Biochemical Dechlorination of Hexachloro-1,3 …...PHB Poly-ß-hydroxybutyrate PTFE...

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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


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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

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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.


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


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


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 (○).


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


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.


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.


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 (■).


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


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.


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).



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.


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.


3.1 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.


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.


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.


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.


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.



2.1.1 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.


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




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.


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.


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).


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).


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


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.


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.


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.


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.


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.


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.


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.


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).


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.


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.


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


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.


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.


109



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


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).



2.4 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.


3.1 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


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.


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


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.


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


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,


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.


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


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).


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


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.


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.


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.


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.


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.


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


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).


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.



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.



2.8 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.


3.1 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


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.


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


156

concentrations of HCBD, and could be used to monitor the dechlorination of other

chlorinated hydrocarbons.

EAg/AgCl affects the rate of HCBD dechlorination.


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


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


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).


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


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.


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.


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

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Curriculum Vitae


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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