Aerobic Degradation of Petroleum Components by Microbial Consortia

This article is a state–of-the-art review on the aerobic degradation of petroleum components that are commonly found in the environment. Numerous microorganisms have been isolated and their phylogeny and metabolic capacity to degrade a variety of aliphatic and aromatic hydrocarbons have been demonstrated. This review focuses on recent progress on how microbes degrade hydrocarbons and heteroaromatic components of petroleum contaminants directed towards better understanding of the aerobic degradation processes and their exploitation for bioremediation. The phylogenetic diversity of the oil-degrading microbes was also discussed. Aerobic Degradation of Petroleum Components by Microbial Consortia


Introduction
The biodegradation of petroleum and other hydrocarbons in the environment is a complex process, whose quantitative and qualitative transformations depend on the nature and amount of the oil or hydrocarbons present, the ambient and seasonal environmental conditions, such as free or dissolved oxygen, optimum temperature for oil degradation (20-35°C), physical and/or chemical dispersion of oil, turbulent conditions as opposed to quiescent conditions, and the composition of the autochthonous microbial community [1][2][3][4][5][6][7]. Microbial degradation of oil has been shown to occur by attack on aliphatic or light aromatic fractions of the oil, with high-molecularweight aromatics, resins, and asphaltenes considered to be recalcitrant or exhibiting only very low rates of biodegradation, although some studies have reported their removal at high rates under optimal conditions [8,9].
Biodegradation rates generally increase with increasing temperature such that ecosystems exposed to extremely low temperatures degrade hydrocarbons very slowly. The microbial degradation of petroleum in aquatic environments is limited primarily by nutrients such as nitrogen and phosphorus; salinity and pressure may be important in estuarine and deep-sea regions, respectively. Oxygen, nutrient concentrations, moisture, and pH are predominant factors in determining biodegradation rates in soil.
Petroleum is a complex mixture of different hydrocarbons including aliphatic (linear or branched), cycloalkanes, mono-and polyaromatics, asphaltenes and resins and majority of these compounds are stable, toxic, and carcinogenic [10,11]. Petroleum compounds such as alkanes, benzene, toluene, ethyl benzene, and xylenes (BTEX) and some polycyclic aromatic hydrocarbons (PAHs) are biodegradable under the proper environmental conditions [3,12] and low salinity marine habitats [13][14][15][16]. However, higher molecular PAHs, polycyclic aromatic sulphur heterocyclics (PASHs), methyl tertiary butyl ether (MTBE), gasoline additive and other components of petroleum products may be recalcitrant to biodegradation. The non-biodegradable components can still pose a high risk in the immediate vicinity of the area in which they remain. Petroleum hydrocarbons are therefore major contaminants in the environment and they cause damages to the surrounding ecosystems. Oil-contaminated soil, groundwater, and/ or wastewater may contain a mixture of contaminant types including salts, organics, alcohols, phenols, acid, radionuclides, PAHs, and trace elements like zinc, cadmium, mercury, copper, chromium, lead etc. at widely varying concentrations [17][18][19]. The present review was focus on aerobic degradation process of various components of petroleum by microbial consortia. The main petroleum components discussed include aliphatics, alicyclic, and aromatics hydrocarbons as well as the N-, S-and O-heterocyclic aromatic hydrocarbons. Also addressed is the phylogenetic diversity of the oil-degrading microbes.

Biodegradation of Petroleum Compounds
Crude oil is a mixture of hydrocarbons composed of mainly heteroatomic and non-heteroatomic hydrocarbons [11]. To date many studies have reported the ability of microorganisms to utilize crude oil components as the growth substrates (Table 1). Zvyagintseva et al. [20] have reported degradation of isoprenoid and n-alkane fractions of crude oils to a significant extent by an enrichment developed from the brines of the Kalamkass oil fields in Kazakhstan. Diaz et al. [21] have enriched microbial consortia, MPD-7 and MPD-M from Cormorant oil fields in North Sea and sediments associated with mangrove roots, respectively. These cultures degraded aliphatic and aromatic hydrocarbons in crude oil. Total oil degradation by MPD-7 ranged from 20 to 38%, while MPD-M degraded much higher amount of crude oil ranging between 45 and 48%. In a subsequent study, Diaz et al. [22] have immobilized the MPD-M culture on polypropylene fibers and the culture was reported to degrade crude oil. Riis et al. [23] showed the degradation of diesel fuel by microbial communities extracted from Argentinean saline soils. In addition, these authors isolated several halotolerant bacteria of the genera Cellulomonas, Bacillus, Dietzia, and Halomonas with the ability to degrade crude oil as the carbon source. Obuekwe et al. [24] reported the isolation of Fusarium lateritium, Drechslera sp, and Papulaspora sp. from a salt marsh in the Kuwaiti desert that are capable of degrading crude oil as the sole carbon source. Similarly, several crude oil or petroleum usually requires the cooperation of more than one single species. Individual microorganisms can metabolize only a limited range of hydrocarbon substrates, so a consortium composed of many different bacterial species with overall broad enzymatic capacities is required to increase the rate of petroleum biodegradation. Hydrocarbon degradation by microbial communities depends on the composition of the community and its adaptive response to the presence of hydrocarbons. Bacteria and fungi are the key agents of degradation, with bacteria assuming the dominant role in marine ecosystems and fungi becoming more important in freshwater and terrestrial environments. Adapted communities, i.e., those which have been previously exposed to hydrocarbons, exhibit higher biodegradation rates than communities with no history of hydrocarbon contamination. The mechanisms of adaptation which include both selective enrichment and genetic changes result in a net increase in the number of hydrocarbon utilizing organisms and in the pool of hydrocarboncatabolizing genes within the community. Petroleum compounds differ in their susceptibility to microbial attack and generally degrade in the following order of decreasing susceptibility [37]: n-alkanes > branched alkanes > low molecular weight aromatics >cyclic alkanes, > polycyclic aromatic hydrocarbons > polar compounds. Although many of these compounds can be relatively easily degraded under soil and fresh water environments [3,12] and low salinity marine habitats [13][14][15][16]. Biodegradation rates have been shown to be highest for the saturates, followed by the light aromatics, with high-molecular-weight PAHs and polar compounds (resins and asphaltenes) exhibiting extremely low rates of degradation or may not be degraded at all [38][39][40].

Biodegradation of petroleum compounds in different ecosystems
In many ecosystems there is already an adequate indigenous microbial community capable of extensive oil biodegradation, provided that environmental conditions are favorable for oil-degrading metabolic activity. There are several advantages of depending on indigenous microorganisms rather than adding microorganisms to degrade hydrocarbons. First, natural populations must have developed through many years. These microorganisms are adapted for survival and proliferation in that environment. Secondly, the ability to utilize reported the isolation of several strains of hydrocarbon-oxidizing bacteria representing the genera Rhodococcus, Gordonia, Dietzia, and Pseudomonas from oil and stratal waters of Tatarstan, western Siberia, and Vietnam oil fields. All these strains oxidized n-alkane fraction of crude oil. A Bacillus sp. strain DHT, isolated from oil contaminated soil, grew and produced biosurfactant when cultured in the presence of variety of hydrocarbons including crude oil, diesel oil, hexadecane, naphthalene, pyrene, dibenzothiophene, salicylate, catechol, and phenanthrene as the sole carbon sources at 30-45°C. However, no growth occurred on toluene, phenol, 2-hydroxyquinoline and carbazole [31]. Similarly, Mnif et al. [32] have reported the isolation of several strains of thermophilic and mesophilic hydrocarbon degrading as well as biosurfactant producing organisms from Tunisian oil fields. Among these, Pseudomonas sp. strain C450R and Halomonas sp. strain C2SS100 were reported to degrade 93-96% of the aliphatic fraction of crude oil (C 13 -C 29 ), while producing biosurfactants. Chamkha et al. [33] have isolated a strain C5 closely related to Geobacillus pallidus from a tyrosol degrading enrichment developed from production water from a hightemperature oil field in Tunisia. The organism degraded crude oil and diesel as the source of carbon. Wang et al. [34] have isolated a moderate halophilic actinomycete, Amycolicicoccus subflavus DQS3-9A1 T from oily sludge at Daqing Oil field, China with the ability to degrade crude oil. Later, Nie et al. [35] studied the genetic capability of the DQS3-9A1 T to metabolize a range of short-chain and long-chain n-alkanes such as propane and C 10 -C 36 alkanes in crude oil, respectively, as the sole carbon sources. Recently, Al-Mailem et al. [36] have isolated Marinobacter sedimentalis and Marinobacter falvimaris from soil and pond water collected from hypersaline Sabkhas in Kuwait. Isolation of these organisms was accomplished using agar plates provided with crude oil vapor as the sole source of carbon. These studies also showed that both organisms were capable of fixing atmospheric nitrogen and such potential is beneficial for effective bioremediation of petroleum compounds without the need of providing fertilizer. Biodegradations of hydrocarbons is distributed among a diverse microbial population. This population occurs in natural ecosystems and either independently or synergistically metabolizes various hydrocarbons. Many times, when the amount of microorganisms is sufficient in the contaminated environment, microbial seeding may not be required. Microorganisms (bacteria and fungi) have different rates at which they utilize and degrade hydrocarbons in the soil or water. This rate is reflected in the multiplication and colony forming units (cfu) for the isolated organisms. The use of microorganisms to degrade petroleum hydrocarbon resulting from oil spillage has been a subject of extensive research since the first publication of bacterial growth on petroleum hydrocarbons [41,42]. Several petroleum hydrocarbon degrading microorganisms have been isolated from both soil and marine sources, which are the two major environments affected by petroleum hydrocarbon pollution [43,44]. Microorganisms are equipped with metabolic machinery to use petroleum products as a carbon and energy source. The metabolic pathways that hydrocarbon-degrading heterotrophs use can be either aerobic (i.e. they utilize oxygen as the primary electron acceptor) or anaerobic (i.e. they utilize an alternative electron acceptor such as nitrate or sulfate). Aerobic degradation usually proceeds more rapidly and is considered to be more effective than anaerobic degradation because of the less free energy required for initiation and energy yield per reaction. This review mainly focuses on aerobic degradation of petroleum components in different ecosystems.
Soil ecosystem: Petroleum compounds bind to soil components, and they are difficult to be removed or degraded [45]. Petroleum contamination in soil results in an imbalance in the carbon-nitrogen ratio at the spill site, because crude oil is essentially a mixture of carbon and hydrogen. This causes a nitrogen deficiency in oil soaked soil, thus retarding the growth of bacteria and the utilization of carbon sources. Furthermore, large concentrations of biodegradable organics in the top layer soil deplete oxygen reserves in the soil and slow down the rates of oxygen diffusion into deeper layers. Many indigenous microorganisms in water and soil are capable of degrading petroleum contaminants [46,47]. Petroleum hydrocarbons in nature are degraded by diverse groups of microorganisms, which are capable of utilizing hydrocarbons as food [48]. The degradation of complex mixtures of hydrocarbons such as crude oil and metals in soil requires mixedpopulations with overall broad enzymic capacities [43,49]. Bacteria are the most active agents in petroleum degradation and they work as primary degraders of spilled oil in environment [50,51]. Several bacteria are known to feed exclusively on hydrocarbons [52].
Biodegradation of crude petroleum oil from petroleum contaminated soil from North East India was reported by Das and Mukherjee [53]. Acinetobacter sp. was reported to be capable of utilizing n-alkanes of chain length C 10 -C 40 as a sole source of carbon [54]. Bacterial genera, namely, Gordonia, Brevibacterium, Aeromicrobium, Dietzia, 3Burkholderia, and Mycobacterium isolated from petroleum contaminated soil proved to be the potential organisms for hydrocarbon degradation [47,[55][56][57][58][59]. The degradation of PAHs by Sphingomonas was reported by Daugulis and McCracken [56]. Tang et al [8] reported the degradation process of crude oil by one artificial microalgal-bacteria consortium. The consortium which was constructed by one axenic Scenedesmus obliquus named GH2 and four oil component-degrading bacteria with known complementary degradative capabilities, including Sphingomonas GY2B, Burkholderia cepacia GS3C, Pseudomonas GP3A and Pandoraea pnomenusa GP3B were reported to completely eliminate alkanes, alkylcycloalkanes and alkylbenzens in 10 and 7 days, respectively. The consortium also preferentially attacked high molecular weight PAHs such as phenanthrene and methylphenanthrenes, a lot of C 2 , C 3 naphthalene isomers and some extra lower molecular substances were produced during the PAHs degradation. Chang et al. [9] investigated the extent of biodegradation of non-volatile petroleum hydrocarbons (C 16 -C 34 ) and the associated microbial activity in predominant aggregate sizes during a pilot-scale biopile experiment conducted at 15°C, with a clayey soil, from a crude oil-impacted site in northern Canada. At the end of 65-d biopile experiment, 42% of the C 16 -C 34 hydrocarbons were reported to be degraded in the nutrientamended macroaggregates, compared to 13% in the mesoaggregates. Higher microbial activity in the macroaggregates of the nutrientamended biopile was inferred from a larger increase in extractable protein concentrations, compared to the other aggregates. Fungal genera, namely, Amorphoteca, Neosartorya, Talaromyces, and Graphium and yeast genera, namely, Candida, Yarrowia, and Pichia were isolated from petroleum contaminated soil and proved to be the potential organisms for hydrocarbon degradation [55]. Singh [60] also reported a group of terrestrial fungi, namely, Aspergillus, Cephalosporium, and Pencillium which were also found to be the potential degrader of crude oil hydrocarbons. Adenipekun [61] reported that Pleurotus tuberregium have the ability to increase nutrient contents in soils polluted with 1 -40% engine-oil concentration after six months of incubation and reduction in heavy metals after six months of incubation. Hence, the fungus can be employed in decontaminating environment polluted with engine oil. In a similar study, Adenipekun and Isikhuemhen [62] revealed the ability of white rot fungus, L. squarrosulus to improve the nutrient contents of the engine oil contaminated soil and an accumulation of Fe, Zn and Ni to an appreciable extent.
Aquatic ecosystem: In aquatic ecosystems, dispersion and emulsification of oil in slicks appear to be prerequisites for rapid biodegradation. Large masses of mousse, tar balls or high concentrations of oil in quiescent environments tend to persist because of the limited surface areas available for microbial activity. Petroleum fractions containing asphalt components are not degraded quantitatively [63]. Therefore, floating tar globules are encountred in the marine environment in increasing quantities because they are practically oxygenated high molecular weight materials that resist furher microbial degradation. When oil spill occurs, a combination of recovery, disposal and the containment of oil is performed thereafter. The conventional methods to remove oil from aquatic ecosystems include; mechanical clean up, chemical clean up and microbial degradation. Mechanical cleaning of spilled oil and dispersant is nearly impossible in "protected" ecosystems. Microbial degradation is the major mechanism for the elimination of spilled oil and dispersants from aquatic environment [64].
The yeast species, namely, Candida lipolytica, Rhodotorula mucilaginosa, Geotrichum sp, and Trichosporon mucoides isolated from contaminated water were noted to degrade petroleum compounds [67,68]. Hidayat and Tachibana [69] demonstrated that Fusarium sp. F092 have the ability to degrade chrysene and the aliphatic fraction of crude oil contaminating liquid culture with artificial sea water (35%).
Though algae and protozoa are the important members of the microbial community in both aquatic and terrestrial ecosystems, reports on their involvement in hydrocarbon biodegradation are scanty. Walker et al. [70] isolated an alga, Prototheca zopfi which was capable of utilizing crude oil and a mixed hydrocarbon substrate and exhibited extensive degradation of n-alkanes and isoalkanes as well as aromatic hydrocarbons. Cerniglia et al. [71] also observed nine cyanobacteria, five green algae, one red alga, one brown alga, and two diatoms that could oxidize naphthalene. None of the Protozoa had been reported to utilize hydrocarbons.
In the aquatic ecosystems, fungi plays an important role during their ability in removing hazardous compounds from the water, whereas sediment particles contaminated with crude oil from oil spills is one of the desired ecological niche to fungi which inhabits such substrate and use carbon source from hydrocarbons in polluted sediment particles to biodegrade crude oil from the sediments in the beaches. Fungi have been found to be better degraders of petroleum than traditional bioremediation techniques including bacteria, and although hydrocarbon degraders may be expected to be readily isolated from a petroleum oil-associated environment [80,81]. The ability of most fungi to produce extracellular enzymes for the assimilation of complex carbohydrates makes possible the degradation of a wide range of pollutants. They also have dvantage of being relatively easy to grow in fermenters, thus being suited for large scale production. Another advantage is the ease of separation of fungal biomass by filtration due to its filamentous structure. They are also less sensitive to variations in nutrients, aeration, pH, temperature and have a lower nucleic content in the biomass as compared to yeasts.

Marine ecosystem:
The microbial response to an oil spill at sea is dependent on numerous factors, including the oil composition and degree of weathering, as well as environmental conditions, particularly temperature and nutrient concentrations [16]. Nevertheless, there are some typical patterns; most notable is the large increase in abundance of Alcanivorax spp., which degrade straight-chain and branched alkanes [84][85][86][87], followed by Cycloclasticus spp., which degrade PAHs [15,52,86,87].
Since the cultivation of Alcanivorax borkumensis [88], functional genomic, biochemical and physiological analyses have revealed the underlying basis of its success [52,[89][90][91]. While it lacks catabolic versatility, utilising alkanes almost exclusively as carbon and energy sources, it has multiple alkane-catabolism pathways, with key enzymes including alkane hydoxylases (a non-haem diiron monooxygenase; AlkB1 and AlkB2) and three cytochrome P450-dependent alkane monooxygenases [89]. Their relative expression is influenced by the type of alkane supplied as carbon and energy source and phase of growth [89]. Alcanivorax borkumensis also possesses a multitude of other adaptations to access oil (e.g. synthesis of emulsifiers and biofilm formation [89] and to survive in open marine environments [87,89]. Acinetobacter spp., which are commonly isolated from oilcontaminated marine environments [92], also have a diverse array of alkane hydroxylase systems enabling them to metabolize both shortand long-chain alkanes [92,93].
Microbial degradation of petroleum hydrocarbons in a polluted tropical stream in Lagos, Nigeria was reported by [94]. Nine bacterial strains, namely, Pseudomonas fluorescens, P. aeruginosa, Bacillus subtilis, Bacillus sp., Alcaligenes sp., Acinetobacter lwoffi, Flavobacterium sp., Micrococcus roseus, and Corynebacterium sp. were isolated from the polluted stream which could be responsible for the degradation of the crude oil.

Aerobic degradation of petroleum compounds
The most rapid and complete degradation of the majority of organic pollutants is brought about under aerobic conditions. Figure  1 shows the main principle of aerobic degradation of hydrocarbons [95]. The initial intracellular attack of organic pollutants is an oxidative process and the activation as well as incorporation of oxygen which is the enzymatic key reaction catalyzed by oxygenases and peroxidases. Peripheral degradation pathways convert organic pollutants step by step into intermediates of the central intermediary metabolism, for example,  the tricarboxylic acid cycle. Biosynthesis of cell biomass occurs from the central precursor metabolites, for example, acetyl-CoA, succinate and pyruvate. Sugars required for various biosyntheses and growth are synthesized by gluconeogenesis.
Aliphatic hydrocarbons: Alkanes, a major group in crude oil, are readily biodegraded in the marine and non-marine environments. Oxidation of alkanes is classified as being terminal or sub-terminal. The degradation of alkanes of medium chain length by Pseudomonas putida containing the OCT plasmid is initiated by alkane hydroxylase [85,[96][97][98]. This enzyme consists of three components: the membranebound oxygenase component and two soluble components called rubredoxin and rubredoxin reductase. The catalytic centre of the oxygenase component contains a dinuclear iron cluster which is also found in other enzymes such as methane monooxygenase and ribonucleotide reductase [96,99]. In P. putida (OCT), oxidation of the methyl group of alkanes by alkane hydroxylase yields alkanols that are further oxidized by a membrane-bound alcohol dehydrogenase to alkanals. The alkanals are subsequently transformed to fatty acids and then to acyl CoA by aldehyde dehydrogenase and acyl-CoA synthetase, respectively (Scheme 1) [99]. An alkane degrading pathway yielding secondary alcohols has also been reported. In this pathway, alkanes are oxidized by monooxygenase to secondary alcohols, then to ketones, and finally to fatty acids (Scheme 1) [100,101]. In Acinetobacter strain M-1, alkanes are transformed to alkyl peroxides, and these molecules would be further metabolized to the corresponding aldehyde. The first enzyme involved in this pathway contains FAD + and Cu 2+ as prosthetic groups (Scheme 1) [54,102].
Many yeast species, e.g. Candida maltosa, Candida tropicalis and Candida apicola, were investigated for use with alkanes [103,104]. The first step of alkane degradation (terminal hydroxylation) and of ω-hydroxylation is catalyzed by P450 monooxygenase. The alcohols thus formed are processed by fatty alcohol oxidase and fatty aldehyde dehydrogenase. The P450 enzyme from some yeast strains can catalyze not only the terminal hydroxylation of long-chain alkanes and the ω-hydroxylation of fatty acids, but also the subsequent two steps to yield fatty acids and α, ω-dioic acids (Scheme 1) [104,105]. Catabolic pathways for the degradation of branched alkanes have been elucidated for a few bacteria; for example, Rhodococcus strain BPM 1613 degraded phytane (2,6,10,14-tetramethylhexadecane), norpristane (2,6,10-trimethylpentadecane) and farnesane

B-oxidation
Acetyl-CoA Fatty acid n-Alkane degradation via alkyl hydroperoxides (2,6,10-trimethyldodecane) via ß-oxidation [96,106]. β-Oxidation of the fatty acids results in the formation of acetyl-CoA. Alkanes with an uneven number of carbon atoms are degraded to propionyl-CoA, which is in turn carboxylated to methylmalonyl-CoA and further converted to succinyl-CoA. The sub-terminal oxidation occurs with lower (C 3 -C 6 ) and longer alkanes with the formation of a secondary alcohol and subsequent ketone (Scheme 1) [95]. Branching, in general, reduces the rate of biodegradation. Methyl side groups do not drastically decrease the biodegradability, whereas complex branching chains, e.g., the tertiary butyl group, hinder the action of the degradative enzymes [95]. Al-Mueini et al. [107] have reported the isolation of an extremely halophilic actino-mycete, Actinopolysporasp. DPD1 from an oil production site in the Sultanate of Oman and was shown to degrade n-alkanes (pentadecane, eicosane, pentacoase) and fluorine. The organism efficiently degraded pentadecane (100% in 4 days) and eicosane (80% in 10 days  Haloferax. In addition, strain MSNC14 also degraded phenanthrene. Three extremely halophilic archaeal strains, Haloferax, Halobacterium and Halococcus isolated on the basis of crude oil utilization also degraded n-alkanes and mono and polyaromatic compounds as the sole sources of carbon and energy [111]. Overall, studies reveal that both bacteria and archaea have the capacity to metabolize n-alkanes with varying chain lengths ( Table 2).
Alicyclic hydrocarbons: Cycloalkanes representing minor components of mineral oil are relatively recalcitrant to microbial attack [95]. The absence of an exposed terminal methyl group as in n-alkanes complicates the primary attack. A few species are able to use cyclohexane as sole carbon source. Cycloalkanes, including condensed cycloalkanes are degraded by a co-oxidation mechanism [96,101,112]. The mechanism of cyclohexane degradation is shown in Scheme 2. The formation of a cyclic alcohol and a ketone has been observed [101]. A monooxygenase introduces an oxygen into the cyclic ketone, and the cyclic ring is cleaved (Scheme 2).
Aromatic compounds: A multitude of catabolic pathways for the degradation of aromatic compounds have been elucidated; for example, toluene is degraded by bacteria along five different pathways. On the pathway encoded by the TOL plasmid, toluene is successively degraded to benzyl alcohol, benzaldehyde and benzoate, which is further transformed to the TCA cycle intermediates [13]. The first step of toluene degradation with P. putida F1 is the introduction of two hydroxyl groups to toluene, forming cis-toluene dihydrodiol. This intermediate is then converted to 3-methylcatechol [113]. With Pseudomonas mendocina KR1, toluene is converted by toluene 4-monooxygenase to p-cresol, this being followed by p-hydroxybenzoate formation through oxidation of the methyl side chain [13,114]. With Pseudomonas pickettii PKO1, toluene is oxidized by toluene 3-monooxygenase to m-cresol, which is further oxidized to 3-methylcatechol by another monooxygenase [13,114]. With Bukholderia cepacia G4, toluene is metabolized to o-cresol by toluene 2-monooxygenase, this intermediate being transformed by another monooxygenase to 3-methylcatechol [13].
Burkholderia sp. strain JS150 is unique in using multiple pathways for the metabolism of toluene (Scheme 3) [115].
The oxygenolytic cleavage of the aromatic ring occurs via o-or m-cleavage [116]. The metabolism of a wide spectrum of aromatic compounds by one species requires the metabolic isolation of intermediates into distinct pathways. The key enzymes of the degradation of aromatic substrates are induced and synthesized in appreciable amounts only when the substrate or structurally related compounds are present. Scheme 4 shows the pathways of the oxygenolytic ring cleavage of phenol to intermediates of the central metabolism. At the branch-point, catechol is either oxidized by the intradiol o-cleavage, or the extradiol m-cleavage. Both ring cleavage reactions are catalyzed by specific dioxygenases. The product of the o-cleavage -cis, cismuconate -is transferred to the unstable enollactone, which is in turn hydrolyzed to oxoadipate. This dicarboxylic acid is activated by transfer to CoA, followed by the thiolytic cleavage to acetyl-CoA and succinate. Protocatechuate is metabolized by a homologous set of enzymes. The additional carboxylic group is decarboxylated and, simultaneously, the double bond is shifted to form oxoadipate enollactone [95].
The oxygenolytic m-cleavage yields 2-hydroxymuconic semialdehyde, which is metabolized by the hydrolytic enzymes to formate, acetaldehyde, and pyruvate (Scheme 4). These are then utilized in the central metabolism. In general, a wealth of aromatic substrates is degraded by a limited number of reactions, which include hydroxylation, oxygenolytic ring cleavage, isomerization, and hydrolysis [117][118][119][120]. The inducible nature of the enzymes and their substrate specificity enable bacteria with a high degradation potential, e.g., Pseudomonads and Rhodococci, to adapt their metabolism to the effective utilization of substrate mixtures in polluted soils and to grow at a high rate.
Several studies have successfully isolated bacteria and archaea that degrade oxygenated aromatics. Table 3 lists organisms that degrade oxygenated hydrocarbons. Woolard and Irvine [121] showed that a halophile isolated from a mixed culture obtained from a saltern at GSL Utah readily degraded phenol. Alva and Peyton [122] isolated a haloalkaliphile, Halomonas campisalis near Soap Lake in central Washington and showed that this organism degraded phenol and catechol as the sole sources of carbon at pH 8-11. A Gram-positive halophilic bacterium, Thalassobacillus devorans isolated from an enrichment culture developed from saline habitats in southern Spain was shown to degrade phenol [123]. The strain C5, closely related to Geobacillus pallidus isolated from a tyrosol-utilizing enrichment also degrades a variety of other oxygenated aromatic compounds including benzoic, p-hydroxybenzoic, protocatechuic, vanillic, p-hydroxyphenylacetic, 3,4-dihydroxyphenylacetic, cinnamic, ferulic, phenol, and m-cresol. However, no degradation of non-oxygenated hydrocarbons such as toluene, naphthalene, and phenanthrene was observed [33]. Recently, Bonfá et al. [124]  halodurans (re-classified as Halomonas halodurans) degrades benzoic acid [125]. Garcia et al. [126,127] have isolated several strains of Halomonas spp. including the Halomonas organivorans from water and sediment of salterns and hypersalines oils collected in different part of the Southern Spain. These isolates degraded a wide range of aromatic compounds including benzoic acid, p-hydroxy benzoic acid, phenol, salicylic acid, p-aminosalicylic acid, phenylacetic acid, phenylpropionic acid, cinnamic acid, ferulic acid, and p-coumaric acid as the sole sources of carbon. Abdelkafi et al. [128] have reported the isolation of a p-coumaric acid degrading Halomonas strain IMPC from a p-coumaric acid degrading enrichment culture obtained from a Table-olive fermentation rich in aromatic compounds. This strain converted p-coumaric acid to p-hydroxybenzaldehyde, p-hydroxybenzoic acid, and then to protocatechuic acid prior to ring cleavage. In addition, the strain also degraded other lignin-related compounds such as cinnamic acid, m-coumaric acid, m-and p-methoxy cinnamic acid, mand p-methylcinnamic acid, and ferulic acid to their corresponding benzoic acid derivatives. Oie et al. [129] have studied the degradation of benzoate and salicylate by Halomonas campisalis isolated from an alkaline Soap Lake. This study showed that the organism degraded benzoate and salicylate to catechol and then to cis, cis-muconate thus indicating degradation via the ortho-cleavage pathway. Kim et al. [130] have isolated a Chromohalobacter sp. strain HS-2 from salted fermented clams that degrades benzoate and p-hydroxybenzoate as the sole carbon and energy sources.
Cuadros-Orellana et al. [131] have reported the isolation of 10 halophilic archaea from Dead Sea that degrades p-hydroxybenzoic acid as the sole carbon and energy source. In addition, strain L1, a member of the unclassified Halobacteriaceae family of the phylum, Euryarchaeota also degrades benzoic acid to gentisate. Erdogmus et al. [132] reported the ability of many archaeal strains belonging to Halobacterium, Haloferax, Halorubrum, and Haloarcula group to degrade p-hydroxybenzoic acid. These studies clearly demonstrate that archaea that metabolize p-hydroxybenzoic are wide spread in the environment. Among bacteria, Halomonas spp. has been frequently reported for their ability to degrade phenolics and benzoates and only few reports exist on their potential to degrade non-oxygenated hydrocarbons.
The most abundant hydrocarbons in produced water are the onering aromatic hydrocarbons, benzene, toluene, ethyl benzene, and xylenes (BTEX) and low molecular weight saturated hydrocarbons [133]. Benzene is a category A carcinogen. Leakage from produced water storage tanks, pipelines, spills, and seepage from surface contaminated sites can cause major BTEX contamination [10]. BTEX are relatively highly soluble in water and hence can contaminate large volumes of ground water. Although there have been many recent reports on the biodegradation of non-oxygenated hydrocarbons, only few reports exist on the biodegradation of BTEX compounds (Table  4). Nicholson and Fathepure [134,135] have reported the degradation of BTEX in microcosms established with soil samples from an oilfield and from an uncontaminated salt flat in Oklahoma. Hassan et al. [136] have reported the isolation of Alcanivorax sp. HA03 from soda lakes in Wadi E1Natrun capable of degrading benzene, toluene, and chlorobenzene as the sole sources of carbon. This observation that Alcanivorax can also degrade aromatic compounds expands the metabolic capability of this group of organisms because Alcanivorax   are primarily known for their ability to degrade aliphatic hydrocarbons. Degradation of benzene was also reported in archaea. For example, the crude oil degrading Haloferax, Halobacterium, and Halococcus isolated from a hypersaline Arabian Gulf coast degraded benzene as the sole source of carbon [111].The complete degradation of PAHs requires a community of microorganisms. PAHs are taken up by microorganisms and are activated in aerobic metabolism by insertion of two oxygen atoms by bacteria and green algae to produce either cis-dihydrodiols or phenols [137]. Simple PAHs such as naphthalene, biphenyl and phenanthrene are readily degraded aerobically. The degradation of these compounds is generally initiated by dihydroxylation of one of the PAH rings, this being followed by cleavage of the dihydroxylated ring. Ring hydroxylation is catalyzed by a multi-component dioxygenase which consists of a reductase, a ferredoxin, and an iron sulphur protein, while ring cleavage is generally catalyzed by an iron-containing metacleavage enzyme. The carbon skeleton produced by the ring-cleavage reaction is then dismantled, before cleavage of the second aromatic ring (Scheme 5) [138].
Plotnikova et al. [139,140] have isolated Pseudomonas sp., Rhodococcus sp., Arthrobacter sp., and Bacillus sp. from soil and sediment contaminated with waste generated by chemical and saltproducing plants. All these isolates degraded naphthalene and salicylate as the sole carbon sources. In addition, some of these organisms also grew on phenanthrene, biphenyl, o-phthalate, gentisate, octane, and phenol as the sole sources of carbon. Zhao et al. [141]  Garcia et al. [126] Abdelkafi et al. [128] Garcia et al. [123] Chamkha et al. [33]   Marinobacter nanhaiticus strain D15-8W from a phenanthrenedegrading enrichment obtained from sediment from the South China Sea. The strain D15-8W degrades naphthalene, phenanthrene or anthracene as the sole source of carbon. Bonfá et al. [144] have isolated several strains of Haloferax that degrade a mixture of the PAHs including naphthalene, anthracene, phenanthrene, pyrene and benzo[a] anthracene. Extremely halophilic archaeal strains of Haloferax, Halobacterium, and Halococcus isolated from a hypersaline coastal area of the Arabian Gulf not only degraded crude oil and n-octadecane as the carbon sources, but also grew on phenanthrene [111]. Erdogmu¸s et al. [132] showed the degradation of naphthalene, phenanthrene and pyrene as the sole carbon sources by several archaeal strains including Halobacterium piscisalsi, Halorubrum ezzemoulense, Halobacterium salinarium, Haloarcula hispanica, Haloferax sp. Halorubrum sp. and Haloarcula sp. isolated from brine samples of Camalt Saltern in Turkey.
All these studies demonstrate the potential of bacteria and archaea to degrade PAHs (Table 5).
Numerous works have been done on the use of fungi for biodegradation of petroleum hydrocarbons [38,44,78,116,145]. Most filamentous fungi is unable to totally mineralize aromatic hydrocarbons; but only transform them into indirect products of lowered toxicity and increased susceptibility to decomposition with the use of bactetria. Among the filamentous fungi capable of aliphatic hydrocarbon biodegradation include Cladophialophoria and Aspergillus, whereas fungi belonging to Cunninghamella, Penicillimum, Fusarium and Aspergillus are capable of degrading aromatic hydrocarbons. Prenafela-Boldu et al. [2,116] have reported the use of Cladophialophoria sp. fungi in monoaromatic hydrocarbons (BTEX) mineralization. Their results showed that the decomposition is more dynamic for toluene, ethylbenzene and m-xylene than for benzene. Common fungi with ability of biodegradation of aromatic compounds (BTEX, PAH) include Phanerochaete chrysosporum. This fungus produces extracellular enzymes (lignic peroxydase) that participate in decomposition of a lignic cell -wall in plants and in oxygenation of various xenobiotics. The microorganism transforms PAH into chinone derivative and later splits the aromatic ring of chinones with their complete mineralization in consecutive stages. Flayyih and AI-Jawhari [146]   and Peniclllium funiculosum were found to be more predominant in the polluted soil. The highest percentage loss of petroleum hydrocarbon concentration by the mixed cultures of fungi were 90% with A. niger and A. fumigatus, but the lowest loss of petroleum hydrocarbon calculated in mixed four fungal strains (A. niger, A. fumigatus, P. funiculosum and F. solani) to 70%. Vanishree et al. [147] used fungus Penicillium sp. for biodegradation of petrol. The efficiency of the fungal strain on the degradation of different concentrations of petrol was studied. The ability of Penicillium sp. to tolerateoil pollutant and grow on them suggest that it can be employed as bioremediation agent and used for restoration of ecosystem contaminated by oil.
Four fungi strains viz. Aspergillus niger, Aspergillus terreus, Rhizopus sp and Penicillium sp were also isolated from soil and tarball samples collected from mangrove forest of Alibaug and Akshi coastal area, Maharashtra, India [148]. These strains were assessed for their degradation capability of petroleum hydrocarbons measuring growth diameter in Potato Dextrose Agar (PDA) solid media for different concentrations of kerosene. Rhizopus sp showed the highest growth diameter in 5% kerosene and Aspergillus niger showed the highest growth diameter in 20% kerosene while, penicillium sp showed the lowest growth diameter at all the concentrations of kerosene as compared to other three strains. A mixed culture consisting of penicillium sp, Rhizopus sp and Aspergillus terreus was reported to show highest growth diameter.
Benzothiophene (BT) and its derivatives are the major sulfur heteroaromatic compounds that are commonly found in higher molecular weight fractions of petroleum. However, no bacterial strain has been found to grow on benzothiophenes as the sole carbon source, and all reported biotransformations of benzothiophenes are based on cometabolism. Attacks on the thiophene ring of benzothiophenes lead to the formation of sulfoxides and sulfones, or to ring opening and the formation of 2-mercaptomandelaldehyde and 2-mercaptophenylglyoxalate [158][159][160]. Catabolism of dibenzothiophene is catalyzed by distinct enzymes in two pathways (Scheme 7). The catabolic branch of initial sulfur oxidation, also called 4S pathway, through which rapid desulfurization can be obtained. Consecutive desulfurization is achieved by desulfinase. These monooxygenases require FAD as a co-factor, accompanied by a specific flavin reductase [161]. FAD containing monooxygenases are very common in all biota and catalyze various detoxification steps. For example, Sutherland et al. [162] reported a FAD containing monooxygenase of high sequence homology with dibenzothiophene desulfurization enzymes. In general, all bacterial species with highdesulfurization activities have the enzymes and broad substrate range [163,164]. However, some bacterial species are defective in some components, which results in accumulation of intermediates. In comparison with desulfurization pathway, the enzymes for lateral dioxygenation and consecutive reactions are very different (Kodama pathway, Scheme 7). Many bacterial species have been reported to metabolize dibenzothiophene through Kodama pathway [120,150,165]. Although no detailed research has been done, the structural similarities of the metabolites suggest that common PAH dioxygenase and enzymes in successive steps may be involved in Kodama pathway.

Phylogeny of the oil-degrading microbes
Oil-degrading microorganisms are ubiquitous in the environment, particularly in the oil-polluted sites. Both fungi and bacteria have been found to be useful in biodegradation process, even though many researches have been on bacteria in the recent times. Although a wide phylogenetic diversity of microorganisms is capable of aerobic degradation of contaminants, Pseudomonas species and closely related organisms have been the most extensively studied owing to their ability to degrade many different contaminants [167]. The oil-degrading populations are widely distributed in the lands and water bodies.

Microbial biodiversity in terrestrial ecosystem
A conceptualization of the functioning of the ancient terrestrial biosphere necessarily requires a general understanding of modern, analog microbial communities to evaluate their living requirements, diversity, physiology, and environmental impact, and to characterize any potential biosignature that could be used to recognize them in the rocks [168,169]. Modern terrestrial microbial communities are found worldwide and in a great variety of local conditions, in surface (solid rock, regolith) and subsurface (caves, groundwater, deep ground) environments. However, it is unclear which one is more productive in terms of biomass [170] and what metabolisms have dominated those systems -and to what extent -over geologic time scales [171]. An understanding of the biology and distribution of modern microbes, which are ubiquitous in today's Earth's biosphere, seems essential for an understanding of their ancient counterparts and their impact on early terrestrial ecosystems. The genetic diversity and biomass distribution in drastically different environments [172][173][174] depict the ample range of strategies that terrestrial organisms, particularly primary producers, have developed for living on the land. Oxygenic photoautotrophy seems to be a particularly important capability of terrestrial organisms, simply because their energy source (light), reductant power (water), and carbon source (CO 2 ) are readily available in these environments. In comparison, other primary producers such as chemolithotrophs are restricted to aqueous environments because they require soluble sources of reductants (e.g., H 2 , Fe 2+ , H 2 S, HS − ) and exergonic reactions to maintain their metabolism [175]. They are also less energy-efficient than oxygenic photoautotrophs [176][177][178], and less likely dominant in subaerial environments.
Cyanobacteria have been the only organisms that developed special pigments and enzymatic capabilities for using water as a source of electrons. This process has allowed them to live outside the water in any suitable environment, even where water might be a limiting factor, such as deserts [179]. Oxygenic photosynthesis also contributed to the oxidation of the atmosphere (both by sequestering CO 2 and by producing O 2 ), a global and ongoing process with profound geochemical, atmospheric, hydrological, and biological implications [166,180,181]. Cyanobacteria and other prokaryotes, can also fix gaseous nitrogen, which seems of great advantage for an independence from dissolved N species, such as NH 4 and NO 3 [183]. The appearance of cyanobacterial akinetes (for N 2 fixation) in the Paleoproterozoic [184] attests to this early adaptation. The limiting nutrients such as P, can be supplied for organisms on land by dust deposition [185,186], which may be an alternative process for replenishment of nutrient loss by runoff and leaching in such environments [168,187]; S can also be acquired from minerals, aerosols, and as gaseous sources, likely present in the early atmosphere [188]. Thus, the nutritional requirements for oxygenic, photoautotrophic, primary producers seem not to have been a limiting factor for the colonization of the land.
On the basis of the rapid achievement of diversity and distribution of early microbial biota and from microbial successions in modern "barren" lands [189][190][191], it is expected that heterotrophic organisms were also part of land communities, as they seem to be an inevitable component in this type of consortia. Under this perspective, primitive microbial ecosystems cannot be understood as composed only of autotrophic primary producers, but also a myriad of other microbes finding their niche within such pre-existent microenvironments. For example, actinobacteria in modern cryptogamic covers (CGC) not only degrade large quantities of organic exudates from cyanobacteria, a process which influences the carbon (C) cycle, but they also seem to be structural components of these sedimentary biostructures [192]. The same applies to other bacteria such as Bacteroidetes and Proteobacteria that secrete large quantities of mucopolysaccharides, which aid in gluing soil particles together and may also have a critical role in the hydraulic conductivity of the surface substrate [193].

Microbial biodiversity in aquatic ecosystem
The emphasis on the organizational level of biodiversity responsible for ecosystem processes is shifting from a species-centered focus to include genotypic diversity. Communities with intermediate species richness show high genotypic diversity while species-poor communities do not [194]. Disturbance of these communities disrupts niche space, resulting in lower genotypic diversity despite the maintenance of species diversity.
Heterotrophic bacteria dwelling in aquatic environments are highly diverse. At coarse level the gram-positive bacteria, the Verrucomicrobiales and the Alpha-and Gamma-Proteobacteria are distributed throughout a range of aquatic habitats including marine and fresh water systems. Some phylogenetic groups appear to be adapted to more narrowly defined niches such as anoxic water and sediments (Delta-Proteobacteria) or aggregates (Bacteroidetes). Betaproteobacteria have been detected throughout freshwater habitats, but these organisms are largely absent from open ocean environments. At narrower level of identification some phylotypes are probably globally distributed as they have been detected in geographically disparate environments. High diversity of heterotrophic bacteria in aquatic environment is explained by high variety of ecological niches occurring and wide spectrum of substrates these organisms utilize.
The abundance of viruses exceeds that of Bacteria and Archaea by approximately 15-fold in the world ocean. However, because of their extremely small size, viruses represent only approximately 5% of the prokaryotic biomass because their content of matter is low. Most abundant groups of viruses found in aquatic environments are bacterioand cyanophages. The first metagenomic studies of viral communities have revealed that viral communities contain large amounts of sequences with very low homology to any described sequences available in the literature [195].

Microbial biodiversity in marine ecosystem
Microbial communities from coastal sediments vary more from one location to another than those from open waters, and have much greater community evenness [196]. Moreover, in sediments, cells are much more concentrated, resulting in a greater likelihood of interactions, which becomes even more prevalent in biofilms where cells are more densely packed. Highly productive photosynthetic microbial mats develop at the water-sediment interface. These multispecies biofilms consist of horizontally stratified layers with extremely steep gradients of light, redox potential, oxygen, sulfur species etc. The exceptionally high microbial diversity within a few microns covers a large range of metabolic groups (oxygenic and anoxygenic phototrophs, sulfate reducers, methanogens etc.) [197]. The communication mechanisms in environments (open water, sediment and biofilms), where small molecules, either diffusing from cell to cell [198], or transported by vesicles [199] or via nanotubes bridging cells [200], elicit intra-and inter-species effects that could be antagonistic or beneficial.
Microbes exhibit all of the types of social behaviour (mutual benefit, selfishness, altruism and spite) seen in multicellular organisms [201]. However, it is often difficult to categorise such behaviour in complex multi-species natural environments. Therefore, a better understanding of crude-oil biodegradation, and thus the capability to more rationally remediate contaminated environments, requires considering the mechanisms of the associations between different hydrocarbondegrading microbes and with non-degrading organisms [15]. Although fungi are considered to be largely terrestrial, they have been found in marine mats [202] and it is known that many can function in saline conditions [203], but in general salt-adapted fungi have received little attention despite a potentially major role in coastal PAH degradation. The ubiquitous co-existence of bacteria and fungi in soil and sediments [204] and their known catabolic cooperation suggests that physical interactions between them may be of importance for PAH degradation.
Marine phototrophs (primarily eukaryotic microalgae and cyanobacteria) contribute half the Earth's primary production and half of the oxygen liberated to the atmosphere [205]. However, they do not exist in isolation, and their phycosphere (loosely defined as the zone around algal cells in which bacteria feed on algal products) constitutes an important habitat that is colonised by an abundant and diverse community of heterotrophic bacteria [206,207].
Bacteria are also found living inside microalgal cells -many with unknown function [145]. The composition of free-living marine microbial communities is frequently very different from those attached to microalgae [208], with certain groups often preferring the attached lifestyle and showing higher levels of activity [194]. Moreover, different species of microalgae host distinct bacterial communities that change with time and environmental conditions [209,210]. However, there is likely to be a large spectrum of bacterial heterotroph-phototroph specificity [211], and certainly many attached bacteria can also live in the absence of a microalgal or cyanobacterial host [212]. While antagonistic interactions occur between marine phototrophs and their attached microbiota [213,214], mutualistic interactions are common. The host supplies carbon and energy sources [215], while the bacteria have been shown to provide iron [216], haem [217], vitamin B 12 [218] to consume oxygen [219] and provide protection from reactive oxygen species [220]. Symbiotic cyanobacteria supply fixed nitrogen to diatoms [221] and other algae and protests [222], and heterotrophic N 2 -fixing bacteria may also be important in interactions with microalgae, as evidenced by the abundance of alpha-proteobacterial diazotrophs in seawater size fractions of >10 μm [223].
Marine microbes offer great opportunities for biodiscovery [224,225], yet that potential is yet to be realised. Despite a huge microbial diversity, there is a lack of laboratory cultures of the microbes that are most abundant in the environment that severely limits development of biodiscovery research. Bacteria probably grow as consortia in the sea and reliance on other bacteria for essential nutrients and substrates is not possible with standard microbiological approaches. Joint et al. [226] highlighted the advantages of novel technologies, such as encapsulation into gel micro-droplets and development of consortia over standard microbiological approaches for biodiscovery programmes. These technologies, according to the authors resulted in the isolation and culturing of many previously uncultured microbes.

Conclusions
This review provides the detailed knowledge on the ability of microorganisms capable of degrading hydrocarbons which has accumulated significantly in the past two decades. Studies show that much richer microbial diversity exists in the environment that can efficiently degrade petroleum compounds. Microbial degradation processes aid the elimination of spilled oil from the environment after critical removal of large amounts of the oil by various physical and chemical methods. This is possible because microorganisms have enzymic systems that degrade and utilize different crude oil compounds as source of carbon and energy.
Microbial degradation of oil has been shown to occur by attack on aliphatic or light aromatic fractions of the oil, with high-molecularweight aromatics, resins, and asphaltenes considered to be recalcitrant or exhibiting only very low rates of biodegradation, although some studies have reported their removal at high rates under optimal conditions. The biodegradation of petroleum compounds depends on the specific microbial population present. Further studies should be carried out to identify new bacterial strains that can metabolize a broad range of compounds contained in crude oil, especially the highly persistent components. Also, a better knowledge of the diversity of catabolic pathways would certainly bring valuable information for the development of robust bioremediation processes.