria were responsible for the fast dechlorination
rates.
Redox potential can be used to monitor HCBD dechlorination in ADE cultures.
The most significant finding from this research is that it demonstrates completely
dechlorinated end-product from HCBD dechlorination in contrast with other studies in
literature where HCBD dechlorination was equated with disappearance rather than the
detoxification of the primary contaminant (here HCBD).
159
It also shows that, in contract to literature where specific bacteria (i.e., pure
strains/cultures) were used for dechlorination, non-specific bacteria are able to
dechlorinate HCBD, and that biomass cultivated under mesophilic conditions are able to
dechlorinate HCBD at 30 - fold faster rates at thermophilic conditions than under
mesophilic conditions.
2 Potential Applications of This Research
The results obtained from this thesis can be applied in several ways for large scale
bioremediation of HCBD from contaminated soils and groundwater. They are
biostimulation, in-situ and ex-situ treatments.
2.1 Biostimulation
The results in this thesis suggest the use of non-specific bacterial source is adequate for
HCBD dechlorination (Chapter 2 and 4). If so, the bacteria inherent in the HCBD
contaminated site could be stimulated to dechlorinate HCBD by providing the crucial
ingredients, cyanocobalamin and acetate. In this site, HCBD dechlorination can be
monitored via redox potential readings as demonstrated in Figure 6.4. This method of
monitoring remediation is also cost-effective as the cost of transporting expensive
analytical equipment on-site is eliminated. Even though, it may not always be feasible to
provide high temperatures to heat vast amounts of contaminated land, this treatment
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
160
option could still be used for contaminated land especially in areas where low levels of
HCBD contamination exist. HCBD dechlorination has been shown at temperatures below
55 °C (Figure 2.8).
2.2 In-situ Remediation
In this method, the bacterial consortia (Activated Sludge or ADE cultures) are added to a
HCBD contaminated site, along with acetate and cyanocobalamin, as shown in Figure
7.1. The principal difference of this method compared to bioaugmentation is the addition
of bacterial consortia along with the other crucial ingredients of cyanocobalamin and
acetate. Again, this addition can be left to dechlorinate the HCBD in the contaminated
site and dechlorination can be monitored via redox potential readings. If the site is dry,
the site would have to be flooded with water to facilitate reductive dechlorination of
HCBD to remove oxygen trapped between soil particles. It would be possible to
remediate the site by containment using commercially available polytetrafluoroethylene
(PTFE) liners or membranes.
Figure 7.1 In-situ bioremediation of a HCBD-contaminated site.
HCBD-Contaminated Site
Redox Potential
Activated Sludge + Acetate + Cyanocobalamin
161
2.3 In-situ Remediation Using Ex-situ Reduced Cyanocobalamin
In this HCBD remediation application, cyanocobalamin is reduced ex-situ via a
bioreactor (Fig. 7.2). This bioreactor is similar to the reactor described in Chapter 3
(Fig.3.1). The modification in this application involves the use of HCBD instead of
oxygen to re-oxidise cyanocobalamin. The bioreactor contains the bacterial consortia
(Activated Sludge or ADE), cyanocobalamin and acetate.
Figure 7.2 In-situ bioremediation of a HCBD-contaminated site using recycled cyanocobalamin from a cyanocobalamin-reducing bioreactor.
Upon cyanocobalamin reduction, the reduced cyanocobalamin exits the bioreactor onto
the HCBD-contaminated site. As it has been demonstrated in Chapter 2 (Fig. 2.13) that
the reduced cyanocobalamin interacts with HCBD to cause dechlorination, reduced
cyanocobalamin can be added directly to the HCBD-contaminated site for dechlorination
to proceed. Therefore, the reduction of cyanocobalamin can be kept independent (i.e., ex-
situ) of the reaction between reduced cyanocobalamin and HCBD; which will occur in
HCBD-Contaminated Site
Activated Sludge + Acetate + Cyanocobalamin
Redox Potential
Redox Potential
162
the contaminated site. The temperature within the reactor can be controlled in this system
which is a useful feature. It has been shown that higher temperatures increase the rate of
cyanocobalamin reduction (Fig. 6.2) which has been demonstrated as the crucial factor/
rate limiting step in biological dechlorination of HCBD (Chapter 2). Cyanocobalamin can
also be recycled. This recycling feature eliminates the need for a constant supply of fresh
cyanocobalamin which would reduce the operational cost. In addition, HCBD entering
the reactor is also dechlorinated within the reactor. Using this system, it is possible to
remediate high levels of HCBD contamination in soil as HCBD is not directly in contact
with the bacterial consortia (except at low levels in water entering the bioreactor). Hence,
there is a reduced risk of bacterial toxicity. Redox potential can be used to monitor
HCBD dechlorination in the contaminated site and for process control. Hence, this
application is in-situ remediation of HCBD using an ex-situ bioreactor that is primarily
concerned with cyanocobalamin reduction.
2.4 Ex-situ HCBD Remediation
HCBD-contaminated soil can be excavated from a contaminated site and dechlorinated in
an ex-situ bioreactor either on-site or off-site (Fig. 7.3). This set-up will contain Activated
Sludge or ADE with acetate and cyanocobalamin in addition to HCBD contaminated soil
in an upflow re-circulated reactor. The redox potential measurements will serve to
monitor dechlorination within the bioreactor. The advantage to this system is the
treatment of HCBD offsite where the conditions necessary for high dechlorination rates
can be controlled (i.e. 55 °C). Mediators can even be solubilised and concentrated from
the lyophilised form to further increase dechlorination rates (Fig. 5.5). All this makes this
163
bioreactor portable. In addition, a fresh source of Activated sludge bacteria from a
wastewater treatment facility can be obtained and inoculated when and where required.
Figure 7.3 Ex-situ bioremediation of HCBD-contaminated soil.
To estimate the time taken to dechlorinate HCBD, the following calculation was
constructed: Given 3.2 g of HCBD/kg of sediment in CPW (General Introduction), 1000
kg of soil contains 3200 g/ (260.76 g/ mole) = 12.3 moles (12,300,000 µmoles) HCBD.
Based on an average conversion rate of HCBD of 100 µmoles/L culture/day to C4 gases
(Fig. 4.1) (provided this rate is maintained and complete desorption of HCBD is
achieved), it should theoretically take 120, 000 days (328.8 years) to convert 1000 kg of
HCBD-contaminated soil per litre of ADE. Alternatively, if 1000 litres of ADE was used,
it would take 120 days (approximately 4 months).
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
Activated Sludge + Acetate +
Cyanocobalamin + HCBD-contaminated soil excavated from a
contaminated site
Redox Potential
164
Although ex-situ treatment of contaminants is energy-intensive and tends to release
volatile compounds into the atmosphere, this application may be best suited to a
contaminated site where high concentrations of HCBD are found in specific spots within
the site. Hence, excavation could be used to remove the HCBD-contaminated soil from
those spots for remediation either on-site or off-site.
2.5 Ex-situ HCBD Remediation Using Adsorption
One other application may involve the use of soil vapour extraction to remove HCBD
from a contaminated site. Soil vapour extraction involves the removal of contaminant
from soil into either an aqueous, solvent or gas phase (Khan et al., 2004). In this
application, HCBD is removed in the gaseous phase using Nitrogen (N2) gas and bubbled
through a bioreactor containing Activated Sludge, ADE, cyanocobalamin and acetate.
Instead of oxygen, N2 gas is best suited for this application to maintain reducing
conditions within the reactor. Any excess HCBD not dechlorinated within the reactor is
expected to be released out of the bioreactor. In order to trap all HCBD exiting the
bioreactor into the atmosphere, activated carbon cartridges may be installed (Fig. 7.4).
165
Figure 7.4 Soil vapour extraction followed by bioremediation of HCBD at a contaminated site. This application may facilitate the immediate removal of HCBD adsorbed onto cartridges
and HCBD dechlorination off-site. This application may suit a contaminated site where
high concentrations of HCBD are confined to certain spots within the site. Redox
potential measurements within the bioreactor will measure HCBD dechlorination (Fig.
6.4).
The possibility of using activated sludge, acetate and cyanocobalamin to dechlorinate
HCBD adsorbed onto Activated Carbon cartridges needs to be studied. This was beyond
the scope of this thesis. However, the potential benefit to HCBD remediation would
warrant its investigation.
The possibility of solvent extraction of HCBD adsorbed onto Activated Carbon and its
dechlorination using abiotic method was subject of a patent filed by Lee and Cord-
Activated Sludge + Acetate +
Cyanocobalamin
Redox Potential
Redox Potential
Activated Carbon cartridge
N2 gas
Gas Release
HCBD-Contaminated Site
166
Ruwisch (2008). The dechlorination rates exhibited in this patent were approximately 100
- fold faster than the rates of dechlorination reported in this thesis. However, the
successful biological dechlorination of HCBD adsorbed onto Activated Carbon will be an
environmentally friendly alternative for large scale application and may find a niche in
contaminated sites with low-levels of HCBD contamination.
2.6 Ex-situ HCBD Remediation Using Absorption
One reason for the repeated stalling of reactions could be due to absorption of HCBD into
biomass. Given that no HCBD was detected in cultures within a few hours of incubation
(Fig. 5.12) and that the concentration of C4 gases* was lower than the initial
concentration of HCBD added to cultures, it is plausible that the HCBD was absorbed
into bacterial biomass. This absorption will limit free HCBD in solution and may have
rendered HCBD unavailable for dechlorination. This absorption is not necessarily a
limitation for HCBD remediation as it could serve to remove HCBD from contaminated
sites. From the previous application (Fig. 7.4), bacterial biomass could be used in place of
Activated Carbon cartridge to remove HCBD. The HCBD removed/stripped from the soil
can then be transported and subsequently treated off-site using solvent extraction and
abiotic dechlorination (Lee and Cord-Ruwisch, 2008). Even though the dechlorination
off-site would occur at a slower rate compared to the rate at which it was removed, but
because it is removed quickly, contained and treated at a facility in a different location,
the immediate risk of exposure to the community is reduced.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
167
The added advantage of the absorption of HCBD into biomass is that the application is
not only limited to removing HCBD from soil from contaminated sites but it can be
extended to the removal of HCBD contaminated in groundwater. Groundwater
contaminated with HCBD can be removed using aerobic Activated Sludge from
wastewater (Fig. 7.5).
Figure 7.5 Removal of HCBD from contaminated groundwater by absorption into bacterial biomass in wastewater.
The advantage to this treatment option is the immediate removal of HCBD from
contaminated water streams or groundwater bodies. The adsorbed/absorbed HCBD can
then be treated anaerobically using activated sludge or ADE and cyanocobalamin in an
ex-situ bioreactor (Fig. 7.3) at a different location. The effect of HCDB dechlorination by
anaerobic bacteria in digestion supplemented with cyanocobalamin should be tested.
2.7 Other Applications
The use of non-specific bacteria in the presence of a mediator to dechlorinate
contaminants can also be applied elsewhere. Mediators reduced by non-specific bacteria
have been used for electricity production (Bullen et al., 2006; Cheng et al., 2009; He and
Angenent, 2006; Lovley, 2006; Wilkinson et al., 2006).
Bacterial Biomass
Inflow Outflow
168
The use of non-specific bacteria in the presence of a mediator has also been observed to
be involved in the decolourisation from textile wastewater (Dos Santos et al., 2005; Field
and Brady, 2003; Robinson et al., 2001; Szpyrkowicz et al., 2005).
3 Future Work
After a few days of C4 gases production, dechlorination ceases in all tests and attempts to
revive the dechlorination resulted in limited success. Further studies are required to
understand why HCBD dechlorination stalls as it proves to be an important obstacle for
maintaining dechlorination. Results indicate that when a minute amount of HCBD is
added (32 µm instead of 1 mM), a greater amount of HCBD is dechlorinated to C4 gases
(Fig. 6.4). HCBD dechlorination was monitored on-line via ORP readings and these
additions could be made to occur via computer-control.
It is also plausible that HCBD and its dechlorination products become toxic to bacteria
after the first few days of dechlorination. Fresh bacteria from activated sludge or ADE,
which are relatively cheap and abundant in supply, could be introduced into a culture that
has ceased dechlorination in order to renew dechlorination.
Benzene and chlorobenzene are both end products from the anaerobic dehalogenation of
beta-hexachlorocyclohexane (Middeldorp et al., 1996; van Doesburg et al., 2005). In this
example, the mixture of end products is not indicative of a reaction that stalled but rather
one that achieved completion. Similarly, the mixture of both partially dechlorinated
169
intermediates as well the chlorine-free end products from HCBD dechlorination may not
be indicative of a reaction that stalled but rather one that achieved completion.
Mass balance currently indicates that up to of 65 % of HCBD added could lead to
completely dechlorinated products (Fig. 6.4). The remaining 35 % or more could be
trapped as partially dechlorinated intermediates or undegraded HCBD*. Throughout the
thesis, partially dechlorinated intermediates were not measured because standards for
such intermediates are not commercially available. Hence, no meaningful data can be
obtained for the levels of partially dechlorinated intermediates that exist both in the gas
and aqueous phases in cultures. A reliable method of accounting for these partially
dechlorinated intermediates would ensure a measured recovery of all the products formed
to enable a more representative mass balance.
A recirculated anaerobic bioreactor, similar to Figure 7.3, with ADE was incubated with
HCBD contaminated soil from CPW, in the presence of cyanocobalamin and acetate to
study the feasibility of ex-situ application of microcosm tests. HCBD dechlorination was
not confirmed, even after repeated incubations. It was suspected that the high
concentration of HCBD in CPW prevented dechlorination. The set-up of a laboratory-
scale on-line monitored bioreactor, incubated with low concentrations of HCBD, to
demonstrate continuous HCBD dechlorination will be beneficial to study bioremediation
of HCBD.
Overall, the levels of completely dechlorinated gases provided throughout the thesis have been underestimated. To obtain the correct rates, the rate/s given here has/have to be multiplied by a factor of 1.5 to consider the effect of Henry’s law on the solubility of C4 gases. Refer to detailed calculations in Addendum (pages 171 - 172).
170
The use of cost effective mediators may prove useful for large scale bioremediation as the
high cost of cyanocobalamin ($400/ 5 grams) may render bioremediation of HCBD
contaminated sites as an unattractive option. In this thesis, a number of electron
mediators were studied for their efficacy in replacing cyanocobalamin as an electron
mediator for HCBD dechlorination (Table 4.2). This study could be further extended to
test a larger number of electroactive species (of cheaper cost), that either mediate alone or
in combination with other such mediators, to enable higher rates of HCBD dechlorination
than reported in this thesis. Moreover, the extent to which bacterially produced mediators
influence and enhance HCBD dechlorination could also be studied further to better
understand the exact mechanism involved.
Lastly, further studies involving the dechlorination of other commonly occurring
pollutants using non-specific bacteria (i.e., Sludge) in the presence of cyanocobalamin
could be undertaken to expand on the findings from this thesis.
171
Addendum
In the examination phase of the thesis it became apparent that the aqueous fraction of the
C4 gases were ignored in this thesis as C4 hydrocarbons were assumed to be reasonably
non-polar gases. If in fact the water soluble fraction is substantial, some quantitative
aspects in the thesis may need to be revisited. It will also interfere with experimental
findings if the amount released from solution was due to the re-establishing equilibrium
and not from renewed dechlorination, when the headspace is flushed. A detailed
calculation quantifying the errors incurred in typical experiments is given below.
Henry’s Law Error Quantification
Henry’s Law states that the partial pressure of a gas in the headspace is proportional to
the mass of that gas dissolved in solution. The dimensionless Henry’s Law constant
(Hg/Haq) (KH,invcc) for 1,3-butadiyne, 1,3-butadiene and 3-buten-1-yne were converted
from kH/[M/atm] values of 1.9 x 10-1(Yaws and Young, 1992), 1.4 x 10-2 (Yaws and
Young, 1992) and 3.8 x 10-2 (Wilheim et al., 1977) mol/kg*bar respectively using a
conversion factor of 4.088 x 10-2 (Sander, 1999) to result in the dimensionless values of
0.21, 2.92 and 1.08 respectively. These values are the same as given by the reviewer.
The predominant species of C4 gases in typical experiments was 1,3-butadiene which
accounted for approximately 80 % of the total C4 gases. The remaining 20 % was made
up by 1,3-butadiyne and 1-buten-3-yne. By accounting for the ignored C4 gases in the
aqueous phase, at 20 °C, the total C4 gases produced in the thesis were underestimated by
a factor of about 2.2 (Table 1 below).
172
The majority of experiments were conducted at a temperature of 55 °C. Henry’s law
constants for 1,3-butdiene and 3-buten-1-yne were calculated as 11.93, 1.81 and assumed
to be 0.42 for 1,3-butadiyne. At 55 °C, by accounting for the ignored C4 gases in the
aqueous phase, the total C4 gases produced in the thesis were underestimated by a factor
of about 1.5 (Table 1 below).
Table 1 Table showing the approximate total error of C4 gases detected throughout the thesis. 1,3-
butadiene 1,3-
butdiyne 1-buten-3-
yne Total C4 gases produced as detected in the headspace of a typical experiment (µmole per 100 mL vial)
10
Individual C4 gases produced in the headspace of a typical experiment (µmole per 100 mL vial)
8 1.5 0.5
Henry’s gas constant (Hg/Haq) at 20 °C 2.92 0.21 1.08 Ignored C4 gas in the aqueous phase assuming 50 % headspace (µmole per 100 mL vial) at 20 °C
2.74
7.14
0.46
Ignored C4 gas in the 60 % aqueous phase with the experimentally used 40 % headspace (µmole per 100 mL vial) at 20 °C
3.29
8.57
0.56
Corrected C4 gases produced per 100 mL vial at 20 °C
11.29 10.07 1.06
Total corrected C4 gases (ignoring 1-buten-3-yne) at 20 °C
22.44
Henry’s gas constant (Hg/Haq) at 55 °C 11.93 0.42* 1.81 Ignored C4 gas in the aqueous phase assuming 50 % headspace (µmole per 100 mL vial) at 55 °C
0.67
3.57
0.28
Ignored C4 gas in the 60 % aqueous phase with the experimentally used 40 % headspace (µmole per 100 mL vial) at 55 °C
0.8
4.28
0.34
Corrected C4 gases produced per 100 mL vial at 55 °C
8.8
5.78
0.84
Total corrected C4 gases (ignoring 1-buten-3-yne) at 20 °C
15.42
*Henry’s Law constant assumed.
173
References
1. Abramowicz, D.A., Brennan, M.J., Van Dort, H.M. and Gallagher, E.L. (1993) Factors Influencing the Rate of Polychlorinated Biphenyl Dechlorination in Hudson River Sediments. Environmental Science and Technology 27, 1125 - 1131.
2. Ahring, B.K., Christiansen, N., Mathrani, I., Hendriksen, H.V., Macario, A.J.L.
and De Macario, E.C. (1992) Introduction of a De Novo Bioremediation Ability, Aryl Reductive Dechlorination, into Anaerobic Granular Sludge by Inoculation of Sludge with Desulfomonile tiedjei. Applied and Environmental Microbiology 58, 3677 - 3682.
3. Aitken, M.D., Massey, I.J., Chen, T. and Heck, P.E. (1994) Characterization of
Reaction Products from Enzyme Catalyzed Oxidation of Phenolic Pollutants. Water Research 28, 1879 - 1889.
4. Akhtar, S, Khan, A.A. and Husain, Q. (2005a) Potential of Immobilized Bitter
Gourd (Momordica charantia) Peroxidases in the Decolorization and Removal of Textile Dyes from Polluted Wastewater and Dyeing Effluent. Chemosphere 60, 291 - 301.
5. Akhtar, S., Khan, A.A. and Husain, Q. (2005b) Partially Purified Bitter Gourd
(Momordica charantia) Peroxidase Catalyzed Decolorization of Textile and Other Industrially Important Dyes. Bioresource Technology 96, 1804 - 1811.
6. Akhtar, S. and Husain, Q. (2006) Potential Applications of Immobilized Bitter
Gourd (Momordica charantia) Peroxidase in the Removal of Phenols from Polluted Water. Chemosphere 65, 1228 - 1235.
7. Alexander, M. (1985) Biodegradation of Organic Chemicals. Environmental
Science and Technology 18, 106 - 111.
8. Allard, A-S., Hynning, P-Å., Remberger, M. and Neilson, A.H. (1992) Role of Sulfate Concentration in Dechlorination of 3,4,5-Trichlorocatechol by Stable Enrichment Cultures Grown with Coumarin and Flavanone Glycones and Aglycones. Applied and Environmental Microbiology 58, 961 - 968.
9. Alvarez-Cohen, L. and McCarty, P.L. (1991a) Product Toxicity and Cometabolic
Competitive Inhibition Modeling of Chloroform and Trichloroethylene Transformation by Methanotrophic Resting Cells. Applied and Environmental Microbiology 57, 1031 - 1037.
10. Alvarez-Cohen, L. and McCarty, P.L. (1991b) Effects of Toxicity, Aeration, and
Reductant Supply on Trichloroethylene Transformation by a Mixed Methanotrophic Culture. Applied and Environmental Microbiology 57, 228 - 235.
174
11. Anonymous (1992) Chemical Review: Hexachlorobutadiene. Dangerous
Properties Industrial Materials Report 12, 2 - 23.
12. Arnold, W.A. and Roberts, A.L. (2000) Pathways and Kinetics of Chlorinated Ethylene and Chlorinated Acetylene Reaction with Fe (0) Particles. Environmental Science and Technology 34, 1794 - 1805.
13. Assaf-Anid, N., Nies L. and Vogel, T.M. (1992) Reductive Dechlorination of a
Polychlorinated Biphenyl Congener and Hexachlorobenzene by Vitamin B12. Applied and Environmental Microbiology 58(3), 1057 - 1060.
14. Baker, K.H. and Herson, D.S. (1994) Bioremediation. McGraw Hill, New York.
15. Ballapragada, B.S., Stensel, H.D., Puhakka, J.A. and Ferguson, J.F. (1997) Effect
of Hydrogen on Reductive Dechlorination of Chlorinated Ethenes. Environmental Science and Technology 31(6), 1728 - 1734.
16. Basheer, S., Kut, O.M., Prenosil, J.E. and Bourne, J.I. (1992) Kinetics of
Enzymatic Degradation of Cyanide. Biotechnology and Bioengineering 39, 629 - 634.
17. Benabdallah El-Hadj, T., Dosta, J. and Mata-A’lvarez (2007) PCB and AOX
Removal in Mesophilic and Thermophilic Sewage Sludge Digestion. Biochemical Engineering Journal 36, 281 - 287.
18. Berededsamuel, Y., Petersen, J.N. and Skeen, R.S. (1996) Effect of
Perchloroethylene (PCE) On Methane and Acetate Production by a Methanogenic Consortium. Applied Biochemistry and Biotechnology 57-8, 915 - 922.
19. Beunink, J. and Rehm, H-J. (1988) Synchronous Anaerobic and Aerobic
Degradation of DDT by an Immobilized Mixed Culture System. Applied Microbiology and Biotechnology 29(1), 72 - 80.
20. Beurskens, J.E.M., Toussaint, M., De Wolf, J., Van Der Steen, J.M.D., Slot, P.C.,
Commandeur, L.C.M. and Parsons, J.R. (1995) Dehalogenation of Chlorinated Dioxins by Anaerobic Microbial Consortium from Sediment. Environmental Toxicology and Chemistry 14, 939 - 943.
21. Bigger, G.W., Elliot, J.M. and Rickard, T.R. (1976) Estimated Ruminal
Production of Pseudovitamin B12, Factor A and Factor B in Sheep. Journal of Animal Science 43, 1077 - 1081.
22. Blum, D.J.W. and Speece, R.E. (1991) A Database of Chemical Toxicity to
Environmental Bacteria and its Use in Interspecies Comparisons and Correlations. Research Journal of the Water Pollution Control Federation 63(3), 193 - 207.
175
23. Booker, R.S. and Pavlostathis, S.G. (2000) Microbial Reductive Dechlorination of
Hexachloro-1,3-butadiene in a Methanogenic Enrichment Culture. Water Research 34, 4437 - 4445.
24. Bose, P. and Sharma, A. (2002) Role of Iron in Controlling Speciation and
Mobilization of Arsenic in Subsurface Environment. Water Research 36, 4916 - 4926.
25. Bosma, T.N.P., Cottaar, F.H.M., Posthumus, M.A., Teunis, C.J., Van Veldhuizen,
A., Schraa, G. and Zehnder, A.J.B. (1994) Comparison of Reductive Dechlorination of Hexachloro-1,3-butadiene in Rhine Sediment and Model Systems with Hydroxocobalamin. Environmental Science and Technology 28, 1124 - 1128.
26. Bullen, R.A., Arnot, T.C., Lakeman, J.B. and Walsh, F.C. (2006) Biofuel Cells
and Their Development. Biosensors and Bioelectronics 21, 2015 - 2045.
27. Brusseau, M.L., Arnold, R.G., Ela, W. and Field, J. (2001) Overview of Innovative Remediation Approaches for Chlorinated Solvents. Arizona Department of Environmental Quality, p.1 - 63.
28. Bryant, F., Hale, D.D. and Rogers, J.E. (1991) Regiospecific Dechlorination of
Pentachlorophenol by Dichlorophenol-Adapted Microorganisms in Freshwater, Anaerobic Sediment Slurries. Applied and Environmental Microbiology 57, 2293 - 2301.
29. Burris, D.R., Delcomyn, C.A., Smith, M.H. and Roberts, A.L. (1996) Reductive
Dechlorination of Tetrachloroethylene and Trichloroethylene Catalyzed by Vitamin B12 in Homogeneous and Heterogeneous Systems. Environmental Science and Technology 30, 3047 - 3052.
30. Caldwell, D.R., Mark, K., Barton, J.S. and Kelley, J.F. (1973) Sodium and Other
Inorganic Growth Requirements of Bacteroides amylophilus. Journal of Bacteriology 114, 782 - 789.
31. Caldwell, S.R. and Raushell, F.M. (1991) Detoxification of Organic Phosphate
Pesticides Using an Immobilized Phosphotriesterase from Pseudomonas diminuta. Biotechnology and Bioengineering 37, 103 - 109.
32. Carr, C.S. and Hughes, J.B. (1998) Enrichment of High-Rate PCE Dechlorination
and Comparative Study of Lactate, Methanol, and Hydrogen as Electron Donors to Sustain Activity. Environmental Science and Technology 32(12), 1817 - 1824.
33. Chang, B-V., Chen, K-S. and Yuan, S-Y. (1995) Dechlorination of 2,4,6-TCP by
an Anaerobic Mixed Culture. Chemosphere 31(8), 3803 - 3811.
176
34. Chang, B-V., Wu, W-B. and Yuan, S-Y. (1997) Biodegradation of Benzene,
Toluene; and Other Aromatic Compounds by Pseudmonas sp. D8. Chemosphere 35(12) 2807 - 2815.
35. Chang, B-V., Chiang, C-W. and Yuan, S-Y. (1998) Dechlorination of
Pentachlorophenol in Anaerobic Sewage Sludge. Chemosphere 36(3), 537 - 545.
36. Cheng, K.Y., Ho, G. and Cord-Ruwisch, R. (2008) Affinity of Microbial Fuel Cell Biofilm for the Anodic Potential. Environmental Science and Technology 42 (10), 3828 - 3834.
37. Chapsal, J.M., Bourbigot, M.M. and Thomas, D. (1986) Oxidation of Aromatic
Compounds by Haemoglobin. Water Research 20, 709 - 713.
38. Chaudhry, G.R. and Chapalamadugu, S. (1991) Biodegradation of Halogenated Organic Compounds. Microbiology and Molecular Biology Reviews 55(1), 59 - 79.
39. Christian, V., Shrivastava, R., Shukla, D., Modi, H.A. and Vyas, B.R. (2005)
Degradation of Xenobiotic Compounds by Lignin-degrading White-rot Fungi: Enzymology and Mechanisms involved. Indian Journal of Experimental Biology 43, 301 - 312.
40. Cord-Ruwisch, R., James, D.L. and Charles, W. (2009) The Use of Redox
Potential to Monitor Biochemical HCBD Dechlorination. Journal of Biotechnology 142, 151 - 156.
41. Cosio, I.G., Fishero, R.A. and Carroad, P.A. (1982) Bioconversion of Shellfish
Waste: Waste Pretreatment Enzyme Production, Process Design and Economic Analysis. Journal of Food Science 47, 901 - 905.
42. De Bruin, W.P., Kotterman, M.J., Posthumus, M.A., Schraa, G. and Zehnder A.J.
(1992) Complete Biological Reductive Transformation of Tetrachloroethene to Ethane. Applied Environmental Microbiology 58(6), 1996 - 2000.
43. Deweerd, K.A. and Bedard, D.L. (1999) Use of Halogenated Benzoates and Other
Halogenated Aromatic Compounds to Stimulate the Microbial Dechlorination of PCBs. Environmental Science and Technology 33, 2057 - 2063.
44. De Wever, H., Cole, J.R., Fettig, M.R., Hogan, D.A. and Tiedje, J.M. (2000)
Reductive Dehalogenation of Trichloroacetic Acid by Trichlorobacter thiogenes gen. nov., sp. nov. Applied and Environmental Microbiology 66(6), 2297 - 2301.
45. Dietrich, G. and Winter, J. (1990) Anaerobic Degradation of Chlorophenol by an
Enrichment Culture. Applied Microbiology and Biotechnology 34, 253 - 258.
177
46. DiMarco, A.A., Bobik, T.A. and Wolfe, R.S. (1990) Unusual Coenzymes of
Methanogenesis. Annual Review of Biochemistry 59, 355 - 394.
47. 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. Applied and Environmental Microbiology 58, 3622 - 3629.
48. Dolfing, J. (1990) Reductive Dechlorination of 3-chlorobenzoate is Coupled to
ATP Production and Growth in an Anaerobic Bacterium, Strain DCB-1. Archives of Microbiology 153(3), 246 - 266.
49. Dolfing, J. and Beurskens, J.E.M. (1995) The Microbial Logic and Environmental
Significance of Reductive Dehalogenation. Advances in Microbial Ecology 14, 143 - 206.
50. Dolfing, J. and Harrison, B.K. (1993) Redox and Reduction Potentials as
Parameters to Predict the Degradation Pathway of Chlorinated Benzenes in Anaerobic Environments. FEMS Microbiology Ecology 13(1), 23 - 29.
51. Doong, R-A., Chen, T-F. and Chang, W-H. (1996) Effects of Electron Donor and
Microbial Concentration on the Enhanced Dechlorination of Carbon Tetrachloride by Anaerobic Consortia. Applied Microbiology and Biotechnology 46(2), 183 - 186.
52. Dos Santos, A.B., Bisschop, I.A., Cervantes, F.J. and Van Lier, J.B. (2004) Effect
of Different Redox Mediators During Thermophilic Azo Dye Reduction by Anaerobic Granular Sludge and Comparative Study Between Mesophilic (30 °C) and Thermophilic (55 °C) Treatments for Decoulorisation of Textile Wastewater. Chemosphere 55, 1149 - 1157.
53. Dos Santos, A.B., Traverse, J., Cervantes, F.J. and Van Lier, J.B. (2005)
Enhancing the Electron Transfer Capacity and Subsequent Color Removal in Bioreactors by Applying Thermophilic Anaerobic Treatment and Redox Mediators. Biotechnology and Bioengineering 89(1), 42 - 52.
54. DSMZ (1983) German Collection of Microorganisms and Cell Cultures.
Catalogue of strains, 5th edition. Braunschweig: Gesellschaft für Biotechnologische Forschung.
55. Durham, R.W. and Oliver, B.G. (1983) History of Lake Ontario Contamination
from the Niagara River by Sediment Radiodating and Chlorinated Hydrocarbon Analysis. Great Lakes Research 9, 160 - 168.
178
56. Duran, N. and Esposito, E. (2000) Potential Applications of Oxidative Enzymes and Phenoloxidase-like Compounds in Wastewater and Soil Treatment: A Review. Applied Catalysis B: Environmental 28, 83 - 99.
57. Duran, N., Rosa, M. A., D’ Annibale, A. and Gianfreda, L. (2002) Applications of
Laccases and Tyrosinases (Phenoloxidases) Immobilized on Different Supports: A Review. Enzyme and Microbial Technology 31, 907 - 931.
58. El Fantroussi, S., Naveau, H. and Agathos, S.N. (1998) Anaerobic Dechlorinating
Bacteria. Biotechnology Progress 14, 167 - 188.
59. Ensley, B.D., (1991) Biochemical Diversity of Trichloroethylene Metabolism. Annual Review of Microbiology. 45, 283 - 299.
60. Fan, S. and Scow, K.M. (1993) Biodegradation of Trichloroethylene and Toluene
by Indigenous Microbial Populations in Soil. Applied and Environmental Microbiology 59, 1911 - 1918.
61. Farwell, S.O., Beland, F.A. and Geer, R.D. (1975) Reduction Pathways of
Organohalogen Compounds: Part I. Chlorinated Benzenes. Journal of Electroanalytical Chemistry 61, 303 - 313.
62. Fathepure, B.Z., Nengu, J.P. and Boyd, S.A. (1987) Anaerobic Bacteria That
Dechlorinate Perchloroethene. Applied and Environmental Microbiology 53(11), 2671 - 2674.
63. Fathepure, B.Z. and Boyd, S.A. (1988) Dependence of Tetrachloroethylene
Dechlorination on Methanogenic Substrate Consumption by Methanosarcina sp. Strain DCM. Applied and Environmental Microbiology 54(12), 2976 - 2980.
64. Fathepure, B.Z., Tiedje, J.M. and Boyd, S.A. (1988) Reductive Dechlorination of
Hexachlorobenzene to Tri- and Dichlorobenzenes in Anaerobic Sewage Sludge. Applied Environmental Microbiology 54, 327 - 330.
65. Fattore, E., Benfenati, E. and Fanelli, R. (1996) Analysis of Chlorinated 1,3-
butadienes by Solid-Phase Microexcitation and Gas Chromatography-Mass Spectrometry. Journal of Chromatography 737, 85 - 91.
66. Fetzner, S. and Lingens, F. (1994) Bacterial Dehalogenases: Biochemistry,
Genetics, and Biotechnological Applications. Microbiological Reviews 58, 641 - 685.
67. Fetzner, S. (1998) Bacterial Dehalogenation. Applied Microbiology and
Biotechnology 50, 633 - 657.
179
68. Field, J.A. and Brady, J. (2003) Riboflavin as a Redox Mediator Accelerating the Reduction of the Azo Dye Mordant Yellow 10 by Anaerobic Granular Sludge. Water and Science and Technology 48(6), 187 - 193.
69. Field, J.A. and Cervantes, F.J. (2005) Microbial Redox Reactions Mediated by
Humus and Structurally Related Quinones. In Use of Humic Substances to Remediate Polluted Environments: From Theory to Practice 52, 343 - 352.
70. FRTR (1999) Thermal Desorption. Federal Remediation Technologies
Roundtable. USEPA, 401 M Street, S.W., Washington, DC, http://www.frtr.gov/matrix2 /section4/4_29.html.
71. Furukawa, K., Suyama, A., Tsuboi, Y., Futagami, T. and Goto, M. (2005)
Biochemical and Molecular Characterization of a Tetrachloroethene Dechlorinating Desulfitobacterium sp. strain Y51: A Review. Journal of Industrial Microbiology and Biotechnology 32, 534 - 541.
72. Gantzer, C.J. and Wackett, L.P. (1991) Reductive Dechlorination Catalyzed by
Bacterial Transition-Metal Coenzymes. Environmental Science and Technology 25, 715 - 722.
73. Gerritse, J., Renard, V., Gomes, T.M.P., Lawson, P.A., Collins, M.D. and
Gottschal J.C. (1996) Desulfitobacterium sp. Strain PCE1, an Anaerobic Bacterium that can Grow by Reductive Dechlorination of Tetrachloroethene or Ortho-chlorinated Phenols. Archives of Microbiology 165(2), 132 - 140.
74. Gibson, S.A. and Sewell, G.W. (1992) Stimulation of Reductive Dechlorination
of Tetrachloroethene in Anaerobic Aquifer Microcosms by Addition of Short-Chain Organic Acids or Alcohols. Applied and Environmental Microbiology 58(4), 1392 - 1393.
75. Glod, G., Angst, W., Holliger, C. and Schwarzenbach, R.P. (1997) Corrinoid-
Mediated Reduction of Tetrachloroethene, Trichloroethene, and Trichlorofluoroethene in Homogeneous Aqueous Solution: Reaction Kinetics and Reaction Mechanisms. Environmental Science and Technology 31, 253 - 260.
76. Guerin, W.F. and Boyd, S.A. (1992) Differential Bioavailability of Soil-Sorbed
Naphthalene to Two Bacterial Species. Applied and Environmental Microbiology 58, 1142 - 1152.
77. Guerrero-Barajas, C. and Field, J.A. (2005) Enhancement of Anaerobic Carbon
Tetrachloride Biotransformation in Methanogenic Sludge with Redox Active Vitamins. Biodegradation 16, 215 - 228.
180
78. Guerrero-Barajas, C. and Field, J.A. (2006) Enhanced Anaerobic Biotransformation of Carbon Tetrachloride with Precursors of Vitamin B12 Biosynthesis. Biodegradation 17(4), 317 - 329.
79. Häggblom, M.M., Apajalahti, J.H.A. and Sallinoja-Salonen, M.S. (1988)
Hydroxylation and Dechlorination of Chlorinated Guaiacols and Syringols by Rhodococcus chlorophenolicus. Applied and Environmental Microbiology, 54(3), 683 - 687.
80. Hashsham, S.A. and Freedman, D.L. (1997) Enhanced Biotransformation of
Carbon Tetrachloride by an Anaerobic Enrichment Culture. In Proceedings of the Fourth International In Situ and On-Site Bioremediation Symposium. New Orleans, L.A., 28 April to 1 May 1997. p. 465 - 470. Battelle Press, Columbus, Ohio.
81. Hashsham, S.A. and Freedman, D.L. (1999) Enhanced Biotransformation of
Carbon Tetrachloride by Acetobacterium woodii Upon Addition of Hydroxocobalamin and Fructose. Applied and Environmental Microbiology 65, 4537 - 4542.
82. He, Z. and Angenent, L.T. (2006) Application of Bacterial Biocathodes in
Microbial Fuel Cells. Electroanalysis 18(19 - 20), 2009 - 2015.
83. He, J., Sung, Y., Dollhopf, M.E., Fathepure, B.Z., Tiedje, J.M. and Löffler, F.E. (2002) Acetate versus Hydrogen as Direct Electron Donors To Stimulate the Microbial Reductive Dechlorination Process at Chloroethene-Contaminated Sites. Environmental Science and Technology 36(18), 3945 - 3952.
84. Heimann, A.C., Batstone, D.J. and Jakobsen, R. (2006) Methanosarcina sp. Drive
Vinyl Chloride Dechlorination via Interspecies Hydrogen Transfer. Applied and Environmental Microbiology 72, 2942 - 2949.
85. Hernandez, M.E. and Newman, D.K. (2001) Extracellular Electron Transfer.
Cellular and Molecular Life Sciences 58(11), 1562 - 1571.
86. Holliger, C., Schraa, G., Stams, A.J. and Zehnder, A.J. (1992a) Enrichment and Properties of an Anaerobic Mixed Culture Reductively Dechlorinating 1,2,3-trichlorobenzene to 1,3-dichlorobenzene. Applied and Environmental Microbiology 58(5), 1636 - 1644.
87. Holliger, C., Kengen, S.W., Schraa, G., Stams, A.J. and Zehnder, A.J. (1992b)
Methyl-Coenzyme M Reductase of Methanobacterium thermoautotrophicum Delta H Catalyzes the Reductive Dechlorination of 1,2-dichloroethane to Ethylene and Chloroethane. Journal of Bacteriology 174, 4435 - 4443.
181
88. Holliger, C., Schraa, G., Stams, A.J. and Zehnder, A.J. (1993) A Highly Purified Enrichment Culture Couples the Reductive Dechlorination of Tetrachloroethene to Growth. Applied and Environmental Microbiology 59(9), 2991 - 2997.
89. Holliger, C. and Schraa, G. (1994) Physiological Meaning and Potential for
Application of Reductive Dechlorination by Anaerobic Bacteria. FEMS Microbiology Reviews 15, 297 - 305.
90. Holliger, C. and Schumacher, W. (1994) Reductive Dehalogenation as a
Respiratory Process. Antoine van Leeuwenhoek 66, 247 - 270.
91. Husain, Q. (2006) Potential Applications of the Oxidoreductive Enzymes in the Decolorization and Detoxification of Textile and Other Synthetic Dyes from Polluted Water: A Review. Critical Reviews in Biotechnology 26, 201 - 221.
92. Husain, M. and Husain, Q. (2008) Application of Redox Mediators in the
Treatment of Organic Pollutants by Using Oxidoreductive Enzymes: A Review. Critical Reviews in Environmental Science and Technology 38, 1 - 42.
93. IPCS (1994) Hexachlorobutadiene. Geneva, World Health Organization,
International Programme on Chemical Safety (Environmental Health Criteria 156).
94. Jablonski, P.E. and Ferry, J.G. (1992) Reductive Dechlorination of
Trichloroethylene by the Co-reduced CO Dehydrogenase Enzyme Complex from Methanosarcina thermophila. FEMS Microbiology Letters 96, 55 - 60.
95. James, D.L., Cord-Ruwisch, R., Schleheck, D., Lee, M.J. and Manefield, M.
(2008) Cyanocobalamin Enables Activated Sludge Bacteria to Dechlorinate Hexachloro-1,3-butadiene to Non-Chlorinated Gases. Bioremediation Journal 12, 177 - 184.
96. Jayachandran, G., Gorish, H. and Adrian, L. (2003) Dehalorespiration with
Hexachlorobenzene and Pentachlorobenzene by Dehalococcoides sp. Strain CBDB1. Archives of Microbiology 180(6), 411 - 416.
97. Jeganathan, J., Bassi, A. and Nakhla, G. (2006) Pre-treatment of High Oil and
Grease Pet Food Industrial Wastewaters Using Immobilized Lipase Hydrolyzation. Journal of Hazardous Materials 137(1), 121 - 128.
98. Johnson, D.R., Park, J., Kukor, J.J. and Abriola, L.M. (2006) Effect of Carbon
Starvation on Toluene Degradation Activity by Toluene Monooxygenase-expressing Bacteria. Biodegradation 17(5), 437 - 445.
99. Jones, E.J.P., Voytek, M.A., Lorah, M.M. and Kirshtein, J.D. (2006)
Characterization of a Microbial Consortium Capable of Rapid and Simultaneous
182
Dechlorination of 1,1,2,2-Tetrachloroethane and Chlorinated Ethane and Ethene Intermediates. Bioremediation Journal 10, 153 - 168.
100. Juteau, P., Beaudet, R., McSween, G., Lépine, F. and Bisaillon, JG. (1995)
Study of the Reductive Dechlorination of Pentachlorophenol by a Methanogenic Consortium. Canadian Journal of Microbiology 41(10), 862 - 868.
101. Kaden, J., Galushko, A.S. and Schink, B. (2002) Cysteine-Mediated
Electron Transfer in Syntrophic Acetate Oxidation by Cocultures of Geobacter Sulfurreducens and Wolinella succinogenes. Archives of Microbiology 178, 53 - 58.
102. Karam, J. and Nicell, J.A. (1997) Potential Applications of Enzymes in
Waste Treatment. Journal of Chemical Technology and Biotechnology 69, 141 - 147.
103. Kargina, O., MacDougall, B., Kargin, Y.M. and Wang, L. (1997)
Dechlorination of Monochlorobenzene Using Organic Mediators. Journal of the Electrochemical Society 144(11), 3715 - 3721.
104. Kataky, R. and Wylie, L.A. (2001) Investigation of Mechanisms for the
Reductive Dechlorination of Chlorinated Ethylenes Using Electroanalytical Techniques. Analyst 126, 1901 - 1906.
105. Kazumi, J., Häggblom, M.M. and Young, L.Y. (1995) Degradation of
Monochlorinated and Nonchlorinated Compounds Under Iron-Reducing Conditions. Applied and Environmental Microbiology 61, 4069 - 4073.
106. Kengen, S.W.M., Breidenbach, C.G., Felske, A., Stams, A.J.M., Schraa,
G. and De Vos, W.M. (1999) Reductive Dechlorination of Tetrachloroethene to cis-1,2-Dichloroethene by a Thermophilic Anaerobic Enrichment Culture. Applied and Environmental Microbiology 65(6), 2312 - 2316.
107. Khan, A.A. and Husain, Q. (2007) Potential for Plant Polyphenol Oxidase
in the Decolorization and Removal of Textile and Non-textile Dyes. Journal of Environmental Science 19, 396 - 402.
108. Khan, F.I., Husain, T. and Hejazi, R. (2004) An Overview and Analysis of
Site Remediation Technologies. Journal of Environmental Management 71, 95 - 122.
109. Khleifat, K.M. (2006) Biodegradation of Linear Alkylbenzene Sulfonate
by a Two Member Facultative Anaerobic Bacterial Consortium. Enzyme and Microbial Technology 39(5), 1030 - 1035.
183
110. Kim, J. and Rhee, G. (1997) Population Dynamics of Polychlorinated Biphenyl-Dechlorinating Microorganisms in Contaminated Sediments. Applied and Environmental Microbiology 63(5), 1771 - 1776.
111. Kim, J. and Rhee, G-Y. (1998) Reductive Dechlorination of
Polychlorinated Biphenyls: Interactions of Dechlorinating Microorganisms with Methanogens and Sulfate Reducers. Environmental Toxicology and Chemistry p. 2696 - 2702.
112. Kim, Y-H. and Carraway, R. (2002) Reductive Dechlorination of PCE and
TCE by Vitamin B12 and ZVMs. Environmental Technology 23(10), 1135 - 1145.
113. Kohring, G.W., Rogers, J.E. and Wiegel J. (1989) Anaerobic
Biodegradation of 2,4-dichlorophenol in Freshwater Lake Sediments at Different Temperatures. Applied and Environmental Microbiology 55(2), 348 - 353.
114. Kuhn, E.P. and Suflita, J.M. (1989) Dehalogenation of Pesticides by
Anaerobic Microorganisms in Soils and Groundwater - A Review. In Reactions and Movement of Organic Chemicals in Soils; America Special Publication No.22. Soil Science Society of America and American Society of Agronomy: Madison, WI, p. 111 - 180.
115. Kulshrestha, Y. and Husain, Q. (2007) Decolorization and Degradation of
Acid Dyes Mediated by Partially Purified Turnip (Brassica rapa) Peroxidases. Toxicological and Environmental Chemistry 89(2), 255 - 267.
116. Larsen, S., Hendriksen, H.V. and Ahring, B.K. (1991) Potential for
Thermophilic (50 degrees C) Anaerobic Dechlorination of Pentachlorophenol in Different Ecosystems. Applied and Environmental Microbiology 57, 2085 - 2090.
117. Lee, C.H., Lewis, T.A., Paszczynski, A. and Crawford, R.L. (1999)
Identification of an Extracellular Catalyst of Carbon Tetrachloride Dehalogenation from Pseudomonas stutzeri Strain KC as Pyridine-2,6-bis(thiocarboxylate). Biochemical and Biophysical Research Communications 261, 562 - 566.
118. Lee, M.J. and Cord-Ruwisch, R. (2008) A Process for the Capture and
Dehalogenation of Halogenated Hydrocarbons. WIPO Patent Application WO/2008/064427. World International Property Office.
119. Lei, Y., Mulchandani, A. and Chen, W. (2005) Improved Degradation of
Organophosphorus Nerve Agents and p-nitorophenol by Pseudomonas putida JS444 with Surface Expressed Organophosphorus Hydrolase. Biotechnology Progress 21(3), 678 - 681.
184
120. Lexa, D. and Saveant, J.M. (1983) The Electrochemistry of Vitamin B12. Accounts of Chemical Research 16(7), 235 - 243.
121. Li, R.T., Going, J.E. and Spigarelli, J.L. (1976) Government Reports and
Announcements Index (US) p. 76, 79.
122. Li, M.T., Hao, L.L., Sheng, L.X. and Xu, J.B. (2008) Identification and Degradation Characterization of Hexachlorobutadiene Degrading Strain Serratia marcescens HL1. Bioresource Technology 99, 6878 - 6884.
123. Löffler, F.E., Ritalahti, K.M. and Tiedje, J.M. (1997) Dechlorination of
Chloroethenes is Inhibited by 2-bromo-ethane sulfonate in the Absence of Methanogens. Applied and Environmental Microbiology 63, 4982 - 4985.
124. Lovley, D.R., Holmes, D.E. and Nevin, K.P. (2004) Dissimilatory Fe(III)
and Mn(IV) reduction. Advances in Microbial Physiology 49, 219 - 286.
125. Lovley, D.R. (2006) Bug Juice: Harvesting Electricity with Microorganisms. Nature Reviews Microbiology 4, 497 - 508.
126. Low, A., Schleheck, D., Khou, M., Aagaard V., Lee M. and Manefield, M.
(2007) Options for In Situ Remediation of Soil Contaminated with a Mixture of Perchlorinated Compounds. Bioremediation 11(3), 113 - 124.
127. Lowe, S.E., Jain, M.K. and Zeikus, J.G. (1993) Biology, Ecology, and
Biotechnological Applications of Anaerobic Bacteria Adapted to Environmental Stresses in Temperature, pH, Salinity, or Substrates. Microbiological Reviews. 57, 451 - 509.
128. Madsen, T. and Licht, D. (1992) Isolation and Characterization of an
Anaerobic Chlorophenol-Transforming Bacterium. Applied and Environmental Microbiology 58(9), 2874 - 2878.
129. Mackay, D., Shiu, W.Y. and Ma, K.C. (1993) Volatile Organic
Compounds. Illustrated Handbook of Physical-Chemical Properties and Environmental Fate for Organic Chemicals 3, 536 - 537. Lewis Publishers, Chelsea, MI, USA.
130. Maloney, S.E., Marks, T.S. and Sharp, R.J. (1997) Degradation of 3-
chlorobenzoate by Thermophilic Microorganisms. Letters in Applied Microbiology 24, 441 - 444.
131. Marchant, R., Banat, I.M., Rahman, T.J. and Berzano, M. (2002) The
Frequency and Characteristics of Highly Thermophilic Bacteria in Cool Soil Environments. Environmental Microbiology 4(10), 595 - 602.
185
132. Masscheleyn, P.H., Delaune, R.D. and Patrick, Jr. W.H. (1991) Effect of Redox Potential and pH on Arsenic Speciation and Solubility in a Contaminated Soil. Environmental Science and Technology 25(8), 1414 - 1419.
133. Master, E.R., Lai, V.W-M., Kuipers, B., Cullen, W.R. and Mohn, W.W.
(2002) Sequential Anaerobic-aerobic Treatment of Soil Contaminated with Weathered Aroclor 1260. Environmental Science and Technology 36, 100 -103.
134. Matto, M. and Husain, Q. (2007) Decolorization of Direct Dyes by Salt
Fractioned Turnip Proteins in the Presence of Hydrogen Peroxide and Redox Mediators. Chemosphere 69(2), 338 - 345.
135. McCarthy, P.L. (1997) Breathing with Chlorinated Solvents. Science 276,
1521 - 1522.
136. McFall-Ngai, M.J. (1999) Consequences of Evolving with Bacterial Symbionts: Insights from the Squid-Vibrio Associations. Annual Review of Ecological Systems 30, 235 - 256.
137. Mensee, A.H., Chen, W. and Mulchandani, A. (2005) Detoxification of
the Organophosphate Nerve Agent Coumaphos Using Organophosphorus Hydrolase Immobilised on Cellulose Materials. Journal of Industrial Microbiology and Biotechnology 32(11-12), 554 - 560.
138. Messmer, M., Wohlfarth, G. and Diekert, G. (1993) Methyl Chloride
Metabolism of the Strictly Anaerobic Methyl-Utilizing Homoacetogen Strain MC. Archives of Microbiology 165, 18 -25.
139. Middeldorp, P.J.M., Jaspers, M., Zehnder, A.J.B. and Schraa, G. (1996)
Biotransformation of α-, β-, γ-, and δ-hexachlorocyclohexane Under Methanogenic Conditions. Environmental Science and Technology, 30 (7), 2345 - 2349.
140. Middeldorp, P.J.M., De Wolf, J., Zehnder, A.J.B. and Schraa, G. (1997)
Enrichment and Properties of a 1,2,4-trichlorobenzene Dechlorinating Methanogenic Microbial Consortium. Applied and Environmental Microbiology 63, 1225 - 1229.
141. Mitchell, P. and Potter, C. (1999) The Future of Waste Treatment in the
Mining Industry, Proceedings of Global Symposium on Recycling, Waste Treatment and Clean Technology, Sept 5 - 9, 1999, San Sebastian, Spain.
142. Miller, E., Wohlfarth, G. and Diekert, G. (1997) Comparative Studies on
Tetrachloroethene Reductive Dechlorination Mediated by Desulfitobacterium sp. Strain PCE-S. Archives of Microbiology 168, 513 - 519.
186
143. Mohn, W.W. and Tiedje, J.M. (1992) Microbial Reductive Dehalogenation. Microbiological Reviews 56, 482 - 507.
144. Morel, F.M.M. and Hering, J.G. (1993) Principles and Applications of
Aquatic Chemistry, Wiley, New York.
145. Morris, P.J., Mohn, W.W., Quensen 3rd, J.F., Tiedje, J.M. and Boyd, S.A. (1992) Establishment of Polychlorinated Biphenyl-Degrading Enrichment Culture with Predominantly Meta Dechlorination. Applied and Environmental Microbiology 58(9), 3088 - 3094.
146. Murray, H.E. and Beck, J.N. (1989) Halogenated Organic Compounds
Found in Shrimp from the Calcasieu Estuary. Chemosphere 19, 1367 - 1374.
147. Nedunuri, K.V., Govindaraju, R.S., Banks, M.K., Schwab, A.P. and Chen, Z. (2000) Evaluation of Phytoremediation for Field-Scale Degradation of Total Petroleum Hydrocarbons. Journal of Environmental Engineering 126 (6), 483 - 490.
148. Neumann, A., Wohlfarth, G. and Diekert G. (1996) Purification and
Characterization of Tetrachloroethene Reductive Dehalogenase from Dehalospirillum multivorans. The American Society for Biochemistry and Molecular Biology 271(28), 16515 - 16519.
149. Newman, D.K. and Kolter, R. (2000) A Role for Excreted Quinones in
Extracellular Electron Transfer. Nature 405, 94 - 97.
150. Nicholson, S.K. and John, P. (2005) The Mechanism of Bacterial Indigo Reduction. Applied Microbiology and Biotechnology 68(1), 117 - 123.
151. Nies, L. and Vogel, T.M. (1990) Effects of Organic Substrates on
Dechlorination of Aroclor 1242 in Anaerobic Sediments. Applied and Environmental Microbiology 56, 2612 - 2617.
152. Olivas, Y., Dolfing, J. and Smith, G.B. (2002) The Influence of Redox
Potential on the Degradation of Halogenated Methanes. Environmental Toxicology and Chemistry 21(3), 493 - 499.
153. Page, W.J. and Knosp, O. (1991) Mutant strains of Azotobacter vinelandii
Used for the Hyperproduction of Poly-β-Hydroxybutyrate During Exponential Growth. US Patent No. 5059536.
154. Pandolfi, D., Pons, M-N. and Motta, M. (2007) Characterization of PHB
Storage in Activated Sludge Extended Filamentous Bacteria by Automated Colour Image Analysis. Biotechnology Letters 29(8), 1263 - 1269.
187
155. Pardue, J.H., Delaune, R.D. and Patrick, Jr. W.H. (1988) Effect of Sediment pH and Oxidation-Reduction Potential on PCB Mineralization. Water, Air and Soil Pollution 37, 439 - 447.
156. Pavlostathis, S.G., Prytula, M.T. and Yeh, D.H. (2003) Potential and
Limitations of Microbial Reductive Dechlorination for Bioremediation Application. Water, Air and Soil Pollution: Focus 3, 117 - 129.
157. Pereira, W.E., Rostand, C.E., Chiou, C.T., Brinton, T.I., Barber, L.B.,
Demchek, D.K. and Demas, C.R. (1988) Contamination of Estuarine Water, Biota and Sediment by Halogenated Organic Compounds: A Field Study. Environmental Science and Technology 22, 772 - 778.
158. Petersen, S.P. and Ahring, B.K. (2006) Acetate Oxidation in a
Thermophilic Anaerobic Sewage-sludge Digestor: The Importance of Non-Aceticlastic Methanogenesis from Acetate. FEMS Microbiology Letters 86(2), 149 - 158.
159. Qiu, X. and Davis, J.W. (2004) Environmental Bioavailability of
Hydrophobic Organochlorines in Sediments - A Review. Remediation 14(2), 55 - 84.
160. Quensen 3rd, J.F., Tiedje, J.M. and Boyd, S.A. (1988) Reductive
Dechlorination of Polychlorinated Biphenyls by Anaerobic Microorganisms from Sediments. Science 242, 752 - 754.
161. Rabaey, K., Boon, N., Siciliano, S.D., Verhaege, M. and Verstraete, W.
(2004) Biofuel Cells Select for Microbial Consortia That Self-Mediate Electron Transfer. Applied and Environmental Microbiology 70(9), 5373 - 5382.
162. Rabaey, K., Boon, N., Höfte, M. and Verstraete, W. (2005) Microbial
Phenazine Production Enhances Electron Transfer in Biofuel Cells. Environmental Science and Technology 39, 3401 - 3408.
163. Reddy, K.R., Admas, J.F. and Richardson, C. (1999) Potential
Technologies for Remediation of Brownfield. Practice Periodical of Hazardous, Toxic, and Radioactive Waste Management 3(2), 61 - 68.
164. Riser-Roberts, E. (1998) Remediation of Petroleum Contaminated Soil:
Biological, Physical, and Chemical Processes, Lewis Publishers, BocaRaton, FL.
165. Robinson T., McMullan G., Marchant R. and Nigam P. (2001) Remediation of Dyes in Textile Effluent: A Critical Review on Current Treatment Technologies with a Proposed Alternative. Bioresource Technology 77(3), 247 - 255.
188
166. Romantschuk, M., Sarand, I., Petänen, T., Peltola, R., Jonsson-Vihanne, M., Koivula, T., Yrjälä, K. and Haahtela, K. (2000) Means to Improve the Effect of In-Situ Bioremediation of Contaminated Soil: An Overview of Novel Approaches. Environmental Pollution 107, 179 - 185.
167. Rostad, C.E., Pereira, W.E. and Leiker, T.J. (1989) Distribution and
Transport of Selected Anthropogenic Lipophilic Organic Compounds Associated with Mississippi River Suspended Sediment. Archives of Environmental Contamination and Toxicology 36(3), 248 - 255.
168. Sahm, H., Brunner, M. and Schoberth, S.M. (1986) Anaerobic
Degradation of Halogenated Aromatic Compounds. Microbial Ecology 12, 147 - 153.
169. Sander, R. (1999) Compilation of Henry's Law Constants for Inorganic
and Organic Species of Potential Importance in Environmental Chemistry (Version 3) http://www.henrys-law.org (Accessed 10-01-2010).
170. Schmidt, H., Akkermans, A.D.L., Van Der Oost, J. and De Vos, W.M.
(2000) Halorespiring Bacteria-Molecular Characterization and Detection. Enzyme and Microbial Technology 27(10), 812 - 820.
171. Schumacher, W., Holliger, C., Zehnder, A.J.B. and Hagen, W.R. (1997)
Redox Chemistry of Cobalamin and Iron-Sulfur Cofactors in the Tetrachloroethene Reductase of Dehalobacter restrictus. FEBS Letters 409(3), 421 - 425.
172. Seedher, N. and Bhatia, S. (2003) Solubility Enhancement of Cox-2
Inhibitors Using Various Solvent Systems. AAPS PharmSciTech 4(3), 36 - 44.
173. Shey, J.A. and Van Der Donk, W.A. (2000) Mechanistic Studies on the Vitamin B12-Catalyzed Dechlorination of Chlorinated Alkenes. Journal of the American Chemical Society 122, 12403 - 12404.
174. Shi, J., Nawaz, H., Pohorly, J., Mittal, G., Kakuda, Y. and Jiang, Y. (2005)
Extraction of Polyphenolics from Plant Material for Functional Foods - Engineering and Technology. Food Reviews International 21, 139 - 166.
175. Shimazu, M., Mulchandani, A. and Chen, W. (2001) Simultaneous
Degradation of Organophosphorus Pesticides and p-nitrophenol by a Genetically Engineered Moraxella sp. with Surface-expressed. Biotechnology and Bioengineering 76(4), 318 - 324.
176. Shimomura, T. and Sanford, R.A. (2005) Reductive Dechlorination of
Tetrachloroethene in a Sand Reactor Using a Potentiostat. Journal of Environmental Quality 34, 1435 - 1438.
189
177. Shulder, S.J. (2006) Experience and Use of Oxidation Reduction Potential
Measurements in Power Plant Applications. PowerPlant Chemistry 9(3).
178. Singh, P.K., Schaefer, A.L., Parsek, M.R., Moninger, T.O., Welsh, M.J. and Greenberg, E.P. (2000) Quorum-Sensing Signals Indicate That Cystic Fibrosis Lungs Are Infected with Bacterial Biofilms. Nature 407, 762 - 764.
179. Singleton, P. (2004) Bacteria in Biology, Biotechnology and Medicine.
Sixth edition. John Wiley and Sons. p. 15. (ISBN 0470090278).
180. Sparling, R. and Daniels, L. (1987) The Specificity of Growth Inhibition of Methanogenic Bacteria by Bromoethanesulfonate. Canadian Journal of Microbiology 33, 1132 - 1136.
181. Stuart, S.L., Woods, S.L., Lemmon, T.L. and Ingle Jr, J.D. (1999) The
Effect of Redox Potential Changes on Reductive Dechlorination of Pentachlorophenol and the Degradation of Acetate by a Mixed, Methanogenic Culture. Biotechnology and Bioengineering 63(1), 69 - 78.
182. Suflita, J.M., Horwitz, A., Shelton, D.R. and Tiedje, J.M. (1982)
Dehalogenation: A Novel Pathway for the Anaerobic Biodegradation of Haloaromatic Compounds. Science 218, 1115 - 1116.
183. Sung, Y., Ritalahti, K.M., Apkarian, R.P. and Löffler, F.E. (2006)
Quantitative PCR Confirms Purity of Strain GT, a Novel Trichloroethene-to-Ethene-Respiring Dehalococcoides Isolate. Applied and Environmental Microbiology 72, 1980 - 1987.
184. Szpyrkowicz, L., Kaul, S.N. and Neti R.N. (2005) Tannery Wastewater
Treatment by Electro-oxidation Coupled with a Biological Process. Journal of Applied Electrochemistry 35(4), 381 - 390.
185. Tandoi, V., Distefano, T.D., Bowser, P.A., Gossett, J.M. and Zinder, S.H.
(1994) Reductive Dechlorination of Chlorinated Ethanes and Halogenated Ethanes by a High-Rate Anaerobic Enrichment Culture. Environmental Science and Technology 28, 973 - 979.
186. Third, K.A., Cord-Ruwisch, R. and Watling, H.R. (2002) Control of
Redox Potential by Oxygen Limitation Improves Bacterial Leaching of Chalcopyrite. Biotechnology and Bioengineering 78(4), 433 - 441.
187. Thomas, L., Jungschaffer, G. and Sproessler, B. (1993) Improved Sludge
Dewatering by Enzymatic Treatment. Water Science and Technology 28, 189 - 192.
190
188. Thomas, O.R.T. and White, G.F. (1991) Immobilization of the Surfactant-Degrading Bacterium Pseudomonas C12B in Polyacrylamide Gel. III. Biodegradation Specificity for Raw Surfactant and Industrial Wastes. Enzyme and Microbial Technology 13, 338 - 343.
189. Truex, M., Powell, T. and Lynch, K. (2007) In Situ Dechlorination of TCE
during Aquifer Heating. Ground Water Monitoring and Remediation 27, 96 - 105.
190. Van Der Honing, M. (2007) Exploration of Management Options for Hexachlorobutadiene (HCBD). Paper for the 6th meeting of the UNECE CLRTAP Task Force on Persistent Organic Pollutants, Vienna, 4 - 6 June 2007, p. 1 - 22.
191. Van Der Zee, F.P., Lettinga, G. and Field, J.A. (2001) Azo Dye
Decolourisation by Anaerobic Granular Sludge. Chemosphere 44, 1169 - 1176.
192. Van Doesburg, W., Van Eekert, M.H.A., Middeldorp, P.J., Balk, M., Schraa, G. and Stams, A.J. (2001) Reductive Dechlorination of beta-hexachlorocyclohexane (beta-HCH) by a Dehalobacter Species in Coculture with Sedimentibacter sp. 54 (1), 87 - 95.
193. Van Eekert, M.H.A. and Schraa, G. (2001) The Potential of Anaerobic
Bacteria to Degrade Chlorinated Compounds. Water Science and Technology 44(8), 49 - 56.
194. Van Pée, K-H. and Unversucht, S. (2003) Biological Dehalogenation and
Halogenation Reactions. Chemosphere 52, 299 - 312.
195. Verschueren, K. (1996) Handbook of Environmental Data on Organic Compounds. 3rd edition. Van Nostrad Reinhold, New York, p. 1070 - 1072.
196. Vogel, T.M., Criddle, C.S. and McCarty, P.L. (1987) Transformations of
Halogenated Aliphatic Compounds. Environmental Science and Technology 21, 722 - 736.
197. Walker, L., Charles, W. and Cord-Ruwisch, R. (2006) Performance of a
Laboratory-Scale DICOM® Reactor - A Novel Hybrid Aerobic/Anaerobic Municipal Solid Waste Treatment Process. Orbit 2006 Conference Proceedings, p. 849 - 857.
198. Whooley, M.A. and McLoughlin, A.J. (1982) The Regulation of
Pyocyanin Production in Pseudomonas aeruginosa. Applied Microbiology and Biotechnology 15, 161 - 166.
199. Wiegel, J. and Wu, Q. (2006) Microbial Reductive Dehalogenation of
Polychlorinated Biphenyls. FEMS Microbiology Ecology 32(1), 1 - 15.
191
200. Wilhelm, E., Battino, R. and Wilcock, R. J. (1977) Low-pressure Solubility of Gases in Liquid Water. Chemical Reviews 77, 219 - 262.
201. Wilkinson, S., Klar, J. and Applegarth, S. (2006) Optimizing Biofuel Cell
Performance Using a Targeted Mixed Mediator Combination. Electroanalysis 18(19 - 20), 2001 - 2007.
202. Wilson, B.H., Wilson, J.T. and Luce, T. (1997) Design and Interpretation
of Microcosm Studies for Chlorinated Compounds. In Proceedings of the Symposium on Natural Attenuation of Chlorinated Organics in Ground Water (May 1997) p. 34. United States Environmental Protection Agency.
203. Wohlfahrt, G. and Diekert, G. (1997) Anaerobic Dehalogenases. Current
Opinion in Biotechnology 8, 290 - 295.
204. Wu, Q., Bedard, D.L. and Wiegel, J. (1996) Influence of Incubation Temperature on the Microbial Reductive Dechlorination of 2,3,4,6-Tetrachlorobiphenyl in Two Freshwater Sediments. Applied and Environmental Microbiology 62, 4174 – 4179.
205. Wu, Q., Bedard, D.L. and Wiegel, J. (1997) Effect of Incubation
Temperature on the Route of Microbial Reductive Dechlorination of 2,3,4,6-Tetrachlorobiphenyl in Polychlorinated Biphenyl (PCB)-Contaminated and PCB-Free Freshwater Sediments. Applied and Environmental Microbiology 63(7), 2836 - 2843.
206. Yamazaki, S., Kano, K., Ikeda, T., Isawa, K. and Kaneko, T. (1999) Role
of 2-amino-3-carboxy-1,4-naphtoquinone, A Strong Growth Stimulator for Bifidobacteria, as an Electron Transfer Mediator for NAD(P)+ Regeneration in Bifidobacterium longum. Biochimica Biophysica Acta 1428, 241 - 250.
207. Yaws, C.L. and Yang, H.-C. (1992) Henry’s Law Constant for Compound
in Water. In C. L. Yaws, editor, Thermodynamic and Physical Property Data, pages 181 - 206. Gulf Publishing Company, Houston, TX.
208. Ye, D., Quensen 3rd, J.F., Tiedje, J.M. and Boyd, S.A. (1992) Anaerobic
Dechlorination of Polychlorobiphenyls (Aroclor 1242) by Pasteurized and Ethanol-Treated Microorganisms from Sediments. Applied and Environmental Microbiology 58, 1110 - 1114.
209. Yeager, C.M., Arthur, K.M., Bottomley, P.J. and Arp, D.J. (2004)
Trichloroethylene Degradation by Toluene-oxidizing Bacteria Grown on Non-aromatic Substrates. Biodegradation 15(1), 19 - 28.
192
210. Yuan, S.Y., Su, C.J. and Chang, B.V. (1999) Microbial Dechlorination of Hexachlorobenzene in Anaerobic Sewage Sludge. Chemosphere 38(5), 1015 - 1023.
211. Yuan, S.Y., Wei, S.H. and Chang, B.V. (2001) Biodegradation of
Polycyclic Aromatic Hydrocarbons by a Mixed Culture. Chemosphere 41(9), 1463 - 1468.
212. Zhang, C. and Bennett, G.N. (2005) Biodegradation of Xenobiotics by
Anaerobic Bacteria. Applied Microbiology and Biotechnology 67, 600 - 618.
213. Zhuang, P. and Pavlosthasis, S.G. (1995) Effect of Temperature, pH and Electron Donor on the Microbial Reductive Dechlorination of Chloroalkenes. Chemosphere 31(6), 3537 - 3548.
214. Zinder, S.H. and Gossett, J.M. (1995) Reductive Dechlorination of
Tetrachloroethene by a High Rate Anaerobic Microbial Consortium. Environmental Health Perspectives 103, 5 - 7.
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Curriculum Vitae
Donny Lawrence James
Education
PhD (Biotechnology) - Murdoch University, Western Australia (2005 - 2009)
BSc (Hons) (Biotechnology) - Murdoch University, Western Australia (2002 - 2003)
BSc (Biotechnology) - Murdoch University, Western Australia (2001 - 2002)
Diploma (Biotechnology) - Ngee Ann Polytechnic, Singapore (1995 - 1998)
Publications
D. L. James (2009) Reductive Dechlorination of Chlorinated Hydrocarbons in Anaerobic Environments - A Literature Review. Soil and Sediment Contamination (Submitted).
R.Cord-Ruwisch and D. L. James (2009) Enrichment of Microorganisms Specific to Cyanocobalamin Reduction. Bioremediation Journal (Submitted).
R.Cord-Ruwisch, D. L. James and W. Charles (2009) The Use of Redox Potential to Monitor Microbial Reductive Dechlorination. Journal of Biotechnology 142, 151 - 156.
D. L. James, R.Cord-Ruwisch, D. Schleheck, M.J. Lee and M. Manefield (2008)
Cyanocobalamin Enables Activated Sludge Bacteria to Dechlorinate Hexachloro-1,3-butadiene to Non-Chlorinated Gases. Bioremediation Journal 12(4), 177 - 184.
G. Garau, W.G. Reeve, L. Brau, P. Deiana, R.J. Yates, D. James, R. Tiwari, G.W.
O’Hara and J.G. Howieson (2005) The Symbiotic Requirements of Different Medicago sp. Suggest the Evolution of Sinorhizobium meliloti and S. medicae with Hosts Differentially Adapted to Soil pH. Plant and Soil 276(1-2), 263 - 277.
Research Papers Presented
Platform Presentations
EBCRC Annual Conference - Perth, Western Australia (2007)
EBCRC Annual Conference - Brisbane, Queensland (2006)
EBCRC Annual Conference - Sydney, New South Wales (2005)
Poster Presentation
International Symposium on Environmental Biotechnology Leipzig, Germany (2006)
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Awards
Commercialisation Training Scheme - Australian Government (2008)
PhD Scholarship - EBCRC (2005)
Student Excellence Award - AusBiotech (2003)
Courses Completed
Media Skills Workshop - University of Western Australia (2009)
Research and Limited Scientific Diver Course - Evaluation Pty Ltd (2008)
Intellectual Property Work Experience - EBCRC (2007)
Science Writing Workshop - Writing Clear Science (2007)
Commercialisation Bootcamp - Australian Institute of Commercialisation (AIC) (2006)
Present Yourself With Impact Workshop - Dr. Bea Duffield and Gavin Blakey (2005)
Unsealed Radioactive Handling Course - University of Western Australia (UWA) (2005)
Fluorochrome In-Situ Hybridisation (FISH) Course - Murdoch University (2005)
Safety in Science Workshop - Murdoch University (2005)
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Acknowledgements
This thesis would not have been possible without the help and support from a number of
people over the last four years. Firstly, I wish to thank my supervisor Dr. Ralf Cord-
Ruwisch for his supervision and guidance in the thesis. Major sections of this project and
thesis would not have been possible if not for his motivation. Secondly, I wish to express
my gratitude to Dr. Matthew Lee for his help in setting up protocols that were the basis of
sampling and analyses in the project.
I would also like to thank the Environmental Biotechnology Co-operative Research
Centre (EBCRC) and Orica Australia Private Ltd for both the studentship and funding of
this research. In particular, I would like to express my appreciation to Dr. David
Schleheck for his recommendation on the use of cyanocobalamin for HCBD
dechlorination, and Dr. Mike Manefield, Dr. David Garman and Mr. James Stenning for
their overall assistance and advice in the project.
Special thanks to Associate Professor Robert Trengrove and Mr. Garth Brookes from
Separation Science laboratory for their expertise and support in maintaining equipment
required for Gas Chromatography analyses. I am also grateful to Dr. Wipa Charles and
Ms. Yingyu Law for setting up experiments associated with dechlorination using
methanogenic cultures.
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I am thankful to Mr. John Snowball for his assistance with the construction of electrical
devices that enabled the setup of automated reactors, and Mr. Murray in the workshop for
his costly but efficient help in the physical construction of my reactors.
Many thanks to my best friend, Thirumurugan (Ocean) and to all my other senior friends
in Singapore for their support throughout the years.
Last but not least, I wish to dedicate this thesis to my parents, especially my father who
was the vital boost to my aspirations in pursuing this degree. My parents, James Devas
and Polin James, brothers Justin Laval James and Elvin Sakayam James were integral in
my successful completion of this degree.
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