Appl Microbiol Biotechnol (1998) 49: 489±499
Ó Springer-Verlag 1998
MINI-REVIEW
R.-M. Wittich
Degradation of dioxin-like compounds by microorganisms
Received: 29 August 1997 / Received revision: 6 January 1998 / Accepted: 8 January 1998
Abstract Polychlorinated dibenzo-p-dioxins (PCDD) and polychlorinated dibenzofurans (PCDF; PCDD/F, dioxins) have not been commercially produced in bulk amounts, as were polychlorinated biphenyls and other haloaromatic organics. Within the past two decades a lot of information has accumulated on the biodegradation of PCDD/F and other dioxin-like compounds because of their toxicity and because of signi®cant environmental concern about many congeners of this class of chemicals. PCDD/F are subjected to reductive dehalogenations leading to less halogenated congeners, which can be attacked eciently by fungal and bacterial oxidases and dioxygenases. In several cases these compounds can be utilized as carbon and energy sources. Pathways for their enzymatic degradation and the organisation of the corresponding degradative genes have been elucidated. Consequently, biotechnological applications will exploit the degradative potential of such microorganisms for bioremediation of contaminated sites.
Introduction Polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans (PCDD/F) are produced by incineration processes and, unintentionally, as by-products during chemical reactions. Many of their congeners are highly toxic haloaromatic compounds with undesirable eects on the environment and animals, including men. The chemical class of dioxin-like compounds should not only comprise the so-called ``dirty dozen'', namely those R.-M. Wittich Bereich Mikrobiologie, Gesellschaft fuÈr Biotechnologische Forschung ± GBF, Mascheroder Weg 1, D-38124 Braunschweig, Germany Tel.: +49 531 6181 557 Fax: +49 531 6181 411 e-mail:
[email protected]
PCDD/F having a 2,3,7,8 substitution pattern, including some higher and less-chlorinated congeners of reduced toxicity. The term should also include their brominated analogues, as well as low-chlorinated and unchlorinated dibenzo-p-dioxin (DD) and dibenzofuran (DF). Because the oxygen bridge connecting the two (halo-)aromatic nuclei is a commonly shared structural element, the (halogenated) diphenyl ethers (DE) and some of their carboxylated or dierently substituted derivatives should be subsumed in this term. In Fig. 1 the basic structures of PCDD, PCDF and polychlorinated diphenyl ethers (PCDE) are shown. The release into the ecosphere of dioxin-like compounds from anthropogenic sources has created a strong demand for legislation and executive activities. Because of the toxicity of dioxins, mentioned above, techniques have been developed to remediate dioxin-contaminated soil, building rubble from production facilities of decomissioned chemical plants, used transformers and capacitors containing polychlorinated biphenyls as insulator or dielectric material, which are highly contaminated with PCDD/F. Sites polluted with pentachlorophenol formerly used as a timber-preserving agent are also highly contaminated with PCDD/F. Up to now incineration has been the only, or the only inexpensive method for almost totally destroying dioxins, although many other physicochemical treatment procedures have been developed. From an economical point of view, thermal treatment of vast masses of contaminated soils and sediments is not feasible and biological alternatives are being sought. The destruction of organic compounds by microbial biocatalysts, comprising fungi as well as bacteria, plays an important role in the global carbon cycle. This biological potential was, and is still being exploited in composting and other degradation processes. Biotechnologies for the remediation of diesel and oil spills and bioremediation technologies for halogenated solvents and other industrial chemicals of environmental concern have been developed. The implementation of such bioremediation processes for the removal of toxic PCDD/F and related critical com-
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Fig. 1 Chemical structures of (polyhalogenated) dibenzo-p-dioxins (PCDD), dibenzofurans (PCDF ) and diphenyl ethers (PCDE). The maximal possible congener numbers are given in brackets; m, n 0± 4(5)
pounds, therefore, represents a challenge for microbiologists and environmental engineers.
Aerobic degradation of dioxin-like compounds: co-oxidation of dioxin-like compounds by naphthalene- and biphenyl-degrading bacteria The ®rst investigations were carried out by Kearny et al. (1972), Matsumura and Benezet (1973), and Ward and Matsumura (1978) who investigated microbial biodegradation of 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8TCDD) several years before the Seveso accident occurred in Italy in 1976. In the course of this accident, several kilograms of 2,3,7,8-TCDD was released into the environment. Studies by Philippi et al. (1981, 1982) and by Quensen and Matsumura (1983) on aerobic biodegradation of 2,3,7,8-TCDD in laboratory systems gave clear evidence of bacterial transformation of this compound in the presence of strains isolated from the polluted Seveso area and elsewhere. Depletion of labelled 2,3,7,8-TCDD from model ecosystems was also reported by Matsumura et al. (1983), who also detected polar metabolites. Cerniglia et al. (1979) and KlecÏka and Gibson (1979, 1980) used naphthalene- and biphenyl-degrading Pseudomonas sp. and Beijerinckia sp. strains (the latter were recently assigned to the genus Sphingomonas) for the transformation of DD, DF and several monochlorinated derivatives. Degradation of all compounds was incomplete and gave rise to dead-end products. The bacteria attacked the compounds at the lateral 1,2 and 2,3 posi-
tions (Fig. 2A) with high regioselectivity. A cis-1,2-dihydrodiol and a cis-2,3-dihydrodiol were identi®ed as the initial products. These dihydrodiols were similar to those already known from biodegradation pathways for benzene, toluene, biphenyl (Fig. 2A) and other aromatic compounds. The dihydrodiols of DD and DF were then dehydrogenated by a dihydrodiol dehydrogenase present in cell-free extracts of the naphthalene-degrading strain Pseudomonas sp. NCIB 9816 (Cerniglia et al. 1979). Enzymes of both Beijerinckia sp. and Pseudomonas sp. strains were able to cleave the aromatic ring, giving rise to a yellow meta-cleavage product indicative for a 2-hydroxymuconate semialdehyde structure. In Fig. 2A the pathway for the aerobic degradation of biphenyl (1) is compared with the pathway for the incomplete degradation by co-oxidation of DF (2). The cleavage of the aromatic ring of DF by a naphthalene-degrading pseudomonad was con®rmed by Selifonov et al. (1991). The authors succeeded in isolating a ring-cleavage product and identi®ed it by GC/MS and NMR investigations as the enol (13) of 4-[2¢-(3¢-hydroxy)benzofuranyl]-2-oxobut-3-enoic acid. The hydrolase encoded by the nahE gene of the nah operon was not able to cleave pyruvate from this molecule, as it does during naphthalene degradation from the corresponding 4-(2-hydroxyphenyl)-2-oxobut-3-enoic acid. The identi®cation of initial dioxygenation products of DF was taken much further by Resnick and Gibson (1996). They showed that attack on this heterocyclic aromatic compound yielded about 60%±70% of the cis-1,2-dihydrodiol and 30%±40% of the cis-2,3-dihydrodiol, and determined their absolute con®guration. Bianchi et al. (1997) found similar ratios of cis-dihydrodiols, which were produced by a salicylate-grown mutant of a naphthalene-degrading Pseudomonas ¯uorescens strain de®cient in its dihydrodiol dehydrogenase. The biphenyl-mineralizing Alcaligenes sp. strain JB1 partially degraded some mono- and dichlorinated DF and DD in batch and in continuous culture. During degradation, mono- and dihydroxylated derivatives of several of the target compounds accumulated. Interestingly, 5-chlorosalicylate was identi®ed during conversion of 2-chlorodibenzofuran (2-CDF; Parsons and Storms 1989; Parsons et al. 1990). Therefore, hydrolytic cleavage of the side-chain must have followed oxidative or hydrolytic cleavage of the furan ring system of the monochlorinated compound 13 (Fig. 2A).
Degradation by bacteria capable of mineralizing DD, DF and DE Degradation of the non-halogenated carbon skeleton The degradation of DF by microorganisms in creosote mixtures, where it is present amongst other polycyclic aromatic hydrocarbons, has already been reported by Lee et al. (1983) and by Foght and Westlake (1988). The enrichment and isolation of DF-, DD- and DE-miner-
Fig. 2A±C Degradative pathways for, (A) the mineralization of biphenyl (1) and the co-oxidation of dibenzofuran (2); (B) the co-oxidation of ¯uorenone (2a) and (C) the mineralization of diphenyl ether (3), dibenzo-p-dioxin (4), and dibenzofuran (2). Bacterial dioxygenation furnishes the cis-dihydrodiols shown in the second row; unstable dihydrodiols (hemiacetals) are shown in brackets. The other intermediates are as follows: the dihydrodiol of ¯uorenone (2b), 2,3-dihydroxybiphenyl (5), 1,2-dihydroxydibenzofuran (6), 2,3-dihydroxydibenzofuran (7), phenol (8), catechol (9), 2,2¢,3-trihydroxydiphenyl ether (10), 2,2¢,3-trihydroxybiphenyl (11), 2-hydroxy-6-oxo-6-phenylhexa-2,4-dienoic acid (12), 4-[2¢-(3¢-hydroxy)benzofuranyl]-2-oxobut-3-enoic acid (13), 2-hydroxy-6-oxo-6-(2-hydroxyphenoxy)hexa-2,4-dienoic acid (14), 2-hydroxy-6-oxo-6-(2-hydroxyphenyl)hexa-2,4-dienoic acid (15), benzoic acid (16), salicylic acid (17), gentisic acid (18), 2-hydroxypenta-2,4-dienoic acid (C5), 2-hydroxy-cis,cis-muconic acid (C6)
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alizing bacterial strains then represented a new approach to the complete degradation of non-halogenated model substrates and some chlorinated derivatives. Bacterial isolates of the gram-positive genus Brevibacterium (Strubel et al. 1989, 1991) later were reassigned to the new genus Terrabacter (Schmid et al. 1997). Other DFand DD-degrading bacteria, formerly identi®ed as pseudomonads (Fortnagel et al. 1989a,b, 1990) by 16S rRNA sequence analyses were assigned to the genus Sphingomonas (Moore et al. 1993, and unpublished results). This new taxon became more important after the identi®cation of several other diaryl-ether-degrading soil bacteria (Schmidt et al. 1992a; Wittich et al. 1992). Another bacterium capable of growth on DF but of unknown taxonomy was isolated by Frahne et al. (1991). Elucidation of the degradative pathways of DF, DD and DE revealed a new type of oxidative attack with high regioselectivity and speci®city for the angular position on the two carbon atoms adjacent to the ether bridge (Fig. 2C). From DF, DD, DE and 3-carboxyand 4-carboxydiphenyl ether an unstable hemiacetal was formed by this type of reaction, which then decayed spontaneously to 2,2¢,3-trihydroxybiphenyl, 2,2¢,3-trihydroxydiphenyl ether, phenol and catechol, or phenol and protocatechuate in the case of a carboxylated DE. The structural analogue of the diaryl ether DF, the aromatic ketone ¯uorenone (2a in Fig. 2B), was transformed to a stable dihydrodiol (2b in Fig. 2B) and therefore was indicative of this new mechanism. The dihydroxylated aromatic rings of the trihydroxy intermediates (Fig. 2C, 10, 11) were then meta-cleaved. Upon hydrolysis of the side-chain, salicylate and catechol respectively were formed. During DF degradation a spontaneous chemical side-reaction, competing with the hydrolase, occurs by addition of a phenolic OH group to the double bond of the aliphatic side-chain, giving rise to a by-product. Other unproductive reactions leading to dead-end products are fortuitous orthocleavage reactions resulting in muconic acid derivatives of 2,2¢,3-trihydroxydiphenyl ether. The degradation pathways that are outlined in Fig. 2C have been proposed for the breakdown of these diaryl ether compounds (Engesser et al. 1989, 1990; Fortnagel et al. 1989a,b, 1990; Harms et al. 1990; Schmidt et al. 1992a; Strubel et al. 1991; Wittich et al. 1990, 1992). The route for the degradation of 3-methyldiphenyl ether, in principle, follows the same course via the intermediates phenol and 4-methylcatechol, the degradation of which proceeds by an unusual ortho-cleavage pathway (Schmidt et al. 1992b). Mechanistic investigations using 18 O2 clearly con®rmed the angular dioxygenation of diaryl ether compounds (Wilkes et al. 1992). Several structural analogues containing another hetero atom (N, S) were attacked by the same angular dioxygenase as well as carbonyl analogues from which stable dihydrodiols were obtained (Engesser et al. 1989, 1990; Fortnagel et al. 1989b; unpublished data). Some of the above results were con®rmed by the work of Monna et al. (1993), who isolated a DF-degrading bacterium,
which was identi®ed as a Staphylococcus auriculans. This strain attacked DD and other related compounds but did not grow on them. Several hydroxylated, methoxylated, acetoxylated DF and nitrodibenzofurans were converted by the enzyme systems of strain HH69 to the corresponding salicylates (Harms et al. 1995). Some of these hydroxylated derivatives, as well as 2-hydroxydibenzo-p-dioxin, are utilized for growth by strains HH69 and RW1. They are degraded via gentisate in the case of 2-hydroxydibenzofuran and via 4-hydroxysalicylate in the case of 3-hydroxydibenzofuran. 2-Hydroxydibenzo-p-dioxin is degraded through a novel pathway via 1,2,4-trihydroxybenzene (hydroxyhydroquinone, unpublished results). Another mechanism for DE degradation, similar to the initial steps in the catabolic pathway of biphenyl (Fig. 2A), was reported by Pfeifer et al. (1989). Upon dioxygenation, dehydrogenation of the dihydrodiol and subsequent meta-cleavage of 2,3-dihydroxydiphenyl ether between C-1 and C-2 the resulting ester was subjected to an intramolecular transesteri®cation. During this reaction, phenol and the dead-end product 2-pyrone-6-carboxylic acid were formed (Pfeifer et al. 1993). Phenol had also been identi®ed as one of the intermediates of DE degradation by a Pseudomonas cruciviae. This strain excreted 2-phenoxymuconic acid during growth, the ortho-cleavage product of 2,3-dihydroxydiphenyl ether (Takase et al. 1986). The mechanism by which a pseudomonad degrades 3-carboxydiphenyl ether, furnishing phenol as the dead-end product (Topp and Akhtar 1990, 1991), remains to be elucidated. Biochemical and genetic analyses of pathways of bacteria for diaryl ether breakdown The initial dioxin dioxygenase system of the DF- and DD-mineralizing strain Sphingomonas sp. RW1 has been puri®ed and characterized (BuÈnz and Cook 1993). It is composed of a FAD-containing reductase of 44 kDa, transferring electrons from NADH via a ferredoxin of the putidaredoxin type (2Fe-2S) of 12 kDa to the terminal heterotetrameric (a2b2) oxidase with subunit sizes of 45 kDa and 23 kDa respectively. This terminal oxygenase binds both the aromatic substrate and dioxygen. The genetic elements coding for individual proteins of this system were identi®ed on the chromosome, cloned, sequenced and functionally expressed (Armengaud and Timmis 1997, and unpublished results). The genes encoding the enzyme sequence necessary for DD and DF degradation by strain RW1 are not clustered as in other operon systems but spread over the chromosome. Even the individual elements of the initial dioxygenase are located at some distance from each other. Close to the dioxygenase, a cluster of several genes was found, which probably code for the degradation of hydroxylated DD and DF (Armengaud et al., unpublished results). The gene coding for the next enzyme, a meta-cleaving dioxygenase (32 kDa) speci®c for the substrates 2,2¢,3-tri-
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hydroxybiphenyl and the corresponding diaryl ether, has been identi®ed on the chromosome of strain RW1. This gene and its product have been characterized in detail (Happe et al. 1993). The catalytic ferrous iron of this extradiol dioxygenase was further investigated; it is coordinated by histidine residues of the polypeptide (Bertini et al. 1995a,b). Two isofunctional hydrolases of 29 kDa and 31 kDa respectively, cleaving o the C-5 side-chain of 2-hydroxy-6-oxo-6-(2-hydroxyphenyl)hexa-2,4-dienoic acid, were puri®ed and characterized by BuÈnz et al. (1993). At least one of them should be active in the degradative pathway of DF, the other probably in that of biphenyl. The products from DF, salicylic acid and 2-hydroxypenta-2,4-dienoic acid, in turn, are degraded by known enzymes. The same is true for the degradation of catechol and 2-hydroxycis,cis-muconic acid, which are obtained through spontaneous cleavage of the preceding ester bond in the pathway of DD degradation. The angular dioxygenase speci®c for the more polar compounds 3-carboxy- and 4-carboxydiphenyl ether was cloned, sequenced and expressed in Escherichia coli (Dehmel et al. 1995). The enzyme system furnishes phenol and protocatechuate upon attack of the above substrates at the angular 1,2 position of the carboxylated aromatic ring. The reductase (33.6 kDa) of this system containes a chloroplast-type (2Fe-2S) centre, whereas the oxygenase component of 46.3 kDa exhibits the typical Rieske-type (2Fe-2S) iron-sulphur centre. Comparison of the amino acid sequence, deduced from the gene sequence, shows this dioxygenase to belong unambiguously to dioxygenase of class IA, also comprising phthalate dioxygenase. The dioxygenase of strain RW1, attacking the more lipophilic DF and DD, was assigned to class IIA of the aromatic ring-activating dioxygenases, to which pyrazon dioxygenase also belongs. Liaw and Srinivasan (1989, 1990a) cloned and expressed a gene in heterologous hosts, termed dpe, of an Erwinia sp. strain that probably encodes the enzyme responsible for the cleavage of DE. Its gene product, of molecular mass 21 kDa, was not shown to be active in vitro. The enzyme cannot be assigned to any known bacterial aromatic ring-activating dioxygenase system nor, because of its small size, to the cytochrome P-450 oxidases, which, with a single exception, are proteins of about 45 kDa (Sariaslani 1991). The products of the cleavage reaction performed by resting cells on DE, 4-CDE, and 4-hydroxydiphenyl ether were not further investigated. A copper-resistant mutant of the wild-type strain was claimed to be capable of degrading 2,7DCDD but not 2-CDD (Liaw and Srinivasan 1990b). The enzymatic mechanism of the initial attack on this compound remains uncertain. Aerobic degradation of halogenated diaryl ethers Mono- and dihalogenated DE were mineralized by Sphingomonas sp. strain SS3 and its derivative SS33 via
4-halophenol and 4-halocatechol, provided that halogen substituents were at positions 4 and 4,4¢. From other isomers, the corresponding halocatechols were formed as the dead-end products (Schmidt et al. 1992a, 1993). A hybrid strain has been constructed by Halden and Dwyer (1996) that degrades 3-chloro- and 4-chloro-3carboxydiphenyl ether. For this purpose the 3-chloroand 4-chlorophenol-degrading strain Pseudomonas sp. B13 was equipped with the carboxydiphenyl ethercleaving dioxygenase system of strain Pseudomonas pseudoalcaligenes POB310 (Dehmel et al. 1995). The almost complete degradation of 2-chloro- and 3-chlorodibenzofuran was performed by a consortium consisting of the DF-mineralizing strain Sphingomonas sp. RW16, the 4-chloro- and 5-chlorosalicylate-mineralizing strain Pseudomonas putida RW10, and two other strains of unknown function, which were necessary for growth of the co-culture. The initial attack by strain RW16 of the halogenated DF proceeded through attack on the halogenated as well as the non-halogenated aromatic ring, in a manner similar to that already described by Harms et al. (1991). Neither strain RW16, carrying out the initial reactions, nor strain RW10 possesses genes encoding the chlorocatechol pathway enzymes. Degradation of the chlorosalicylate(s) excreted by strain RW16 took place in strain RW10 through protoanomine (Blasco et al. 1995) via the intermediate 4-chlorocatechol. Both compounds and the product of the hydrolysis of protoanemonin, 3-acetylacrylic acid, are utilized for growth by strain RW10 (unpublished results). The isomeric diaryl ether compound 4-CDF was attacked solely at the non-halogenated aromatic ring because the other position is sterically blocked. The dead-end metabolite 3-chlorosalicylate is excreted in stoichiometric amounts and, in turn, is mineralized by the 3-chlorosalicylate-degrading strain Burkholderia sp. JWS (Arfmann et al. 1997). Thus the potential of DFdegrading bacterial strains, such as Sphingomonas sp. RW1, for the productive conversion of PCDD/F seems to be restricted to less-halogenated congeners because of the relatively narrow substrate range of the initial dioxygenase (Wilkes et al. 1996). An extensive study on the in¯uence of the substitution pattern on the depletion of many of the 210 congeners of PCDD/F was performed with bacterial strains of dierent taxa. Some of these strains were degraders of chlorobenzenes and only Sphingomonas sp. strain HH69 was known to be capable of mineralizing DF and of cooxidizing DD and 3-CDF (Fortnagel et al. 1990; Harms et al. 1990, 1991). In this study by Schreiner et al. (1997), strain HH69 depleted 31% of 2,3,7,8-TCDF and 15% of 2,3,7,8-TCDD within 84 days, whereas the 1,2,4,5tetrachlorobenzene-mineralizing Burkholderia (formerly Pseudomonas) sp. strain PS12 (Sander et al. 1991) depleted 64% of the 2,3,7,8-TCDF from the culture broth in the same period, but 100% of 2,3,7,8-TCDD within 25 days. Concentrations of PCDD/F in the assay systems were between 1 nM and 100 nM. In another study on 2,3,7,8-TCDD degradation, with resting cells of the DD-
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and DF-mineralizing strain Sphingomonas sp. RW1, no signi®cant degradation was detected. Neither metabolites nor 14CO2 evolution from the labelled compound were monitored within 8 days of incubation (W. Jonas, NATEC Institut Hamburg, personal communication).
Anaerobic dehalogenation of chlorinated dibenzo-p-dioxins and dibenzofurans Studies on the anoxic degradation of polychlorinated biphenyls and chlorobenzenes, for example, in anaerobic sediments containing methanogenic consortia, indicated the accumulation of less chlorinated congeners. Haloaromatics and haloalkenes are used as electron acceptors in anoxic respiratory processes, releasing chloride from the electron-rich organic substrates (Mohn and Tiedje 1992). The ®rst indications of biological anaerobic dehalogenating processes for PCDD/F were reported by Townsend (1983) and Beurskens et al. (1993), who analyzed contaminated sediments and detected altered congener distribution patterns. Earlier reports on 2,3,7,8TCDD degradation by Meyer and Bartha (1988) demonstrated poor mineralization to CO2 (0.1%±0.5%) within 4±6 months. The recovery rates with regard to the initial TCDD concentration were between 48% and 93%.
Fig. 3A, B Anoxic reductive dehalogenations of PCDD by mixed-sediment cultures. Further dehalogenations of (A) the upper hexachlorodibenzo-p-dioxins (HxCDD) (Adriaens and GrbicÏ-GalicÏ 1994) have to be expected, but were not demonstrated. Almost identical results for identi®ed intermediates from 1,2,3,4-tetrachlorodibenzo-p-dioxin (1,2,3,4-TCDD) (B) were obtained by Beurskens et al. (1995) and Ballerstedt et al. (1997)
The halogens of PCDD/F were predominantly removed from carbon atoms 1, 4, 6 and/or 9 (relative substituent positions, Fig. 3A). These dehalogenations led to much more toxic congeners, which were still substituted at the lateral 2,3,7,8 positions (Adriaens and GrbicÏ-GalicÏ 1992, 1994, Adriaens et al. 1995; Barkovskii and Adriaens 1996). Similar experiments were performed with the much less toxic congener 1,2,3,4-TCDD which, therefore, was used as a model for anoxic dehalogenation experiments. Results demonstrated reductive dehalogenation at the lateral positions too, leading to 2-CDD as the ®nal end-product (Toussaint et al. 1992; Beurskens et al. 1995). The same compound was identi®ed as the product in studies on the anaerobic dehalogenation of 1,2,3,4-TCDD (Fig. 3B) by Saale river sediments (Ballerstedt et al. 1997). From these experiments a pure culture was isolated, which was capable of dehalogenating 1,2,4-trichlorodibenzo-p-dioxin in the presence of formate, acetate and thiosulphate (Ballerstedt and Lechner 1998). A pathway for the anaerobic dehalogenation of PCDD is shown in Fig. 3B. It clearly indicates that monohalogenated aromatics are not dehalogenated further. Dwyer and Tiedje (1986) have reported that DD and DE were completely mineralized to methane via benzene and catechol, and benzene and phenol, respectively.
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Reductive dehalogenations are not only mediated by living anoxic microorganisms. The involvement of corrinoids, such as vitamin B12, has been shown and there are strong indications that PCDD/F are also dehalogenated by such reaction systems in the absence of biologically active microorganisms (Adriaens et al. 1996).
Fungal degradation of diaryl ether structures Degradation of diaryl ethers by P-450 monooxygenases Fungi are able to hydroxylate aromatic diaryl ether compounds directly or upon hydrolysis of epoxides formed as intermediates by cytochrome P-450 oxidase, a haemoprotein that they share with other higher eukaryotic organisms as well as prokaryotes. P-450 is used by all of these organisms for detoxi®cation of aromatics and xenobiotics and for biosynthesis and/or transformations of steroids, for example (reviewed by Guengerich 1990; Sariaslani 1991). The use of microbial cytochrome P-450 systems for biodegradation and bioremediation of pollutants has been reviewed and discussed by Guengerich (1995). Kellner et al. (1997) have summarized the potential for the engineering of this enzyme for bioremediation purposes in a very recent review paper on new developments in protein engineering. This is of interest since a promising mechanistic model for cytochrome P-450-mediated cleavage of DE and derivatives has been proposed by Ohe et al. (1994). Many natural and xenobiotic compounds are subjected to oxidation by fungal cytochrome P-450 oxidases. Reports on diaryl ether degradation that can unambiguously be attributed to fungal P-450 enzyme activity, however, are negligible. Although no direct proof has been published, the involvement of P-450 enzymes has to be assumed because of its presence in other strains of the genera Aspergillus, Cunninghamella and Trichosporon capable of hydroxylating similar aromatic structures. A review of fungal P-450 activities towards organic compounds in general has been written by Sariaslani (1991). The oxidation of DF by Cunninghamella elegans yielded the stable 2,3-dihydrodiol to which the trans con®guration was assigned. Only under rigid conditions did the compound decay into 2-hydroxy- and 3-hydroxydibenzofuran and water (Cerniglia et al. 1979). Cell extracts also catalyzed the oxidation of the trans dihydrodiol to the corresponding 2,3-dihydroxydibenzofuran. One has to assume that the initial product of the oxidation is an epoxide, which, upon hydrolysis by epoxide hydrolase or hydratase, furnishes this trans dihydrodiol. Hydroxylations of DE by Cunninghamella echinulata have been described by Seigle-Murandi (1991). This conversion predominantly furnished 4-hydroxydiphenyl ether and small yields of 4,4¢-dihydroxydiphenyl ether. The formation of sulphate ester conjugates has been discussed, since derivatives of structurally similar (di-)hydroxylated biphenyls were previously found after
oxidation of biphenyl by Aspergillus toxicarius (Golbeck et al. 1983). The oxidation of DE by the yeast Trichosporon beigelii resulted in a dierent set of hydroxylated products, probably because of the dierent regioselectivity and speci®city of the oxidizing enzyme system (Schauer, personal communication). Hydroxylations took place only at one of the aromatic rings, yielding all possible monohydroxylated DE. Only those substituted at positions 3 and/or 4 then underwent further hydroxylation to 3,4-dihydroxydiphenyl ether. The ®rst report on the cleavage of this dihydroxylated aromatic structure was by Schauer et al. (1995). Intra- or extradiol cleavage by a dioxygenase may have occurred, with subsequent oxidation of the aldehyde to the postulated derivative 2hydroxymuconic acid, which then lactonized chemically. The same strain was able to oxidize 4-CDE, DD and DF to hydroxylated derivatives. Further degradation to phenol was obtained through oxidation of DE via the same intermediates by the yeast Cryptococcus humicolus SBUG 517 (Schauer et al. 1993). It is not clear whether phenol was obtained through further degradation of a ring-cleavage product or by initial cleavage of the ether bridge by a monooxygenase. It is probably also true for fungi, as for hepatic microsomes, that the cleavage of the ether bond of DE by P-450 monooxygenase requires the presence of two highly polar substituents at the 4,4¢ position. This reaction yields (substituted) phenols as end-products (Ohe et al. 1994). The utilization, by a fungus of the genus Fusarium, of DF as the only carbon and energy source was reported by Hofmann et al. (1992). The site of attack on the DF as well as the mechanism of ring cleavage and further catabolism will represent an interesting feature of fungal catabolism of heteroaromatics and therefore remains to be elucidated. Fortuitous transformations by multiple hydroxylations of DF are widespread amongst yeast and ®lamentous fungi (Hammer and Schauer 1997). Degradation by peroxidases of white-rot fungi The second, free-radical oxidation system is predominantly found within white-rot fungi. In contrast to the P-450 system, these peroxidases are highly unspeci®c in the chemical structure of their organic substrates. Their role in the life of white-rot fungi is in the oxidation and depolymerisation of the lignin polymer in order to gain access to cellulose, which is used by the fungus for energy and biomass production. The evolution of carbon dioxide from lignin is low. Owing to the low speci®city towards organic compounds, xenobiotics such as nitroaromatic explosives, polyhalogenated ¯ame retardants, pesticides of dierent structures, and many other aliphatic and aromatic compounds such as PCDD/F are also attacked and degraded to some extent. An excellent mechanistic overview was published by Barr and Aust (1994), irrespective of its focus on 2,4,6-trinitrotoluene degradation.
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The ®rst study on 2,3,7,8-TCDD degradation by the white-rot fungus Phanerochaete chrysosporium was by Bumpus et al. (1985). The evolution of carbon dioxide from 2,3,7,8-TCDD was less than 2.5% when recovery was calculated with respect to the initial substrate concentration. More detailed mechanistic studies of the reaction mechanism of the lignin peroxidase (ligninase) of the fungus were performed by Hammel et al. (1986). The authors identi®ed a cationic radical formed from DD. They found that 2-CDD was also attacked and postulated a reaction sequence for further oxidation of aromatic molecules using pyrene as a model for anellated aromatic compounds. By subsequent abstraction of a single electron along with a proton and addition of water, followed by elimination of a pair of electrons and protons, hydroxylated and dihydroxylated derivatives are formed. They undergo further oxidation to quinone structures. In nitrogen-limited cultures of P. chrysosporium, degradation of numerous polycyclic aromatic hydrocarbons of a technical anthracene oil mixture, also containing DF and methylated DF, was demonstrated by Bumpus (1989). A mechanistic approach towards DD degradation, using H218O in the ligninase reaction, demonstrated cleavage of the ether bond as the next step following lateral hydroxylations. This led to the formation of catechol and higher hydroxylated derivatives. Some of these compounds might be used as additional carbon sources by the basidiomycete before being oxidized to more toxic quinones or before being repolymerized (Joshi and Gold 1994). Studies by Valli et al. (1992), who used the model compound 2,7-DCDD, which is signi®cantly less toxic than 2,3,7,8-TCDD, clearly demonstrated the elimination of one of the chlorines from the parent molecule and the concomitant formation of 4-chlorocatechol which, to a certain extent, was also dehalogenated. Dimethyl ethers formed as intermediates from 4-chlorocatechol, were subsequently demethylated again by the lignin peroxidase or manganese-dependent peroxidase of P. chrysosporium. This radical reaction furnished 4-chlorocatechol again, plus traces of the corresponding chloroquinone, and a 1,2,4-trihydroxybenzene monomethyl ether. Recent studies by Takada et al. (1996) were performed with 2,3,7,8-TCDD/F and higher halogenated congeners as target compounds for the peroxidases produced by the mycelium of a Phanerochaete sordida strain. They resulted in signi®cant degradation rates and metabolite formation. Initial concentrations of about 50 ng/l were degraded from approximately 10% to about 60% within 7 days. In the case of higher concentrations of both 2,3,7,8-TCDD and octachlorodibenzo-p-dioxin (5 parts per billion) only 10% were degraded under similar conditions within 10 days. In the culture medium 4,5-dichloro- and tetrachlorocatechol were identi®ed. An interesting feature of oxygen-radical-mediated polymerization is the coupling of two or more molecules of 2-hydroxydibenzofuran by laccases of the white-rot fungi Trametes versicolor and Pycnoporus cinnabarinus (Jonas et al. 1998).
Degradation of dioxins and dioxin-like compounds in soils and sediments Arthur and Frea (1989) discussed some features and conditions relevant to the eective degradation of dioxins in soil systems with regard to their biodegradation and, consequently, for bioremediation. PCDD/F in the environment are often bound to soil organic matter and surfaces of particles. The reduced bioavailability of (halogenated) DD/F, therefore, is of interest and was examined by Harms and Zehnder (1994, 1995). Investigations by Parsons (1992) had demonstrated that biodegradation rates for mono- to trihalogenated DD were lower in sediment slurries than in sediment-free solution. The adsorbed PCDD were readily desorbed and degraded through co-oxidation. Unchlorinated DD and DF were found to be readily and completely mineralized in soil upon the addition of speci®c biocatalysts (Figge et al. 1991, 1993). It was shown very recently (Megharaj et al. 1997) that Sphingomonas sp. strain RW1 grew on these compounds in soil. For a short period of time the strain grew exponentially at relatively high concentrations of DD and DF when preadapted prior to inoculation. Adaptation, by preculturing the strain in soil extract medium, prolonged the survival of this strain in soil. Studies by Halden et al. (1997) con®rmed the degradation of DD and DF in soil by strain RW1. Biotransformation in soil of 2-CDD led to signi®cantly reduced survival of the strain. Earlier depletion studies with 2,3,7,8-TCDD in laboratory soil systems and model ecosystems have showed some decline of the concentration of the target compound which, at least partially, was attributed to microbial activity (Kearny et al. 1972; Philippi et al. 1981, 1982). Degradative activities towards DD seem to be present in soil in general and were shown to be enhanced by the addition of active cultures of white-rot fungi or of organic substrates stimulating microbial activities (Rosenbrock et al. 1997). An estimation of the microbial potential for bioremediation of dioxin-polluted soils has been published recently by Halden and Dwyer (1997). Many soils contaminated with PCDD/F often contain high amounts of chlorophenols, which can be destroyed by the activity of fungal peroxidases already used as a tool for bioremediation (Chung and Aust 1995; Dec and Bollag 1990). Chlorophenols, however, can undergo polymerization into dioxins in the presence of such enzyme systems (OÈberg et al. 1990; Svenson et al. 1989; Wagner et al. 1990).
Conclusions and prospects The attack by bacterial angular dioxygenases seems to be restricted to all mono- and some dichlorinated DD and DF. Extension of the substrate range of DD- and DF-mineralizing bacteria can probably be achieved by
497
mutagenesis of the catalytically active a subunit of the dioxygenase. The stable incorporation into the bacterial chromosome of the regulated chlorocatechol pathway operon structure is currently being investigated in our laboratory and should allow the mineralization of halogenated salicylates and catechols from co-oxidative turnover of PCDD/F. An optimistic strategy for the degradation of 2,3,7,8-TCDD by such enzyme systems and other constructs was outlined by Omori (1995). However, the author did not take into account the necessity for productive mineralization of highly halogenated intermediates that are not completely degradable to energy-yielding Krebs-cycle intermediates in those organisms known up to now. Therefore, the organisms may be useful for overnight co-oxidation experiments in the laboratory only, but not in ®eld trials requiring the long-term performance of these ®eld-application vectors. More ecient hybrid organisms have to be constructed in the laboratory if one cannot isolate them from the environment. At present a process based on anaerobic dehalogenation of PCDD/F and subsequent degradation of less halogenated products in aerobic environments by specialized biocatalysts may be less expensive than other technical processes. Provided that biodegradation rates can be increased signi®cantly by the use of (bio-)emulsi®ers, the increased bioavailability of the target compounds may lead to more ecient degradation rates. Acknowledgements A part of this work was supported by grants from the Bundesministerium fuÈr Bildung, Wissenschaft, Forschung und Technologie.
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