Open access peer-reviewed chapter

Microbial Degradation of Some Halogenated Compounds: Biochemical and Molecular Features

By Yu-Huei Peng and Yang-hsin Shih

Submitted: July 10th 2012Reviewed: March 4th 2013Published: June 14th 2013

DOI: 10.5772/56306

Downloaded: 1787

1. Introduction

Due to the advance of organic and synthetic chemistry and many applications of man-made organic compounds, lots of xenobiotic chemicals are produced and benefit our life. However, some of them are persistent in the environment and their toxicities are accumulated through the food webs. Due to the potential hazard for human and the ecosystem, the regulations on the usage of these persistent organic pollutants (POPs) and the development of safe decomposition methods are now in great request.

Tetrachloroethene (PCE) and trichloroethene (TCE) (Table 1) have been widely used as dry cleaning solvents and degreasing agents. Due to poor disposal practices and accidental release, they are within the most-abundant groundwater contaminants. Exposure to PCE is injurious to epidermis, kidney and nervous system [1]. It has been classified as a probably carcinogen [2]. Exposure to TCE leads to acute effects on liver, kidney, central nervous, and endocrine systems. It is also associated with several types of cancers based upon epidemiological research [3]. PCE and TCE are regulated in U.S.A. to a maximum contaminant level of 5 ppb. The use of PCE and TCE in the food and pharmaceutical industries has been banned across much of the world since the 1970s. However, these chemicals are still used as a degreasing agent for other demand. Besides, the mono-chlorinated ethene, vinyl chloride (VC), is known as carcinogen that causes liver cancer [4] and the cis-1,2-dichloroethene (cDCE) is harmful to nervous system, liver and blood cells [5].

Polybrominated diphenyl ethers (PBDEs), composed by two phenol rings and linked by one oxygen atom (Table 1), allow maximum ten bromide atom incorporated on the phenol rings to form 209 possible congeners. They have been widely used as flame retardants in many products over more than three decades. Their usage has protected both human lives and their properties from fire damage. PBDEs disrupt the balance of thyroid hormone, lead to reproductive toxicity, hepatic toxicity, immunotoxicity and developmental neurotoxicity in mammals [6, 7]. The toxicity of PBDEs and their metabolites are due to elevated free radicals, DNA damages, cell cycle blockage and apoptosis rate [8, 9]. Among the congeners of PBDEs, the usage of penta-BDEs and octa-BDEs has been banned in the European Union and several states of the USA; tetra-BDE to hepta-BDE have also been classified as POPs and the production of decabromodiphenyl ether (DBDE) will cease in 2013 in USA. However, the concentration of PBDEs in environment remains exponentially increasing because of the consequence of long-term usage [10]. Due to their ubiquitous distribution in the environment, potential toxicity, tendency for bioaccumulation, and the increased accumulation amount in the environment, the fates of PBDEs in the nature is serious concern for public health.

Hexabromocyclododecane (HBCD), a brominated aliphatic cyclic hydrocarbon (Table 1), is another widely used brominated flame retardant (BFR). It has animal thyroidal and developmental toxicity. The toxicity is due to altering the expression and function of metabolic enzymes, increasing hormone turn over and apoptosis [11, 12]. HBCD has been detected widely in biota and abiotic samples [11]. Due to its persistent, bioaccumulative, and toxic properties, HBCD has been proposed as a Substance of Very High Concern under the REACH regulations [13] and included on the USEPA's lists of Chemicals of Concern [14]; It is also under screening-level risk assessments to determine if it meets criteria of compounds in the Stockholm Convention and in the UN-ECE Protocol on POPs [15].

Most traditional remediation methods are not suitable for degrading chloroethenes, PBDEs and HBCD. For example, dehalogenation processes under oxidative, alkaline, or irradiation conditions are high cost of energy and treatment reagents [16, 17]. Pyrolysis is only limited for specific contaminated media with high heat conductance [18]. The generation of hazardous by-products is also a problem [19] It has been reported that PBDEs and HBCD could be photo-degraded [20]; however, pollutants accumulated in the soils, sediments, water bodies are not easy approach to light. Recently, the permeable reactive barrier made by zerovalent iron offers a new direction for halogenated compounds remediation [21]. Electrons offered from iron can reduce the halogenated compounds through reductive dehalogenation.

Zerovalent iron is cheaper than above processing methods, and the shortage of low efficiency is compensated by newly developed nanotechnology. Nanoscale zerovalent metals can degrade chloroethenes, PBDEs, and other contaminants with a fast kinetics and high efficiency [22-26]. There are still some limitations by using nano-metals, such as: toxic by-products generation due to incomplete dehalogenation [23], potential hazardous effect from nanoparticles [27], and large requirement for metals. Therefore, it comes to be one of the recent trends in developing nanomaterials with high efficiency and low environmental impact, and combining with other treatment technologies.

Microbiological approaches produce less intervention to the environment and are less expensive than physical or chemical methods. Biodegradation of chloroethenes has been extensively investigated and reviewed [28-30]. Bioremediation for PBDEs and HBCD are just at the beginning. The main objective of this review is to summarize current knowledge of microbial degradation of chloroethene, PBDEs and HBCD, especially from the biochemical and molecular point of view. We also attempt to compare the advantages and drawbacks of the combined approaches which may apply to field remediation.

2. Biodegradation of chloroethenes, PBDEs and HBCD

Biodegradation of chloroethenes, PBDEs and HBCD occurs in various environmental or living samples [31, 32]. In the environment, microorganisms play major roles in the degradation reactions; while intrinsic detoxicification systems in plants and animals bodies metabolize these compounds [31, 33]. In this article, the diverse and complex microbial degradation machineries are presented and compared.

Biotransformation of chloroethenes, PBDEs and HBCD in aerobic or anaerobic environments has been demonstrated (Fig. 1). In aerobic environment, chloroethenes and PBDEs are metabolized with the generation of energy or degraded cometabolically without energy-yield. In anaerobic condition, they are reduced through the energy yielded from the oxidation of electron donors, i.e. reductive dehalogenation or dehalorespiration. Biotransformation of HBCD might mediate through hydrolytic dehalogenation, which may occur either in aerobic or anaerobic conditions. Detail for the each type of reaction and the degraders will be described in the following sections.

2.1. Aerobic oxidative degradation

Under aerobic conditions, chloroethenes and PBDEs can be oxidized both cometabolically and metabolically (Fig. 1, left part). Metabolic degradation indicates the use of the above compounds as growth substrate. Chloroethenes and PBDEs more easily undergo aerobic transformation with less numbers of halogen substituent.

Metabolic degradation of cDCE and VC as sole carbon and energy source has been reported by many bacteria, such as the Pseudomonas sp. and Bacillus sp. [34, 35]. Using cDCE as auxiliary substrate for growth is much less [35] and not shown in TCE and PCE. After oxidative transformation, the auxiliary substrate may be mineralized or the carbon atoms may be incorporated into biomass. Microbial growth can be confirmed by monitoring the stable isotope fractionation and is suitable for field assessment [36].

On the contrary, cometabolic degradation of chloroethenes occurs fortuitously during the degradation of growth dependent substrates (auxiliary substrates), such as methane, ammonia, or aromatic hydrocarbons. Even cDCE can be cometabolized when VC is metabolic degraded [34]. Cometabolic degradation of TCE, DCE and VC is common [37, 38]. So far Pseudomonas stutzeri OX1 is the only one that could aerobically cometabolize PCE [39]. Therefore, without primary substrate supplement, intrinsic bioremediation with air or nutrients injection alone could not enhance the aerobic cometabolic mechanism and would not cause the microbial degradation of PCE and TCE contaminated sites.

PBDEs could be degraded into phenol or catechols by aerobic microbial through hydroxylation or bond cleavage [33] (Table 2). Sphingomonas sp. SS3 and SS33 can transform mono- or di-halogenated DEs for growth [40, 41]. In addition, Sphingomonas sp. PH-07 could break down several lower-bronimated BDE congeners (up to tri-BDEs) [42]. Other PBDEs degradation bacteria are reported [43]: Rhodococcus jostii RHA1 and Burkholderia xenovorans LB400 transform several lower-brominated BDE congeners (up to penta- and hexa-BDEs); Rhodococcus sp. RR1 transforms di- and mono-BDEs and the Pseudonocardia dioxanivorans CB1190 only degrades mono-BDEs. The transformation by-products include phenol, catechol, halophenol and halocatehol, indicating nonspecific attractions. These degraders might transform PBDEs through cometabolic reactions because auxiliary substrates such as diphenyl ether are supplemented. The Lysinibacillus fusiformis strain DB-1, cometabolically debrominate DBDE with the metabolism of lactate, pyruvate and acetate, is isolated [44].

So far, there is only one degrader been reported can transform HBCD: Pseudomonas sp. HB01 [45]. Since bromide atom was not detected after degradation reaction, such transformation might not through haloelimination. In general, cometabolism requires supplement of auxiliary substrates and there is no energy yielded. Therefore, microorganisms do not favor proceed this kind of reaction. Besides, the dehalogenation reaction is usually incomplete, resulting accumulation of toxic intermediates. The contribution on the bioremediation from cometabolism is limited [30].

2.2. Anaerobic reductive dehalogenation

Reductive dehalogenation is an anaerobic respiration process. Electron donors are oxidized and the halogenated compounds are reduced through accepting the electrons. The free energy generated from this reaction supports the growth of microbial degraders. Hydrogen atom replaces the halogen atoms one after another resulting in the dehalogenation sequence from higher-numbered compounds to lower-numbered ones. Contrary to aerobic degradation, the potential for reductive dehalogenation increases with the number of halogenated substituent [29].

Hydrogen gas is generated primary by fermentative and acetogenic bacteria (Fig.1, right part). The dehalorespiration bacteria compete with hydrogenotrophs, such as sulphate-reducers, nitrate-reducers, methanogens, acetogens, and other reducers [46, 47]. Except hydrogen, other electron donors also can be used for reductive dehalogenation, ex. Sulfurospirillum multivorans can also use pyruvate and formate [48].

Several mixed cultures and pure strains are known to reductively transform chloroethenes. Mixed culture could cooperate and transform PCE to ethene. The pure strains belong to different genus, such as Bacillus, Dehalobacter, Dehalococcoides, Desulfitobacterium, Geobacter, and Sulfurospirillum (Table 2). Most of them only dechlorinate PCE and TCE to cDCE. Only Dehalococcoides ethenogenes strain 195 can reductively dechlorinate PCE to ethene. The accumulated hazardous DCE and VC is a major obstacle in bioremediation of chloroethene contaminated sites. Dehalococcoides sp. strain BAV1, dechlorinates DCE and VC and cometabolizes PCE and TCE; the accumulated toxic compounds can be transformed into benign ethene [49].

Reductive debromination of PBDEs has been reported through pure strains (Table 2) or mixed cultures. Most of the debromination processes require TCE to be co-substrate. 20 mixed microcosms can degrade octa-BDE mixture to hexa- to mono-BDEs within 2 months [50]. Sulfurospirillum multivorans could debrominate DBDE into hepta- and octa-BDEs after 2 months of incubation. D. ethenogenes strain 195 could debrominate the octa-BDE mixtures into hepta- to di-BDEs after 6 months of incubation [51]. Dehalococcoides sp. Strain DG could degrade octa-BDE mixture into terta- and penta-BDEs or transform penta-BDE mixture into terta-BDE [52]. Several dechlorinating bacteria, Desulfitobacterium hafniense PCP-1, Dehalobacter restrictus PER-K23, Desulfitobacterium chlororespirans Co23 and Desulfitobacterium dehalogenans JW/IU-DCI debrominate the octa-BDE mixture and the most frequently detected congeners, penta 99 and tetra 47 when PCP, PCE, 3-chloro-4-hydroxybenzoate, or 3-chloro-4-hydroxyphenylacetate are applied as co-substrates [53]. Some mixture cultures do not need halogenated compounds to stimulate PBDEs transformation [50]. Recently, a lactate-dependent bacterium, Acetobacterium sp. strain AG, was isolated and can transform penta-BDE mixtures without other halogenated electron acceptors [52]. We also found that the cometabolism with glucose facilitated the biodegradation of mono-BDE, in terms of kinetics and efficiency in one anaerobic sludge in Taiwan [54].

In a mix microcosm, anaerobic environment necessary for dehalorespiration could be established by other symbiotic microorganisms. In our previous study, the mono-BDE is transformed to diphenyl ether in an aerobic culture from sewage sediment, indicating an anaerobic debromination reaction occurred. The enriched Clostridiales specie shown in the denatured gradient gel electrophoresis (DGGE) may responsible for such reaction [55].

2.3. Degradation enzymes

The metabolic pathway of VC is much clearer than that of cDCE. Alkene monooxygenase (AkMO) involves in the initial epoxidation step. The encoded genes (etnABCD) and the structures have been identified. Downstream events of the transformation are mediated through coenzyme M transferase (encoded by etnE gene), alcohol/aldehyde dehydrogenase, CoA transferase and CoM reductase/carboxylase. The final product, acyl-CoA, is then metabolized through TCA cycle [56]. Proteomic and transcriptomic analyses have confirmed the roles of above enzymes in aerobic VC transformation process.

Aerobic cometabolic degradation of chloroethenes is supposed through several kinds of oxygenases: toluene monooxygenase, toluene dioxygenase, phenol monooxygenase and methane monooxygenase [57]. P. stutzeri OX1 depletes PCE and releases chloride irons when toluene is applied as an auxiliary substrate [39]. PCE, DCE, and VC could be transformed by the purified toluene-o-xylene monooxygenase (ToMO). ToMO is a four-component enzyme which consists a catalytic oxygen-bridged dinuclear center encoded by touABE, a NADH ferredoxin oxidoreductase (encoded by touF), a mediating protein (encoded by touD), and a Rieske-type ferredoxin (encoded by touC). The touA~F genes cloned into E. Coli could make it to be PCE-degradable.

Different dioxygenases are supposed to involve in aerobic degradation of lower numbered PBDEs. 1,2-dioxygenase is involving in the initial dihydroxylation step when mono-halogenated DEs to be degraded [40]. Downstream degradation processes are supposed through phenol hydroxylases and catechol 1,2-dioxygenase. The transformation by-products range from phenol, catechol, halophenol and halocatechol, indicating nonspecific attack reactions [40, 41]. 2,3-dioxygenase is responsible to dihydroxylate lower numbered PBDEs and their similar chemicals such as DE in the close species Sphingomonas sp. PH-07 [42]. The range of PBDEs transformed by R. jostii RHA1 depends on the types of growth substrate. The enzymes responsible for degradation are inducible [43]. The expression of biphenyl dioxygenase (BPDO) and ethylbenzene dioxygenase (EBDO) are upregulated during PBDEs degradation. Ectopically expression of these enzymes in closed bacteria that bears no PBDEs degradation activity could transform PBDEs. EBDO depleted mono- through penta-BDEs and BPDO only depleted mono-, di- and one tetra-BDEs. [58].

The structures of HBCD and hexachlorocyclohexanes (HCHs) are quiet similar. Heeb et al. purified the HCH-converting haloalkane dehalogenase LinB, from Sphigobium indicum B90A and applied the enzyme for HBCD degradation. LinB transforms HBCD into pentabromocyclododecanols (PBCDOHs) and further tetrabromocyclododecadiols (TBCDDOHs) [59]. Whether LinB or other haloalkane dehalogenase are the de novo HBCD degradation enzyme is unknown. What enzyme responsible for HBCD degradation in Pseudomonas sp. HB01 is also waited to be uncovered.

Reductive dechlorination reactions are catalyzed by reductive dehalogenases (RDases) The purified PCE RDase, PceA, has proved to transform PCE and TCE to cDCE [60]. The function of TCE RDase, TceA, in transforming TCE to ethene has also been identified [61]. VcrA and BvcA catalyze the transformation of DCE to ethene [62, 63]. In addition to chloroethenes, RDases also could reduce other chlorinated compounds.

RDases which could debrominate PBDEs have not yet been identified. However, some PBDEs degradation bacteria also could transform chloroethene (Table 2), such as Dehalococcoides sp., Desulfitobacterium sp., and Sulfurospirillum sp. Whether these microorganisms use chloroethene RDases to transform PBDEs is unknown. It is also possible that enzymes with different degradation activities or substrate specificity within single degrader may cooperatively transform different PBDEs congeners.

2.4. The structure and function of reductive dehalogenase

Most RDases presented similar features and conserved motifs [28, 29]. In the N-terminus, RDases possess a putative signal sequence containing the twin-arginine translocation (Tat) motif. Such motif is presented in secretary proteins to be transported across the cytoplasmic membrane through the Tat export system. It is proposed that newly synthesized RDase proteins is folded with cofactors (corrinoid and iron-sulfur clusters) in the cytoplasm with the aids of chaperone proteins. The Tat sequence is then proteolytically cleaved during the maturation process. In the C-terminus, two iron-sulfur cluster binding motifs are presented. The Fe-S clusters cooperate with corrinoid, transfering electrons from upstream donors to chloroethenes and thus catalyze the dehalogenation reaction [28].

The localization of chloroethene RDases is supposed in the membrane, where they could accept electrons from proton producing hydrogenase via menaquinone. The membrane-bound characteristics of RDases has been proved, such as PceA of D. ethenogenes. The localization of constitutively expressed PceA in S. multivorans was initially found in the cytoplasmic fraction [65]. John et al. used freeze-fracture replica immunogold labeling technique and found it would be at the cytoplasmic membrane when cells grown on pyruvate or formate as electron donors [66].

2.5. Genomic structure and transcription regulation of reductive dehalogenases

The major catalysis reaction of RDases is directed by subunit A, encoded by reductive dehalogenase homologous A (rdhA) genes. Over 650 rdhA genes have been identified based upon genomic sequence annotation or homologous cloning [67]. However, most of them are not yet been functional characterized. It is common for one dehalorespiration bacterium baring multicopy of rdhA genes in the genome. Besides tceA and pceA genes, there is still 17 RDase genes with unknown function in the genome of D. ethenogenes strain 195 [68]. Whether the roles of these genes are relevant to dehalogenation remains unclear.

Most rdhA genes are organized with genes encoding for accessory proteins. The pce-gene cluster from D. hafniense strain Y51 constitutes pceA followed by pceB, pceC and pceT [69]. PceT is the trigger factor involving in folding newly synthesized polypeptides. It interacts with the Tat motif of PceA, thus solubilizing and stabilizing PceA polypeptide proceeding downstream maturation and transportation processes [62]. PceB protein contains three transmembrane domains and is assumed as a membrane anchor protein of PceA. PceC contains six-transmembrane domains, an FMN binding domain, and a C-terminal polyferredoxin-like domain. It is similar to the membrane-binding transcription regulators [28, 60]. More examples for the organization of different RDase gene clusters are presented in [28]. RdhA and rdhB genes usually locate adjacent and are the basic components of the rdh gene cluster. They would co-express in order to perform dehalorespiration together.

The expression and silence of RDases during dechlorination reaction is dynamic and regulated. It could be monitored through the amount of RNA or protein. Dehalococcoides sp. strain MB bares 7 RDase genes. Only dceA6 is highly expressed when PCE and TCE are transformed into tDCE. Transcription regulation protein binding site related to gene expression is detected in the upstream of dceA6 gene [70]. A shotgun metagenome microarray is created to investigate gene transcription in a mixed culture. rdhA14 and rdhB14, are the only two with higher transcript levels during VC degradation, while another 4 rdhA genes has higher transcript levels in the absence of VC [71]. The absolute quantification of RDase proteins during the dechlorination process is performed by using nano-liquid chromatography-tandem mass spectrometry in two PCE/TCE degradation consortia. Within 5 selected RDases, only the quantities of PceA and TceA are detectable [72]. The regulation on the expression of rdh genes during dechlorination reaction or steady state is not clear. How the physiological environments affect the gene expression is also unclear. Uncover these questions would be helpful for environmental monitoring and remediation.

2.6. The dynamic of degrading population and the evolution of degradation ability

The complete dehalogenation requires different microorganisms which bear various functions in degradation or growth support. Besides, it competes with methanogens and other reducers for H2. The snapshot of the microbial composition stands for specific ecological condition. The dynamic of composition reveals the effect of various remediation treatments and the interaction between microorganisms. The microbial compositions when co-incubated with zerovalent iron (ZVI) are analyzed by DGGE. The enrichment of iron-reducing bacteria would support the reduction activity of iron for multiple rounds of reactions; the enrichment of nitrate-reducing bacteria also facilitates the cometabolic dehalogenation. These may due to the synergistic effect [54]. Terminal restriction fragment length polymorphism (TRFLP) analysis is also used to analyze the microbial compositions [73]. The resolution limitation of these techniques makes underestimating the complexity of a community. Therefore, new technique is needed to detect specific microbes that are responsible for a key biodegradation process while present in the communities in low numbers. The 16S rRNA genes within a community could be analyzed by recently evolved pyrosequencing or phylogenetic microarray (PhyloChip). PhyloChip composes ten thousands unique 16S rRNA genes. The microbial compositions in TCE contaminated groundwater that is biostimulated or bioaugmented are analyzed. The increase of methanogens at late treatment stage coincident with the increase in methane concentration [74].

There is no close phylogenic relation among diverse dehalorespration degraders. Horizontal gene transfer (HGT) though transposable elements, transmissible plasmids or phage infections is assumed for such convergent evolution. Phylogenetic analyses of the sequences of rdh operon and the adjacent genomic structures support HGT. The pceABC operon in D. hafniense strain TCE1 has been shown to be presented in a circularized transposable element, Tn-Dha1 [75].The single-copy transfer messenger RNA gene (ssrA) essential in bacteria is a common target for mobile element. Integration of mobile element results in the duplication of ssrA gene around transported gene cluster. Many strain-specific rdhA genes collocates within such structures and in a region of high genomic variability between Dehalococcoides strains [76]. According to the metagenomic sequence analysis, one prophage element is located adjacent to tceAB genes in the Dehalococcoides-containing consortium, KB-1. The failure in detecting tceA gene expression in virus and the more closed transposase genes indicating higher possibility for HGT through transposable elements [71]. It seems that dehalorespiration degraders do no acquire RDase genes through single way. This would increase the diversity of degraders and function of RDase, which is advantageous for remediation.

3. Integration of biodegradation with other remediation methods

The degradation rate of natural attenuation is slower than chemical or physical treatments. Biostimulation or bioaugement are common strategies for bioremediation of chloroethene [30] Chemical supplements such as potassium permanganate or oxygen injection which can increase oxygen concentration are benefit for microbial dechlorination [77]. Kuo and his collages set biosparging wells for injection substrate and air into TCE contaminated area. Above 95% TCE was removed through cometabolic reactions because the elevated chemical oxygen demand (COD), microbial population, oxidation-reduction potential (ORP) and specific degrading genes after the supplement of substrate [78]. Shortage of auxiliary substrates or accumulation of toxic intermediates also decrease the dehalogenation effect, combined remediation methods may recover the above drawbacks.

Sequential anaerobic/aerobic biodegradation is one of the approaches to accelerate the degradation of recalcitrant halogenated compounds. Anaerobic degraders could target the higher-numbered halogenated compounds. Aerobic degraders only process lower-numbered halogenated compounds. They could transform the by-products produced from anaerobic degradation to antoxic compounds through metabolic or cometabolic reactions. Integration of these two systems makes it possible for complete mineralization.

Chloroethene and PBDEs can be depleted by microscale or nanoscale zerovalent metals [79, 80]. Preliminary dehalogenation of highly halogenated compounds by the reduced metal to generate less halogenated byproducts those are susceptible for microbial degradation. Therefore, the integration of zerovalent metals with biodegradation promotes the dehalogenation efficiency of each type of remediation methods. Reductive debromination of DBDE with nanoscale ZVI (nZVI) results various intermediates ranging from nona-BDEs to tri-BDEs. The known aerobic PBDEs degrader, Sphingomonas sp. strain PH-07, which is able to grow in the presence of nZVI, aerobically mineralizes the low brominated-DEs (tri-BDEs – mono-BDEs) from nZVI treatment [81]. The interactions between metals and microbes are complicated and delicate. H2 generated from the oxidation of metals promotes the growth of some dehalorespiration microbes. Some microorganisms could reduce oxidized metals for multiple runs of reductive reactions or degrade target compounds through cometabolism. Co-incubation with ZVI, microbes in the DBDE-degrading anaerobic sludges hinders the accessibility of MZVI to DBDE and reduced the removal ability in initial stage. However, the synergistic effect in DBDE degradation appears later on. According to the analysis of the microbial community change, co-incubation with MZVI leads to the enrichment of heterotrophic microbial populations bearing nitrate- or iron-reducing activities. The interaction between MZVI and microbes contributed to the synergistic effect [54]. Not only is the growth of microbes affected by metals, but also the expression of functional RDases. Bare nZVI down-regulate the expression of tceA and vcrA genes while coated particles up-regulate their expression [82]. In addition to the reduced metals, combining the electro-fenton process in aerobic degradation is also a newly evolved and potential way in bioremediation. Application of electrolysis also stimulate the microbial reductive dechlorination and oxidative activities [83].

There are advantages by using each type of remediation approach, while there are also limitations and drawbacks. Combining biodegradation with other abiotic/biotic degradation approaches could overcome their weakness and accelerate the degradation efficiency. The recalcitrant halogenated compounds could be completely mineralized. The impact on environment might also be minimized. Integration of different approaches is a new direction for future investigation.

4. Conclusion

The current knowledge of microbial degradation of chloroethenes, PBDEs and HBCD, has been summarized and reviewed. The biodegradation of these halogenated compounds through aerobic oxidation, aerobic cometabolization, or reductive dehalogenation are introduced. The correspondent enzymes are discussed from the biochemical and molecular point of view. The structure and function of RDases, as well as gene expression regulation and genomic evolution are the major focus. Integration and sequential anaerobic/aerobic biodegradation or (electro)chemical/microbial degradation are suggested for overcoming the disadvantages of single type of treatment. It is possible to completely mineralize these halogenated pollutants by the combination of bio- and abiotic processes and shows promise for site remediation in natural settings and in engineered systems.

Figure 1.

Aerobic degradation and anaerobic reductive dehalogenation reactions of chloeoethenes, PBDEs and HBCD. Circle dot indicates above compounds; star indicates reductive dehalogenation driven by the oxidation of electron donors or occurring cometabolically with other dehalorespiration process; gray arrow indicates the hydrolytic dehalogenation of HBCD.

PCE/TCEPBDEsHBCD
or H
Molecular weight165/131249.1 ~ 959.2641.7
Water solubility (mg l-1)150/12804.8 ~0.020.003
ToxicityaEpidermis, liver and kidney damage, immune- and neuro-toxicity, reproductive and endocrine effects, probably cancerDisrupt the balance of thyroid hormone, reproductive, hepatic, and immunotoxicity, developmental neurotoxicityThyroidal and developmental toxicity
Abiotic degradationChemical oxidation ,
irradiation,
reduced metals (Fe, Fe/Pd)
Pyrolysis,
photolysis,
reduced metals (Fe, Fe/Pd),
Pyrolysis,
photolysis
Biological degradationAerobic/anaerobicAerobic/anaerobicAerobic/anaerobic(?)

Table 1.

Physicochemical properties, biological impacts and degradation routes of chloroethenes, PBDEs and HBCD.

a: the toxicity of HBCD is based upon investigation in animal model.


SubstrateEnd-productsGenes creferences
Aerobic
Burkholderia xenovorans LB400Hexa-BDE to mono-BDE,Hydroxylated-BDE[43]
Lysinibacillus fusiformis strain DB-1DBDEND[44]
Pseudonocardia dioxanivorans CB1190Mono-BDEND[43]
Pseudomonas sp. HB01HBCDPBCDOHs, TBCDDOHs[45]
Pseudomonas stutzeri OX1PCECl-,ToMO-touABCDEF[39]
Rhodococcus jostii RHA1Penta-BDE to mono-BDEHydroxylated-BDEBPDO- bphAa,
EBDO- etbAa1, etbAc
[43, 84]
Rhodococcus sp. RR1Di-, and mono-BDEND[43]
Sphingomonas sp. PH-07Tri-, di-, and mono-BDECatechol, dibromophenol, dihydroxy mono- and dibromo-BDE[42]
Sphingomonas sp. SS3Fluoro-, chloro-, and bromo-DEPhenol, catechol, Halophenol and Halocatehol[40]
Sphingomonas sp. SS33Di-, and mono-; fluoro-, chloro-, and bromo-DEPhenol, catechol, di-, and mono-halophenol, di-, and mono-halocatehol[41]
Anaerobic
Acetobacterium sp. strain AGPenta-BDE mixtureaTetra-, tri-, and di-BDEs[52]
Bacillus sp JSK1PCECis-DCE[85]
Dehalobacter restrictusPCE, TCECis-DCEpceA,[67, 86]
Dehalococcoides sp. strain BAV1DCE, VCEthenebvcA[49, 87]
Dehalococcoides sp. strain DGTCE
Octa-BDE mixtureb
Penta- BDE mixturea
Ethene
Tetra- and penta-BDEs
Tetra- BDE
[52]
Dehalococcoides ethenogenes strain 195PCE, TCE, cis -DCE, and VC
Octa-BDE mixtureb
Ethene
Tetra-, penta-, hexa- and Penta-BDEs
pceA, tceA[51, 67, 89]
Dehalococcoides sp. strain MBPCE, TCETrans-DCE[90]
Desulfitobacterium chlororespirans strain Co23Octa-BDEND[53]
Desulfitobacterium dehalogenans strain JW/IU-DC1PCE, TCE,
Octa-BDE,
ND[53, 67, 91]
Desulfitobacterium hafniense PCP-1Octa-BDEND[53]
Geobacter lovleyi strain SZPCE, TCEcDCE[92]
Sulfurospirillum multivoransPCE, TCE
DBDE
DCE
Octa- and hepta-BDEs
pceA[51, 67]

Table 2.

Selected bacteria which degrade chloroethens, PBDEs and HBCD.

a penta- BDE mixture: hexa-, penta-, and tera-BDEs.


b Octa-BDE mixture: nona-, octa-, hepta-, and hexa-BDEs.


c Genes that only relevant to degradation of chloroethens, PBDEs and HBCD are listed.


ND: data not shown. BPDO: biphenyl dioxygenase. EBDO: ethylbenzene dioxygenase. PBCDOHs: pentabromocyclododecanols. TBCDDOHs: tetrabromocyclododecadiols.


Acknowledgments

The authors thank National Science Council (NSC), Taiwan, ROC for financial support.

© 2013 The Author(s). Licensee IntechOpen. This chapter is distributed under the terms of the Creative Commons Attribution 3.0 License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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Yu-Huei Peng and Yang-hsin Shih (June 14th 2013). Microbial Degradation of Some Halogenated Compounds: Biochemical and Molecular Features, Biodegradation of Hazardous and Special Products, Rolando Chamy and Francisca Rosenkranz, IntechOpen, DOI: 10.5772/56306. Available from:

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