Bacteria, which have evolved more than three billion years ago, have developed strategies for obtaining energy from virtually every compound and have been considered as nature’s ultimate scavengers. Because of their quick adaptability, bacteria have largely been used to degrade or remediate environmental hazards. Various bacteria have been found to degrade PAHs, in which degradation of naphthalene and phenanthrene has been most widely studied. Numerous unique metabolic pathways for the bacterial degradation of PAHs have been well documented in a number of excellent review articles (Cerniglia, 1992; Peng et al., 2008; Seo et al., 2009; Mallick et al., 2011). So, the biochemical pathways for bacterial degradation of PAHs will not be discussed in details in this communication. Nevertheless, there are two major strategies to degrade PAHs depending on the presence or absence of oxygen. In the aerobic catabolism of aromatics, the oxygen is not only the final electron acceptor but also a co-substrate for the hydroxylation and oxygenolytic ring cleavage of the aromatic ring. In contrast, the anaerobic catabolism of aromatic compounds uses a completely different strategy to attack the aromatic ring, primarily based on reductive reactions (Foght, 2008; Carmona et al., 2009). While the aerobic catabolism of aromatic compounds has been studied for several decades, the anaerobic degradation of aromatic compounds is a more recently discovered microbial capacity that still awaits a deeper understanding. However, anoxic conditions dominate in many natural habitats and contaminated sites (e.g., aquifers, aquatic sediments and submerged soils) where biodegradation is carried out by anaerobes using alternative final electron acceptors such as nitrate, sulfate or ferric ions (Foght, 2008; Carmona et al., 2009).

Principally, bacteria favor aerobic conditions for degradation of PAHs via oxygenase-mediated metabolism (involving either monooxygenase or dioxygenase enzymes). Usually, the first step in the aerobic bacterial degradation of PAHs is the hydroxylation of an aromatic ring via a dioxygenase, with the formation of a cis-dihydrodiol, which gets rearomatized to a diol intermediate by the action of a dehydrogenase. These diol intermediates may then be cleaved by intradiol or extradiol ring-cleaving dioxygenases through either an ortho-cleavage or meta-cleavage pathway, leading to intermediates such as catechols that are ultimately converted to TCA cycle intermediates (Evans et al., 1965; Cerniglia, 1992; Eaton and Chapman, 1992; Mueller et al., 1996; Mallick et al., 2011). Dioxygenase is a multicomponent enzyme generally consisting of reductase, ferredoxin, and terminal oxygenase subunits (Mallick et al., 2011). Bacteria can also degrade PAH via the cytochrome P450-mediated pathway, with the production of trans-dihydrodiols (Sutherland et al., 1990; Moody et al., 2004) or under anaerobic conditions, e.g., under nitrate-reducing conditions (Foght, 2008; Carmona et al., 2009).

Naphthalene degrading bacteria are ubiquitous in nature and there are enormous numbers of reports documenting the bacterial degradation of naphthalene including the elucidation of the biochemical pathways, enzymatic mechanisms and genetic regulations (Cerniglia, 1992; Peng et al., 2008; Seo et al., 2009; Lu et al., 2011; Mallick et al., 2011). The naphthalene catabolic genes present in the plasmid NAH7 in Pseudomonas putida G7 are well characterized (Simon et al., 1993). In the plasmid NAH7, the naphthalene catabolic genes (nah) are organized into two operons: the nal operon containing the genes for the upper pathway enzymes involved in conversion of naphthalene to salicylate, and the sal operon containing the genes for the lower pathway enzymes involved in the conversion of salicylate to pyruvate and acetaldehyde (Simon et al., 1993). The operons are positively regulated by a common regulator NahR, a LysR type of positive transcriptional regulator and are widely dispersed in bacteria. NahR is induced in presence of salicylate leading to high-level expression of the nah genes in bacteria (Yen and Gunsalus, 1985; Peng et al., 2008). There are several reports on the nucleotide sequences of genes encoding the upper pathway enzymes in different Pseudomonas strains, and the genes are more than 90% identical (Menn et al., 1993; Simon et al., 1993; Yang et al., 1994; Bosch et al., 2000; Peng et al., 2008). Moreover, in Ralstonia sp. U2, the naphthalene catabolic operon (nag) contains all of the upper pathway genes similar to that of the classical nah genes of Pseudomonas strains in the same order, with the exception of two extra genes named nagG and nagH, which are structural subunits of salicylate-5-hydroxylase enzyme required for the conversion of naphthalene to gentisate (Zhou et al., 2001). Additionally, in Comamonas testeroni strain GZ42, the naphthalene catabolic genes are similar to that found in Ralstonia sp. U2 (Goyal and Zylstra, 1997). The lower pathway genes for the naphthalene catabolism are also similar in different Pseudomonas strains like P. putida G7, NCIB9816-4, ND6, and P. stutzeri AN10 (Habe and Omori, 2003; Peng et al., 2008). In those organisms, lower pathway of naphthalene operon contains 11 genes present in the order nahGTHINLOMKJY, in which nahY represent a naphthalene chemotaxis gene. However, in AN10 and ND6 strains another salicylate hydroxylase gene (nahW) has been found to be present outside, but near to the sal operon.

Along with Pseudomonas, a high proportion of the PAH-degrading isolates belong to the sphingomonads, which comprise a physiologically versatile group within the Alphaproteobacteria: mainly Sphingomonas, Sphingobium and Novosphingobium that are frequently found as aromatic degraders. Species belonging to these genera show great catabolic versatility, capable of degrading a wide range of natural and xenobiotic compounds including HMW PAHs (Basta et al., 2005; Peng et al., 2008; Stolz, 2009; Vila et al., 2015). In a few reports, it has been shown that the catabolic versatility of sphingomonads relies on the large plasmids present in those organisms (Basta et al., 2005; Peng et al., 2008). However, the plasmid-encoded degradative genes have been found to be largely scattered, or they are not organized in coordinately regulated operons (Romine et al., 1999). It may be possible that this kind of ‘flexible’ gene organization i.e., different combinations of conserved gene clusters allows sphingomonads to adapt easily and proficiently to degrade various aromatic compounds including xenobiotics (Basta et al., 2005; Peng et al., 2008). Recently an illustrative report on the catabolic versatility of sphingomonads has shown that strain PNB, acting on diverse monoaromatic and polyaromatic compounds, contains seven sets of ring-hydroxylating oxygenases (RHO) with different substrate specificities (Khara et al., 2014). In general, it has been seen that mobile genetic elements (MGEs) like plasmids and transposons play a crucial role in the biodegradation of organic pollutants like PAHs. The presence of foreign compounds can often lead to the selection of mutant bacteria that are capable of metabolizing them. Apart from vertical gene transfer, aromatics catabolic genes are often harbored by MGEs that successfully disseminate the catabolic traits to phylogenetically diverse bacteria by horizontal gene transfer (Top and Springael, 2003; Nojiri et al., 2004).

Among PAHs degrading bacteria the genus Rhodococcus is very unique, having an enormous catabolic versatility. In contrast to Pseudomonas and other Gram-negative bacteria whose naphthalene catabolic genes are usually clustered, the Gram-positive Rhodococcus strains usually exhibit only three structural genes required for naphthalene degradation (narAa, narAb and narB) (Kulakov et al., 2005; Larkin et al., 2005). It has been seen that the nar region is not arranged into a single operon, and there are several homologous transcription units in different Rhodococcus strains separated by non-homologous sequences containing direct and inverted repeats (Larkin et al., 2005). The narAa and narAb genes codes for the α- and β- subunit of the naphthalene dioxygenase (NDO). Both subunits of NDO in Rhodococcus sp. strain NCIMB12038 showed only 30% amino acid identity to the corresponding P. putida NDO subunits. In addition, there is no gene in Rhodococcus strains which is similar to the genes encoding the electron transport components reductase and ferredoxin of NDO in Pseudomonas strains (Kulakov et al., 2005; Larkin et al., 2005). Moreover, the narA and narB genes are transcribed as a single unit through different start sites, and their transcription is induced in the presence of naphthalene in contrast to salicylate-inducible naphthalene catabolic genes in Pseudomonas species. There are two putative regulatory genes (narR1 and narR2, GntR-like transcriptional regulator) which are shown to be transcribed as a single mRNA in naphthalene-induced cells (Kulakov et al., 2005; Larkin et al., 2005). Figure 3 represent the gene organization of the nar gene cluster of various naphthalene degrading Rhodococcus sp.

Along with naphthalene, a number of reports on phenanthrene degradation by various Gram negative and Gram positive bacterial species have been reported (Peng et al., 2008; Seo et al., 2009; Mallick et al., 2011). In a study, Mallick et al. (2007) reported the degradation of phenanthrene by Staphylococcus sp. strain PN/Y, by initiating the dioxygenation specifically at the 1,2-position followed by meta-cleavage of phenanthrene-1,2-diol, leading to the formation of 2-hydroxy-1-naphthoic acid as the metabolic intermediate; while the ortho-cleavage could yield the naphthalene-1,2-dicarboxylic acid. Authors also reported that 2-hydroxy-1-naphthoic acid was metabolized by a meta-cleavage enzyme 2-hydroxy-1-naphthoate dioxygenase leading to the formation of trans-2,3-dioxo-5-(2′-hydroxyphenyl)-pent-4-enoic acid, a novel metabolite in the phenanthrene degradation pathway, and was subsequently degraded via salicylic acid and catechol (Mallick et al., 2007). Later on, Ghosal et al. (2010) reported the assimilation of phenanthrene by Ochrobactrum sp. strain PWTJD, isolated from municipal waste contaminated soil sample using phenanthrene as a sole source of carbon and energy. The strain PWTJD could also degrade phenanthrene via 2-hydroxy-1-naphthoic acid, salicylic acid and catechol. The strain PWTJD was found to utilize 2-hydroxy-1-naphthoic acid and salicylic acid as sole carbon sources, while the former was metabolized by a ferric-dependent meta-cleavage dioxygenase. In the lower pathway, salicylic acid was metabolized to catechol and was further degraded by catechol 2,3-dioxygenase to 2-hydroxymuconoaldehyde acid, ultimately leading to TCA cycle intermediates. The metabolic pathway involved in the degradation of phenanthrene by Ochrobactrum sp. strain PWTJD is shown in Figure 4 (Ghosal et al., 2010). This was the first report of the complete degradation of a PAH molecule by Gram-negative Ochrobactrum sp. describing the involvement of the meta-cleavage pathway of 2-hydroxy-1-naphthoic acid in phenanthrene assimilation.

Other LMW PAHs like anthracene, fluorene, acenaphthene and acenaphthylene are also found in high quantities in PAH-polluted sites and various bacterial species have the capability to utilize these compounds individually as sole carbon and energy sources (Moody et al., 2001; Peng et al., 2008; Seo et al., 2009; Mallick et al., 2011). Lately, Ghosal et al. (2013) reported assimilation of acenaphthene and acenaphthylene by the Acinetobacter sp. strain AGAT-W, isolated from municipal waste contaminated soil sample using acenaphthene as a sole source of carbon and energy. The strain AGAT-W could degrade acenaphthene via 1-acenaphthenol, naphthalene-1,8-dicarboxylic acid, 1-naphthoic acid, salicylic acid and cis, cis muconic acid ultimately leading to TCA cycle intermediates. The metabolic pathways involved in the degradation of acenaphthene and acenaphthylene by Acinetobacter sp. strain AGAT-W is shown in Figure 5 (Ghosal et al., 2013). This was the first report on the complete degradation of acenaphthene and acenaphthylene individually by strain AGAT-W belonging to the genus Acinetobacter.

Polycyclic aromatic hydrocarbons with more than three rings viz. pyrene, benzo[a] pyrene (Bap), are generally referred to as HMW PAHs (Figure 1), which are of significant environmental concern due to their long persistence and high toxicity as well as their mutagenic and/or carcinogenic properties (Cerniglia, 1992; Kanaly and Harayama, 2000; Peng et al., 2008; Seo et al., 2009). In the last few decades research on microbial degradation of HMW PAHs has advanced significantly and a number of HMW PAH-degrading isolates have been reported (Cerniglia, 1992; Kanaly and Harayama, 2000, 2010; Peng et al., 2008; Seo et al., 2009). Biodegradation of HMW PAHs by microorganisms is discussed adequately in many excellent reviews and the pathways for their degradation have also been depicted (Kanaly and Harayama, 2000; Peng et al., 2008; Seo et al., 2009). However, further investigations are prerequisite in several areas of HMW PAH biodegradation, namely, research related to the regulatory mechanisms of HMW PAH biodegradation, biodegradation of HMW PAH associated with other hydrocarbons in mixtures; and the interactions of complex microbial community during HMW PAH degradation can be exploited in more details. These will enrich our understanding on the microbial ecology of HMW PAH-degrading communities and the mechanisms by which HMW PAH biodegradation occur. In addition, the outcome will also help in predicting the environmental fate of these recalcitrant compounds and aid for the development of convenient as well as cost-effective bioremediation strategies in near future.

Environmental pollution due to anthropogenic activity has spread to all types of ecosystems and marine ecosystem is not excluded from this list. Hypersaline environments are regularly being polluted with organic pollutants including PAHs, through industrial and municipal effluents. Industrial effluents, specifically, petroleum industry where one of the main sources of contaminants in the waste waters are aromatic hydrocarbons including PAHs. Contamination and biodegradation in extreme environments like high salinity has been receiving increased attention in recent times (Oren et al., 1992; Margesin and Schinner, 2001a,b; Le Borgne et al., 2008; Bose et al., 2013; Cravo-Laureau and Duran, 2014; Elango et al., 2014; Fathepure, 2014; Genovese et al., 2014; Kappell et al., 2014; Kostka et al., 2014; Thomas et al., 2014; Torlapati and Boufadel, 2014). In addition, a survey of relevant literatures indicates that along with numerous biotechnological applications, halophilic microorganisms have more extended catabolic versatility than previously thought about (Margesin and Schinner, 2001b; Garcia et al., 2005; Smith et al., 2013; Kappell et al., 2014; Lamendella et al., 2014; Roling and Bodegom, 2014; Scott et al., 2014; Singh et al., 2014).

The biological treatment of industrial hypersaline waste waters and the bioremediation of polluted hypersaline environments are not possible with conventional microorganisms because they are incapable to function efficiently at salinities that of seawater or above. Thus, halophilic microorganisms are the best alternatives to overcome this problem (Oren et al., 1992; Garcia et al., 2005; Fathepure, 2014; Kostka et al., 2014). In the last decade there has been an increasing interest in the development and optimization of bioremediation processes via halophiles to deal with hypersaline environments that are contaminated with organic pollutants (Margesin and Schinner, 2001a; Mellado and Ventosa, 2003; Peyton et al., 2004; Garcia et al., 2005; Cui et al., 2008; Mnif et al., 2009; Zhao et al., 2009; Bonfa et al., 2011; Smith et al., 2013; Fathepure, 2014; Kappell et al., 2014; Lamendella et al., 2014; Singh et al., 2014; Thomas et al., 2014). Some halophiles that have shown PAHs degrading property are depicted in Supplementary Table S1. However, studies concerning the ability of this group of microbes to degrade PAHs are still in their infancy. Although in the last few years, a few studies on the degradation of PAHs by halophiles have been published. In this context, Gauthier et al. (1992) described the isolation of Marinobacter hydrocarbonoclasticus which can degrade various organic compounds including naphthalene (Gauthier et al., 1992). The presence of naphthalene dioxygenase, similar to those from Pseudomonas and Burkholderia spp. was reported in a naphthalene-degrading Marinobacter sp. strain NCE312 (Hedlund et al., 2001). Again, based on the analysis of the shoreline sand and rocks (Costa da Morte, northwestern Spain) affected by the Prestige oil spill, unveiled the high relative abundances of Sphingomonadaceae and Mycobacterium that could be associated with PAH degradation (Alonso-Gutierrez et al., 2009). In a similar study, Vila et al. (2010) reported the isolation of a marine microbial consortium from a beach polluted with the Prestige oil spill which is efficient in removing different hydrocarbon present in heavy fuel oil including three to five-ring PAHs (for example anthracene, fluoranthene, pyrene, benzo(a)anthracene, chrysene, and BaP) (Vila et al., 2010). In addition based on community dynamics analysis, it has been contemplated that Alphaproteobacteria (Maricaulis and Roseovarius), could be associated with the degradation of LMW PAHs, whereas Gammaproteobacteria (Marinobacter and Methylophaga), could be associated with the degradation of HMW PAHs. A similar bacterial community associated with the degradation of various organic compounds was found in a beach affected by the Deepwater Horizon oil spill (Kostka et al., 2011; Lamendella et al., 2014). Recently, Gallego et al. (2014) reported the isolation of a marine pyrene-degrading microbial consortium from a beach polluted by an oil spill and found that an uncultured Gordonia sp. is the key pyrene degrader in the consortium based on community structure and PAH ring-hydroxylating genes analyses (Gallego et al., 2014). Nevertheless, detailed information on the bioremediation of PAHs under high salinity by halophilic/halotolerant bacteria and archaea is still in its initial stage of exploration, but it is expected that this emerging field is ready to flourish in near future.

The community analysis of indigenous microorganisms capable of degrading various aromatic compounds in diverse environments has been of great interest in the recent years (Watanabe et al., 2002; Brakstad and Lodeng, 2005; Gerdes et al., 2005; Yakimov et al., 2005; Coulon et al., 2007; Cui et al., 2008) and different strategies have been developed for the study of associated microbial communities (Widada et al., 2002). Previously it was thought that PAH degradation capabilities may be associated with certain genera or groups of bacteria independent of the origin of the source sample (Kastner et al., 1994; Mueller et al., 1997). But consequently, with time it has been reported that PAH-degrading bacteria are far more diverse (Widada et al., 2002; Andreoni et al., 2004; Cui et al., 2008; Hilyard et al., 2008). Currently, there are enormous reports of various bacterial as well as some archaeal genera capable of degrading PAHs. Supplementary Table S1 represents the bacterial/archaeal (halophilic) strains involved in the degradation of various PAHs. Figure 6 illustrates the 16S rRNA gene-based phylogenetic analysis of all the bacterial/archaeal strains cited in Supplementary Table S1. It has been seen that bacterial/archaeal strains having PAHs catabolic property are distributed mainly in six major groups: Alphaproteobacteria, Betaproteobacteria, Gammaproteobacteria, Actinomycetes, Firmicutes and Archaea (Halophiles). It may be mentioned here that many strains listed in Supplementary Table S1 have not been cultured in laboratory. Some PAHs degrading isolates represent novel strains which are isolated from various geographically diverse habitats. Thus, all this information illustrate that PAH-degrading machinery is not confined to a few particular genus reported so far but is more likely to be distributed widely in the prokaryotic kingdom found in diverse geographical niches.

Polycyclic aromatic hydrocarbons metabolic pathway of non-ligninolytic fungi is similar to those formed by mammalian enzymes. In this case, the predominant pathway of initial oxidation of PAHs by non-ligninolytic fungi involves the activity of the cytochrome P450 monooxygenase enzymes. These enzymes catalyze a ring epoxidation to form an unstable arene oxide, which is further transformed to trans-dihydrodiol via an epoxide-hydrolase catalyzed reaction (Jerina, 1983; Sutherland et al., 1995). Non-ligninolytic fungi such as Cunninghamella elegans and ligninolytic fungi such as Pleurotus ostreatus, metabolize PAHs via this pathway (Bezalel et al., 1996; Tortella et al., 2005). For example, the transformation of fluoranthene by C. elegans produce fluoranthene trans-2,3-dihydrodiol, 8- and 9-hydroxyfluoranthene trans-2,3-dihydrodiols (Tortella et al., 2005). Similarly, P. ostreatus metabolizes pyrene into pyrene trans-4,5-dihydrodiol (Bezalel et al., 1996). Arene oxide produced by cytochrome P450 can also be rearranged to phenol derivatives by non-enzymatic reactions and are subsequently conjugated with sulfate, xylose, glucoronic acid, or glucose (Mueller et al., 1996; Pothuluri et al., 1996). In some fungi, cytochrome P450 monooxygenases oxidize PAHs to epoxides and dihydrodiols which are potent carcinogens and more toxic than the respective parent PAHs, while on the other hand peroxidase-mediated oxidation of PAHs produces quinines which are less toxic than the parent PAHs (Cerniglia and Sutherland, 2010). This is why oxidation of PAHs by ligninolytic enzymes could be a more logical strategy for the detoxification of PAHs contaminated environment.

White-rot fungi are ubiquitous in nature and can produce ligninolytic enzymes which are secreted extracellularly. These enzymes can degrade lignin present in wood and other organic substances. Ligninolytic enzymes are mainly of two types, peroxidases and laccases. On the basis of reducing substrate types, peroxidase enzyme can be classified again into two types, lignin peroxidase and manganese peroxidase. Both types of peroxidases have the ability to oxidize PAHs (Hammel, 1995; Cerniglia and Sutherland, 2010). Laccases, the phenol oxidase enzymes, also have the ability to oxidize PAHs (Cerniglia and Sutherland, 2010). Compared to bacterial PAHs degrading enzymes, ligninolytic enzymes are not induced in the presence of PAHs or by its degradation products (Verdin et al., 2004). As ligninolytic enzymes are secreted extracellularly, they are able to diffuse toward the immobile PAHs, and that is why they are more useful than bacterial intracellular enzymes in making initial attack on PAHs in soil. In addition compared to bacterial PAH-degrading enzymes, ligninolytic enzymes have broad substrate specificity and are therefore able to transform a wide range of substrates, even those which are most recalcitrant (Tortella et al., 2005; Cerniglia and Sutherland, 2010). Ligninolytic enzymes can transform PAHs by producing hydroxyl free radicals by the donation of one electron, which oxidizes the PAH ring (Sutherland et al., 1995). As a result, PAH-quinones and acids are formed instead of dihydrodiols. It has been seen that ligninolytic fungi mineralize PAHs by a combination of ligninolytic enzymes, cytochrome P450 monooxygenases, and epoxide hydrolases (Bezalel et al., 1997). Numerous reports have been documented on the degradation of PAHs by white-rot fungi (Tortella et al., 2005; Cerniglia and Sutherland, 2010). The metabolic pathway for the degradation of phenanthrene by the ligninolytic fungus Pleurotus ostreatus is illustrated in Figure 7 (Bezalel et al., 1997). PAHs degradative potential of wood-rotting fungi Plrurotus ostreatus and Antrodia vaillantii was also examined in soil, artificially contaminated with fluorene, phenanthrene, pyrene, and benz[a]anthracene (Andersson et al., 2003). It has been reported that although P. ostreatus significantly increased the degradation of PAHs in soil, but in the process, accumulated toxic PAH metabolites. It has also been seen that this white-rot fungus inhibits the indigenous microbial populations in the soil, which may have prohibited the complete mineralization of the PAHs. Conversely, the brown-rot fungus A. vaillantii did not generate any dead-end metabolites, although its PAHs degradation rate was similar to that of P. ostreatus (Andersson et al., 2003). Oxidation of pyrene, anthracene, fluorene and BaP to the corresponding quinines by lignin peroxidase and manganese peroxidase was reported by the white-rot fungus Phanerochaete chrysosporium (Bogan et al., 1996a,b). In another study, complete mineralization of BaP (HMW PAHs) by the white-rot fungus P. chrysosporium was reported in a two-stage pilot scale reactor (May et al., 1997). Clemente et al. (2001) studied PAHs degradation by thirteen ligninolytic fungal strains and reported that the rate of degradation varies with a variation of lignolytic enzymes (Clemente et al., 2001). In this study, highest naphthalene degradation (69%) was reported by the strain 984 which have Mn-peroxidase activity, followed by strain 870 (17%) having lignin peroxidase and laccase activities. The ability of soil fungi to degrade PAHs that produce ligninolytic enzymes was also studied under microaerobic and very-low-oxygen conditions (Silva et al., 2009). It was reported that Aspergillus sp., Trichocladium canadense, and Fusarium oxysporum can degrade LMW PAHs (2–3 ring compounds) most extensively, whereas highest degradation of HMW PAHs (4–7 rings) was observed in T. canadense, Aspergillus sp., Verticillium sp., and Achremonium sp. These results suggest that, along with bacteria, fungi can be exploited as a valuable endeavor for the bioremediation of PAH contaminated sites.

Temperature has a profound effect on the biodegradation of PAHs in contaminated sites since those places are not always at ambient temperature for activity of the inhabitant microorganisms. When temperature increases, the solubility of PAHs also increases, which in turn increases the bioavailability of PAH molecules. On the other hand, with increasing temperature, dissolved oxygen level decreases, as a result of which the metabolic activity of aerobic mesophilic microorganisms reduces. There are some microorganisms called psychrophiles, which can operate at low temperature, and are in contrast to thermophiles that function at high temperature. At high temperature, some pollutants can get transformed into a new compound and often the daughter compound appears to be more toxic than the parent compound, which in turn inhibits its biodegradation rate (Müller et al., 1998). Generally in most of the studies biodegradation of PAHs have been examined under moderate temperatures, however, there are reports on PAHs biodegradation under extreme temperatures. For example, biodegradation of petroleum hydrocarbons including PAHs was reported in seawater at low temperatures (0–5°C) (Brakstad and Bonaunet, 2006) while the biodegradation of PAHs along with long chain alkanes was documented at very high temperature (60–70°C) by Thermus and Bacillus spp (Feitkenhauer et al., 2003).

pH also plays a significant role in biodegradation processes including that of PAHs. Usually, microorganisms are pH-sensitive and near neutral conditions (6.5–7.5) are favored by most of them for their normal activity. However, it has been seen that in many PAHs contaminated sites, pH is very far from neutral settings. For example, some abandoned gasworks sites frequently contain a huge amount of demolition waste like concrete and brick. Leaching of these materials can result in the increase of pH of the native soil. Moreover, the leaching of some material like coal spoil can generate an acidic environment due to the oxidation of sulfides. These acidic or alkaline conditions can create adverse conditions for microbial activities, which in turn decrease the biodegradation of PAHs in those sites. Hence, it is recommended to adjust the pH at those sites by adding chemicals: basic soils can be treated with ammonium sulfate or ammonium nitrate while acidic soils can be treated by liming with calcium or magnesium carbonate (Bowlen and Kosson, 1995) to generate an environment for effective biodegradation.

Biodegradation of organic contaminants including PAHs can proceed under both aerobic and anaerobic conditions; however, most of the studies focus bioremediation of PAHs under aerobic condition, where oxygen acts as a co-substrate and a rate limiting factor. During aerobic biodegradation of PAHs, oxygen is required for the action of both mono- and dioxygenase enzymes in the initial oxidation of the aromatic rings. For in situ bioremediation of contaminants, sometimes oxygen is added from an external source. The addition of oxygen externally may be achieved by drainage, tilling, and addition of chemicals that release oxygen or by injection of air into the contaminated site (Pardieck et al., 1992; Atlas and Unterman, 1999; Bewley and Webb, 2001). In a study, enhanced intrinsic bioremediation of BTEX compounds in aquifer using an oxygen releasing compound (magnesium peroxide; ORC^®) was reported (Odencrantz et al., 1996). Again, in a separate study, in situ bioremediation of an aquifer contaminated with various pollutants, including phenols, BTEX compounds and PAHs was reported, in which sodium nitrate was circulated through the aquifer as the oxygen source, via a series of injection and abstraction wells (Bewley and Webb, 2001).

Availability of nutrients is another rate-limiting factor for successful bioremediation of PAHs contaminated environments. Along with easily metabolized carbon source, microorganisms require various minerals like nitrogen, phosphorus, potassium, and iron for normal metabolism and growth. Thus, supplementation of nutrients, in the poor nutrient content pollutant contaminated sites is required to stimulate the growth of autochthonous microorganisms, to enhance bioremediation of pollutants (Atagana et al., 2003). In marine environments, poor biodegradation of petroleum hydrocarbon is due to a low level of nitrogen and phosphorous (Floodgate, 1984). On the other hand, excessive nutrient availability can also inhibit the bioremediation of pollutants (Chaillan et al., 2006). There are several reports on the negative effects of high nutrient levels on the biodegradation of organic pollutants especially PAHs (Carmichael and Pfaender, 1997; Chaîneau et al., 2005).

In case of a biological system, the term bioavailability is defined as the fraction of a chemical that can be taken up or transformed by living organisms during the course of the experiment. This fraction can vary under the influence of mass transfer parameters, which include physicochemical processes governing dissolution, desorption and diffusion, hydrological processes like mixing and, finally, biological processes, such as uptake and metabolism (Bosma et al., 1997; Semple et al., 2003). Bioavailability is regarded as one of the most crucial factors in bioremediation of pollutants. Biodegradation of PAHs in the environment is often limited, due to their low aqueous solubility and their strong tendency to sorbs to mineral surfaces (like clays) and organic matters (like black carbon, coal tar humic acids) in the soil matrix, processes which reduce bioavailability. There are numerous reports on poor or unsuccessful remediation of PAHs contaminated sites due to absorption of PAHs in black carbon, coal-tar substances which decreases bioavailability of PAHs (Volkering et al., 1992; Luthy et al., 1993, 1994; Nelson et al., 1996; Northcott and Jones, 2001; Hong et al., 2003; Bucheli et al., 2004; Cornelissen et al., 2005; Liu et al., 2009; Benhabib et al., 2010; Rein et al., 2016). Again, the aqueous solubility of PAHs decreases with increasing molecular weight, which in turn reduces the bioavailability of PAHs. It has been seen that longer any hydrophobic organic contaminants (like PAHs) is in contact with soil, the more irreversible the sorption, and lesser is the chemical and biological extractability of the contaminants, which is known as ‘aging’ of the contaminant (Alexander, 2000; Luo et al., 2012). It has been reported that extractability and bioavailability of PAHs decrease significantly with time in aging processes which can be significantly rate-limiting for in situ bioremediation (Alexander, 2000; Megharaj et al., 2002; Luo et al., 2012; Abdel-Shafy and Mansour, 2016).

The major goal of bioremediation is the removal or detoxification of contaminants from a given environment. However, sometimes it is possible that the contaminant gets transformed into a more toxic dead end product. This is why it is essential to confirm that the pollutant is completely mineralized at the end of the treatment (Mendonca and Picado, 2002; Lundstedt et al., 2003). A study using a bioreactor to treat PAH-contaminated gasworks soil using in situ bioremediation monitored both the removal of PAHs and the accumulation of oxy-PAHs, such as PAH-ketones, quinines and coumarins as dead end products (Lundstedt et al., 2003). As oxy-PAHs are more toxic than the parent PAHs, this study highlights the significance of monitoring the metabolites during bioremediation, specifically for toxic dead-end products, and determining the toxicity of the metabolites both before and after treatment (Lundstedt et al., 2003).

Since PAHs in the environment are present as mixtures, the effect of substrate interaction during biodegradation is crucial in determining the fate of PAHs in nature. It has been reported that, when present as a mixture, the high molecular weight PAHs are degraded followed by the degradation of low molecular weight PAHs, by the bacterial community (Mueller et al., 1989). It has been also observed that a high concentration of naphthalene have inhibitory effect on the degradation of other PAHs, by a defined bacterial coculture (Bouchez et al., 1995). Similarly, a competitive inhibition of phenanthrene degradation by naphthalene, methylnaphthalene and fluorene in binary mixtures using two pure cultures was reported (Stringfellow and Aitken, 1995). Thus during in situ bioremediation of PAHs, concentration of PAHs in the contaminated sites, and a chance for the formation of the toxic dead end product(s) need to be considered for effective process development.

There are some reports on the treatment of PAH contaminated soils and sediments by in situ or ex situ bioremediation methods. Straube et al. (2003) reported that a pilot-scale landfarming treatment of PAH-contaminated soil from a wood treatment facility was achieved by biostimulation of the soil with water, ground rice hulls as a bulking agent, and palletized dried blood as a nitrogen source and bioaugmentation of the microbial community with an inoculum of Pseudomonas aeruginosa strain 64 (Straube et al., 2003). It has been seen that within 1 year of treatment ~86% of total PAHs were removed from the soil including a moderate reduction in HMW PAHs such as pyrene, BaP etc. In a separate study on the treatment of an aged gasworks soil contaminated with aromatic compounds including PAHs, it has been observed that LMW PAHs and heterocyclic compounds were degraded more quickly than the HMW counterparts (Lundstedt et al., 2003). In addition, the unsubstituted PAHs were degraded faster than the related alkyl-PAHs as well as nitrogen-containing heterocyclics. Treatment of soil at a tar contaminated site via composting along with conventional land treatment process revealed that composting led to more extensive PAH removal than did by two different land treatment processes (Guerin, 2000). Sasek et al. (2003) reported the remediation of a manufactured-gas plant soil contaminated with PAHs via composting. In this study, treatment of soil was performed in a thermally insulated chamber using mushroom compost containing wheat straw, chicken manure and gypsum, where at the end of 54 days, removal of ~20–60% of individual PAHs was reported, while ~37–80% of individual PAHs degradation was observed after another 100 days of composting (Sasek et al., 2003). Moreover, studies on the removal of PAHs present in contaminated soil using the associated bacterial communities in two aerobic, lab-scale, slurry-phase bioreactors, which were run semi-continuously and fed either on weekly or monthly basis, showed that most of the PAHs, including HMW PAHs were biodegraded to a greater extent in the weekly fed bioreactor for up to 76% (Singleton et al., 2011).

Polycyclic aromatic hydrocarbons-polluted groundwater can also be bioremediated through both in situ and ex situ bioremediation procedures. However, due to the high cost associated with the extraction and shipment, ex situ treatment of contaminated groundwater is not generally performed. Few reports are documented on the treatment of PAH contaminated water. Bewley and Webb (2001) reported the in situ bioremediation of an aquifer which was contaminated with diverse pollutants, including phenols, BTEX compounds and PAHs. The site was bioremediated using a procedure involving a combination of bioaugmentation and biostimulation. Nutrients (a commercial mixture of urea and diammonium phosphate), a commercially available phenol-degrading mixed bacterial inoculum (PHENOBAC, Microbac Ltd, Durham), and sodium nitrate (an oxygen source) were circulated through the aquifer via a series of injection and abstraction wells. After treatment for over two and half years, the mean concentration of PAHs was reduced to 0.9 μgL^-1 from 11 μgL^-1 along with reduction of other contaminant (Bewley and Webb, 2001). Apart from that, groundwater remediation technology of petroleum-derived compounds (PDCs) including PAHs based on enhanced solubility of PDCs in humic acid was also reported (Van Stempvoort et al., 2002). Zein et al. (2006) reported the treatment of groundwater contaminated with PAHs, gasoline hydrocarbons, and methyl tert-butyl ether using an ex situ aerobic biotreatment system and it was observed that after 10 months of treatment, the concentration of PAHs was reduced to >99% along with other pollutants (Zein et al., 2006).

The bioremediation of PAHs contaminated site is generally very slow because there are a number of biotic and abiotic factors responsible for successful bioremediation. In addition, the incomplete success rates might be due to the fact that some places are heavily contaminated, and hence, the microorganisms are incapable to grow and degrade the contaminant at the same rate at which they are introduced into the environment. Although there is very few information on the use of GEMs in bioremediation, they can be a promising candidate for such processes (Ang et al., 2005; Singh et al., 2008; Megharaj et al., 2011). Using genetic engineering it is possible to enhance the activity or broad substrate specificity of certain enzymes associated with PAH-degrading pathways, which in turn will improve the mineralization of those pollutants in the environment (Timmis et al., 1994; Timmis and Pieper, 1999; Ang et al., 2005; Singh et al., 2008). In recent times, various molecular biology tools such as gene conversion, gene duplication, and transposon or plasmids mediated gene delivery are available, which might play vital roles to boost up the biodegrading potential of naturally occurring microorganisms (Timmis et al., 1994; Timmis and Pieper, 1999; Ang et al., 2005; Singh et al., 2008; Megharaj et al., 2011). However, it is crucial to ensure the stability of GEMs prior to their field application since the catabolic activity of released GEM is associated largely with the stability of the recombinant plasmid introduced into the organism (Brunel et al., 1988; Shaw et al., 1992; Samanta et al., 2002). As GEMs that are released into a contaminated site can spread to other places and multiply under favorable conditions, so before releasing into the environment there should be a clear understanding of the possible side effects associated with GEMs, which would possibly guide the restriction of their application in pollutant abatement (Samanta et al., 2002; Ang et al., 2005; Singh et al., 2008; Megharaj et al., 2011).