Bacterial Degradation of Aromatic Compounds

Biodegradation is a viable bioremediation technology for organic pollutants. It has long known that microorganisms degrade environmental pollutants in various matrices and environments. Bioremediation utilizes the metabolic versatility of microorganisms to degrade hazardous pollutants. A goal of bioremediation is to transform organic pollutants into harmless metabolites or mineralize the pollutants into carbon dioxide and water [ 1 ]. A feasible remedial technology requires microorganisms being capable of quick adaptation to and efficient uses of pollutants of interest in a particular case in a reasonable period of time. Many factors influence microorganisms to use pollutants as substrates or cometabolize them. Therefore, understanding catabolic pathways and mechanisms and responsible enzymes is an effective means to define important factors for efficient cleanup of pollutants. Research has been conducted to understand bioremediation for environmental pollutants such as aromatic compounds that are among the most prevalent and persistent environmental pollutants. Biodegradation is a very broad field and involves uses of a wide range of microorganisms to break chemical bonds. It has been well reviewed [ 1 , 2 ], however, it is a very active field and new data are rapidly contributed to the literature. This review is focused on bacterial catabolic pathways of selected aromatic pollutants under aerobic culture conditions ( Table 1 ). The selected aromatic pollutants include the PAHs naphthalene, fluorene, phenanthrene, fluoranthene, pyrene, and benzo[ a ]pyrene, the heterocycles dibenzofuran, carbazole, dibenzothiophene, and dibenzodioxin, and alkylated PAHs. Metabolomics and proteomics in elucidation of mechanisms of microbial degradation of aromatics are also briefly discussed.

3.1. Naphthalene Naphthalene has often been used as a model compound to investigate the ability of bacteria to degrade PAHs because it is the simplest and the most soluble PAH [ 24 ]. Therefore, information of bacterial degradation of naphthalene has been used to understand and predict pathways in the degradation of three- or more ring PAHs. Many bacteria that have been isolated and utilize naphthalene as a sole source of carbon and energy belong to the genera Alcaligenes , Burkholderia , Mycobacterium , Polaromonas , Pseudomonas , Ralstonia , Rhodococcus , Sphingomonas , and Streptomyces ( Table 1 ) [ 3 , 10 , 25 – 36 , 152 , 153 ]. Degradation of naphthalene starts through the multicomponent enzyme, naphthalene dioxygenase, attack on the aromatic ring to form cis -(1R, 2S)-dihydroxy-1,2-dihydronaphthalene ( cis -naphthalene dihydrodiol) ( Scheme 1 ) [ 24 , 37 ]. The cis -naphthalene dihydrodiol formed by naphthalene dioxygenase is subsequently dehydrogenated to 1,2-dihydroxynaphthalene by a cis -dihydrodiol dehydrogenase [ 24 , 25 ]. Subsequently, 1,2-dihydroxynaphthalene is metabolized to salicylate via 2-hydroxy-2 H -chromene-2-carboxylic acid, cis - o -hydroxybenzalpyruvate, and 2-hydroxy-benzaldehyde ( Scheme 1 ) [ 24 , 26 , 27 , 33 ]. Also, 1,2-dihydroxynaphthalene is nonenzymatically oxidized to 1,2-naphthaquinone [ 25 ]. Salicylate is typically decarboxylated to catechol, which is further metabolized by ring fission in meta - and ortho -pathways. Fuenmayor et al. [ 29 ] reported that salicylate is converted to gentisate by salicylate-5-hydroxylase. Recently, Jouanneau et al. [ 31 ] purified the salicylate 1-hydroxylase from Sphingomonas sp. strain CHY-1 and characterized its biochemical and catalytic properties. The bacterial degradation of naphthalene has been well characterized for the catabolic enzyme system encoded by the plasmid NAH7 in Pseudomonas putida G7 [ 27 , 37 , 38 ]. NAH7 has two operons that contain the structural genes for naphthalene degradation. One operon contains the gene for the upper catabolic pathway encoding the enzymes necessary for the conversion of naphthalene to salicylate. The second operon contains the gene for the lower catabolic pathway encoding the enzymes necessary for the metabolism of salicylate through the catechol meta-cleavage pathway to pyruvate and acetaldehyde [ 24 , 27 , 37 , 38 ]. The entire sequence structure of plasmid NAH7 (82,232 bp) was determined by Sota et al . [ 154 ]. Also, the complete 83,042 bp sequence of the circular naphthalene degradation plasmid pDTG1 from Pseudomonas putida NCIB 9816–4 was determined by Dennis and Zylstra [ 155 ]. Parales et al. [ 39 ] reported that aspartate 205 in the catalytic domain of naphthalene dioxygenase is a necessary residue in the major pathways of electron transfer to mononuclear iron at the active site.

Phenanthrene, a three aromatic ring system, is found in high concentrations in PAH-contaminated sediments, surface soils, and waste sites [ 49 ]. Bacterial degradation of phenanthrene has been extensively studied. A variety of bacterial strains in Acidovorax , Arthrobacter , Brevibacterium , Burkholderia , Comamonas , Mycobacterium , Pseudomonas , and Sphingomonas have been isolated and have the ability to utilize phenanthrene as a sole carbon and energy source ( Table 1 ) [ 10 , 26 , 32 , 34 , 36 , 40 , 49 – 59 ]. Phenanthrene contains bay- and K-regions able to form an epoxide, which is suspected to be an ultimate carcinogen [ 56 , 60 ]. Therefore, it is used as a model substrate for studies on the catabolism of bay- and K-region containing carcinogenic PAHs such as benzo[ a ]pyrene, benzo[ a ]anthracene, and chrysene [ 56 ]. In general, bacterial degradation of phenanthrene is initiated by 3,4-dioxygenation to yield cis -3,4-dihydroxy-3,4-dihydrophenanthrene, which undergoes enzymatic dehydrogenation to 3,4-dihydroxyphenanthrene ( Scheme 3 ) [ 57 , 58 , 61 ]. The diol is subsequently catabolized to naphthalene-1,2-diol through both ortho -cleavage to form 2-(2-carboxy-vinyl)-naphthalene-1-carboxylic acid and meta -cleavage to form 4-(1-hydroxy-naphthalen-2-yl)-2-oxo-but-3-enoic acid [ 57 ]. Recently, Seo et al. [ 57 ] elucidated that the ortho -cleavage product, 2-(2-carboxy-vinyl)-naphthalene-1-carboxylic acid, is degraded to naphthalene-1,2-diol through naphthalene-1,2-dicarboxylic acid and 1-hydroxy-2-naphthoic acid. Pagnout et al. [ 62 ] isolated and characterized a gene cluster involved in phenanthrene degradation by 3,4-phenanthrene dioxygenation and meta -cleavage. It is possible that 1,2-dioxygenation of phenanthrene forms cis -1,2-dihydroxy-1,2-dihydrophenanthrene, which undergoes enzymatic dehydrogenation to 1,2-dihydroxyphenanthrene ( Scheme 3 ). This diol is also subsequently catabolized to naphthalene-1,2-diol through both ortho - and meta -cleavages. In general, phenanthrene-1,2- and 3,4-diols mainly undergo meta -cleavage due to the rapid accumulation of 5,6- and 7,8-benzocoumarin [ 57 ]. Naphthalene-1,2-diol converged from 1-hydroxy-2-naphthoic acid and 2-hydroxy-1-naphthoic acid is further degraded in a phthalic acid pathway through ortho -cleavage and a salicylic acid pathway through meta -cleavage [ 55 , 57 , 58 ]. Mallick et al. [ 63 ] reported that a novel meta -cleavage of 2-hydroxy-1-naphthoic acid to form trans -2,3-dioxo-5-(2’-hydroxyphenyl)-pent-4-enoic acid in Staphylococcus sp. PN/Y. Interestingly, phenanthrene degradation starts from 9,10-dioxygenase to yield phenanthrene cis -9,10-dihydrodiol that is further catabolized to 2,2’-diphenic acid via 9,10-dihydroxyphenanthrene [ 49 , 57 , 61 ].

Pyrene possessing four benzene rings is a byproduct of gasification processes and other incomplete combustion processes. Many bacterial isolates capable of degrading pyrene have been studied ( Table 1 ). Mycobacterium as Gram-positive species has been most widely studied for degrading pyrene by using it as a sole carbon and energy source [ 4 , 34 , 35 , 51 , 61 , 75 , 78 , 79 , 83 – 85 ]. Mycobacterium spp. are known to have high cell surface hydrophobicity and adhere to the emulsified solvent droplets [ 83 ]. Other pyrene degrading strains isolated include Rhodococcus sp. [ 86 ], Bacillus cereus [ 87 ], Burkholderia cepacia [ 80 ], Cycloclasticus sp. P1 [ 88 ], Pseudomonas fluorescens [ 64 ], Pseudomonas stutzeri [ 87 ], Sphingomonas sp. VKM B-2434 [ 26 ], Sphingomonas paucimobilis [ 89 ], and Stenotrophomonas maltophilia [ 64 , 66 , 90 ]. Heitkamp et al. [ 76 ] found the three products of ring oxidation, pyrene- cis -4,5-dihydrodiol, pyrene- trans -4,5-dihydrodiol, and pyrenol, and four products of ring fission ( Scheme 5 ), 4-hydroxyperinaphthenone, 4-phenanthroic acid, phthalic acid, and cinnamic acid by multiple analyses, including UV, infrared, mass spectrometry, NMR, and GC. The formation of pyrene- cis -4,5-dihydrodiol by dioxygenase and pyrene- trans -4,5-dihydrodiol by monooxygenase suggested multiple initial oxidative attacks on pyrene. Pyrene-1,2-diol derived from the dioxygenation at pyrene 1,2-C positions is metabolized to 4-hydroxyperinaphthenone via cis -2-hydroxy-3-(perinaphthenone-9-yl)-propenic acid and 2-hydroxy-2 H -1-oxa-pyrene-2-carboxylic acid ( Scheme 5 ). Kim and Freeman [ 61 ] found 1,2-dimethoxypyrene as a subproduct of pyrene-1,2-diol. Pyrene-4,5-diol is degraded by ortho -cleavage to phenanthrene-4,5-dicarboxylic acid, which is further metabolized through phenanthrene-3,4-diol pathway and 6,6’-dihydroxy-2,2’-biphenyl-dicarboxylic acid pathway. Meta -cleavage of pyrene-4,5-diol lead to 5-hydroxy-5 H -4-oxa-pyrene-5-carboxylic acid via 2-hydroxy-2-(phenanthrene-5-one-4-enyl)-acetic acid ( Scheme 5 ). A novel metabolite, 6,6’-dihydroxy-2,2’-biphenyl-dicarboxylic acid, was identified by Vila et al. [ 85 ] from the degradation of pyrene by Mycobacterium sp. strain AP1. Liang et al. [ 91 ] reported pyrene-4,5-dione formation and identified almost all the enzymes required during the initial steps of pyrene degradation in Mycobacterium sp. KMS. Kim et al. [ 92 ] identified 27 enzymes necessary for constructing a complete pathway for pyrene degradation from both genomic and proteomic data.

Various types of heterocycles containing oxygen, nitrogen, and sulfur are found in the environment, originating from anthropogenic or natural sources. Dibenzofurans, dibenzodioxins, and dibenzothiophene are among the most important environmental pollutants and are well reviewed [ 2 , 101 , 102 ]. Therefore, only aerobic bacterial degradation of non-halogenated heterocyclic aromatics is briefly discussed in this review to make a general relation with PAH degradation. Many bacterial species have been reported to decompose dibenzofurans and carbazole, a structural analogue with nitrogen instead of oxygen [ 103 – 108 ], although some white rot fungi are well known to decompose halogenated dibenzofurans. Guo et al. [ 106 ] isolated a stable carbazole-degrading microbial consortium consisting of Chryseobacterium sp. NCY and Achromobacter sp. NCW. As for the degradation of unsubstituted PAHs, bacterial catabolism of dibenzofurans starts at insertion of two oxygen atoms catalyzed by dioxygenases. Although many bacterial PAH dioxygenases are evolutionarily related to phenylpropionate dioxygenase, striking diversities have been reported in catalytic activity, mechanisms, regulations, and substrates of these important enzymes. In short, initial reactions of dibenzofuran and carbazole can be classified into angular and lateral dioxygenation, which may be catalyzed by different enzymes ( Scheme 7 ). These enzymes were found in Gram-positive and negative bacteria [ 109 – 112 ]. Some bacterial dioxygenases from Pseudomonas sp. CA10 and Sphingomonas wittichii RW1, for example, catalyze predominantly angular insertion of oxygen [ 109 , 112 ] while the commonly known naphthalene dioxygenase from Pseudomonas sp. NCIB 9816-4 catalyzes only lateral dioxygenation [ 113 ]. According to the metabolite analyses accompanied by biochemical studies, angular and lateral dioxygenases were considered from different origins. However, recent studies with cloned dioxygenase from Norcardioides aromaticivorans IC177 revealed that some dioxygenase can catalyze both reactions [ 110 ]. It is well known that PAH dioxygenases can catalyze various reactions, including reduction, mono- and di-oxygenation [ 113 ]. In addition to the multiple reactions with specific dioxygenases, current genomic or proteomic research with several PAH degrading bacteria (e.g., Burkholderia spp. and Mycobacterium spp.) revealed that multiple dioxygenases, probably playing different roles in PAH degradation, exist in a single bacterium [ 91 , 114 , 115 ]. For example, metabolite profiling of culture supernatant of Janibacter sp. YY1 treated with dibenzofuran showed that it may have both angular and lateral dioxygenases [ 116 ]. Multiple catabolic pathways can produce various metabolites with structural diversities, which probably are metabolized by different types of enzymes. For example, Nojiri et al. [ 107 ] reported that Terrabacter sp. DBF63 can metabolize fluorene and dibenzofuran by the same dioxygenase but the metabolites, namely phthalate and salicyclate from fluorene and dibenzofuran, respectively, are further catabolized by different enzymes, of which the related genes are located in different areas of chromosomes. Dioxygenases involving in dibenzo- p -dioxins degradation catalyze predominantly an angular dioxygenation to produce trihydroxy diphenyl ethers ( Scheme 8 ). PCDDs are well known pollutants with extreme toxicity, especially tetra to hexachloro analogues. Some bacterial species can catabolize up to hexachlorodibenzo- p -dioxins [ 112 , 117 ]. In comparison with dibenzofuran catabolism, chlorophthalates or salicylates and chlorophenols are common metabolites of PCDDs [ 112 ]. Chlorophenols are well known to be toxic to various organisms, including their degrading bacteria. Toxic effects of chlorometabolites are also reported from the catabolism of PCBs [ 118 ]. Although it is not clear how the bacteria can cope with the toxicity of chlorophenols during dioxin metabolism, glutathione or other sulfur-containing primary metabolites may be involved in its detoxification [ 118 ]. Dibenzothiophene and its higher molecular weight analogues are commonly found in higher molecular weight fractions of petroleum, but are not pyrogenic and biogenic PAH components. Although the chemical profiles of sulfur heterocycles are well documented in petroleum products, studies of their biodegradation in petroleum contaminated natural environments are very limited. Because of the economic importance of desulfurization of unprocessed petroleum, detailed research was performed with various bacterial species [ 119 – 121 ]. Catabolism of dibenzothiophene is catalyzed by distinct enzymes in two pathways ( Scheme 9 ). The catabolic branch of initial sulfur oxidation is also called 4S pathway, through which rapid desulfurization can be obtained. Flavin-containing monooxygenases, noted as DszA and DszB (or SoxA and SoxB), are widely distributed in bacteria and catalyze a consecutive

Proteomics and metabolomics have been recently employed in studies of environmental microbiology and have shown their high impact on the field of biodegradation and bioremediation [ 158 , 159 ]. Proteomics is an effective technique to identify proteins and their functions involved in the biodegradation of aromatics while metabolomics can be used to profile degradation products of PAHs and primary metabolites in response to PAH exposures. Intent of a brief discussion of proteomics and metabolomics here is to bring attention to these emerging fields rather than offering a comprehensive review and a long list of references. A good number of genomic sequences or expressed sequence tags (ESTs) of bacteria are currently available. These include several PAH degrading bacteria in the genus of Mycobacterium , Acinetobacter , Arthrobacter , and Burkholderia . Detailed transcriptomes in Burkholderia xenovorans LB400 during PCB catabolism were analyzed by microarray techniques [ 114 , 133 – 135 ]. The results revealed that large changes at genomic and proteomic levels occurred during the PAH catabolism, compared with the catabolism of natural carbon sources. These include up-regulation of the genes related to a) PAH catabolism, b) removal of reactive oxygen species, c) primary carbon metabolism, including one-carbon metabolism, d) energy production, e.g., ATP synthesis, e) various transporters, and f) synthesis of energy reservoir, e.g., polyphosphate. Many of these phenomena have been demonstrated by specific studies. For example, polyphosphate accumulation during PCB catabolism was confirmed by electron microscopy [ 136 ]. The accumulation of polyphosphate and polyphosphate kinase ( Ppk ) under stressed conditions are common in all biota. Up-regulation of superoxide dismutase or catalase/peroxidase is also common phenomena during PAHs or toxic chemical catabolism. The activation of one-carbon metabolism is another common phenomenon in bacteria during xenobiotic metabolism [ 137 ]. The exact mechanism of the gene activation is not clear. The up-regulation of one-carbon metabolism may be induced by several factors, including limitation of primary metabolite pools, introduction of one carbon unit, e.g., formate in PAH metabolism, and the loss of concerted regulation of general metabolism. Another common phenomenon is the changes of fatty acid content or composition in bacteria [ 138 ]. However, transcriptome analysis usually did not show a significant change in fatty acid metabolism-related genes. Discrepancies between genomic and proteomic data also occur. In general, proteomic analyses in relation to aromatics catabolism are very limited. However, a few proteomic studies of PAH catabolism in Mycobacterium spp. have been reported [ 61 , 68 , 91 , 115 , 139 ]. For example, Lee et al. [ 68 ] studied fluoranthene catabolism and associated proteins in Mycobacterium sp. JS14. An increased expression of 25 proteins related to fluoranthene catabolism is found with 1D-PAGE and nano liquid chromatography tandem MS/MS. Detection of fluoranthene catabolism associated proteins coincides well with its multiple degradation pathways that are mapped via metabolites identified. Kim et al. [ 115 ] reported the protein profiles of Mycobacterium vanbaalenii PYR-1 grown in the presence of pyrene, pyrene-4,5-quinoline, phenanthrene, anthracene, and fluoranthene after two dimensional electrophoresis (2DE) separation of proteins and mass spectrometry analysis. They found PAH-induced proteins (catalase-peroxidase, putative monooxygenase, dioxygenase small subunit, small subunit of naphthalene-inducible dioxygenase, and aldehyde dehydrogenase), carbohydrate metabolism related proteins (enolase, 6-phosphogluconate dehydrogenase, indole-3-glycerol phosphate synthase, and fumarase), DNA translation proteins, heat shock proteins, and energy production protein (ATP synthase). Proteomics studies showed that several Mycobacterium species have multiple dioxygenases and related enzymes, which may be involved in different substrates [ 61 , 68 , 91 ]. Those studies also revealed that many of stress related proteins, including enzymes related to reactive oxygen removal and transcriptional regulators are commonly up-regulated during PAH metabolism. The workflow of proteomics has been extended from qualitative analyses after 1DE or 2DE to quantitative studies. Systematic quantitation of proteomes has become common by using methods such as isotope-coded affinity tag (ICAT), isobaric tags for relative and absolute quantitation (iTRAQ), and stable isotope labeling by essential amino acid culture (SILAC). The labeled proteins or peptides by those methods are analyzed by tandem MS [ 140 , 141 ]. Recently, Kim et al. [ 142 ] conducted the proteome analysis of Pseudomonas putida KT2440 induced with monocyclic aromatic compounds using 2-DE/MS and cleavable isotope-coded affinity tag (ICAT) to determine the changes of proteins involved in aromatic degradation pathway