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7.1 Introduction

7.1.1 Sources of PAHs in the Environment

Polycyclic aromatic hydrocarbons (PAHs) are a group of organic compounds consisting of two and more than two fused benzene rings. PAHs are naturally present in the fossil fuel. However, the increased level of PAHs in the environment over the last few decades is due to the huge increment in production and use of petroleum and petroleum products. Point sources of PAHs originate from accidental discharges during production, transportation and disposal of petroleum and its products and industrial processes such as, liquefaction and gasification of coal and waste incineration. Creosote and coal tar, which are by-products of coking, are rich source of PAHs containing 85–90% of it (Cerniglia 1992).

At contaminated sites, PAHs are often present along with other contaminants, such as, BTEX (benzene, toluene, ethyl-benzene and xylene) compounds, aliphatic hydrocarbons and heavy metals. Depending on the source of contamination, the level of PAHs in soil has been found to range from 1 to 300 ?g/kg. Air-borne PAHs originating from incomplete combustion of organic materials can give rise to high PAH levels in the atmosphere (60 ?g/m3–3 mg/m3 of air) (Bamforth and Singleton 2005). PAHs as well as numerous PAH derivatives, i.e., alkylated PAHs, nitro-PAHs, oxygenated PAHs, quinones, hydroxy and hydroxynitro compounds, are formed during incomplete combustion of various fuels such as, coal, gasoline, diesel and biofuels (Mukherji et al. 2002). Natural processes, such as volcanic eruption and forest fires are other sources of PAHs in the environment. PAHs are also formed through geochemical processes, such as, exposure of sediments to high temperature during sediment diagenesis.

7.1.2 Physical and Chemical Properties of PAHs

PAHs are a group of compounds composed of two or more fused aromatic rings in linear, angular, and cluster arrangements as shown in Fig. 7.1. Based on structural diversity, PAHs can be broadly categorized into alternant and non-alternant classes. Alternant PAHs contain only fused benzenoid rings (e.g., anthracene, phenanthrene and pyrene), whereas non-alternant PAHs contain four or five carbon aromatic rings in addition to benzene ring (e.g., fluorene and fluoranthene) (Harvey 1991). PAHs can also be classified as low molecular weight (LMW) and high molecular weight (HMW) PAHs depending upon the number of aromatic rings. PAHs containing three or less than three aromatic rings are known as LMW PAHs (e.g., naphthalene and phenanthrene) and PAHs having greater than three aromatic rings are known as HMW PAHs (e.g., pyrene, fluoranthene, chrysene). Most of the PAHs have melting points above room temperature and boiling points above 100°C and are characterized by relatively low aqueous solubility.

Fig. 7.1
figure 1figure 1

Chemical structure of some polynuclear aromatic hydrocarbons

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Greater thermodynamic stability of PAHs arises from the delocalization of ?-electron density. Thermodynamic stability and low aqueous solubility make them recalcitrant in the environment. The impact of PAH structure on its chemical behavior can generally be categorized as follows: PAHs with a linear structure are more unstable compared to their angular counterparts. Increased size and angularity of the PAH structure increase hydrophobicity and electrochemical stability which, in turn, affect their chemical and photochemical reactivity, as well as their ionization potential, vapor pressure, solubility and adsorption characteristics. The environmental fate of a PAH molecule depends on the number of aromatic rings and also on the pattern of ring linkage (Peters et al. 1999). The persistence of PAHs increases with an increase in the number of benzene rings and this is inversely correlated with the environmental biodegradation rates of PAHs.

Some PAHs contain a “bay-region” and a “K-region” (Mrozik et al. 2003). The bay- and K-region epoxides are reported to be highly reactive. Both these regions are present in phenanthrene as illustrated in Fig. 7.2. The bay-region of phenanthrene is between carbon atoms 4 and 5 whereas the K-region is between the carbon atoms 9, 10. The bay-region is a sterically hindered area. According to the Schmidt-Pullman electronic theory, K-region epoxides are likely to be more carcinogenic than the parent hydrocarbon.

Fig. 7.2
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Structure of phenanthrene showing bay-region and K-region

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LMW PAHs are characterized by higher volatility, higher solubility and greater ease of degradation than the higher molecular weight PAHs. HMW PAHs strongly sorb on to soil and sediments due to their high hydrophobicity and are therefore less available to indigenous microorganisms for degradation. The association of a contaminant with organic matter in soil (expressed as the organic matter normalized partition coefficient Koc) is a function of hydrophobicity of the compound which is expressed in terms of the octanol–water partitioning coefficient Kow. Some properties of selected PAHs are presented in Table 7.1.

Table 7.1 Physical and chemical properties of selected PAHs
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7.1.3 Toxicity of PAHs

Removal of PAHs from the environment is a major concern since they can exert acute as well as chronic toxicity (Juhasz et al. 2000). The highly hydrophobic and lipophilic nature of PAHs is responsible for their accumulation in the fatty tissue. Thus, PAHs are biomagnified through the food chain. Lower molecular weight PAHs impose greater acute toxicity because of their higher aqueous solubility compared to the higher molecular weight PAHs. PAHs can enter the human body through dermal exposure and food intake. It has been found that PAHs, especially, pyrene can get absorbed through dermal administration (75%). Acute toxicity is often reported based on lethal dose required to kill 50% of the test animals (LD50, mg/Kg-Body weight) in animal assays. LD50 value of naphthalene based on mice bioassay with oral administration is found to lie in the range of 533–710 mg/Kg while that for phenanthrene is 750 mg/Kg. The LD50 value for anthracene, fluoranthene, pyrene and benzo(a)pyrene (BaP) are > 450, 100, 514 and 232 mg/Kg, respectively, based on mice bioassays with intraperitonial exposure (Bamforth and Singleton 2005). Exposure through inhalation may occur near combustion sources since PAHs are formed during incomplete combustion of carbonaceous fuels. Bacterial assays, human cell mutagenicity assays, animal assays and epidemiological studies have demonstrated the mutagenicity and carcinogenicity of PAHs associated with combustion and ambient aerosols. Although unsubstituted PAHs are not directly genotoxic to the mammalian system, a cytochrome P450 monooxygenase enzyme mediated reaction activates them to reactive epoxides that are genotoxic (Mukherji et al. 2002). However, various substituted PAHs are directly genotoxic. USEPA has listed 17 PAHs as priority pollutants for their toxic, mutagenic and carcinogenic properties. BaP is recognized as a potent human carcinogen while several other PAHs, such as, chrysene, are also classified as possible human carcinogens (IARC 1983, 1998).

7.2 Bioremediation of PAHs

Over the past two decades, numerous researchers have demonstrated the ability of microorganisms, such as, bacteria, fungi and algae to degrade PAHs (Cerniglia 1992). Such PAH-degrading microorganisms are commonly found in soil and in the aquatic environment. The first step in bacterial degradation of PAHs involves an initial oxidative attack in the presence of dioxygenase enzyme to form cis dihydrodiol. The introduction of two hydroxy groups facilitates further degradation. Cleavage of the aromatic ring also occurs in the presence of molecular oxygen. Microbial degradation can potentially transform toxic PAHs, into benign compounds at a relatively low cost. Hence, biodegradation has gained attention for the clean-up of PAH contaminated sites. Bioremediation was adopted at numerous PAH contaminated Superfund sites in United States. Biodegradation of PAHs composed of three rings is well established and the degradation of HMW PAHs (composed of more than three rings) by bacteria has also been reported. Several bacterial genera have been identified for their ability to degrade PAHs, including Pseudomonas, Alcaligenes, Mycobacterium, Rhodococcus, Sphingomonas and Cycloclasticus. However, the microbial degradation of PAHs containing four or more aromatic rings is energetically less favorable compared to degradation of LMW PAHs (Mueller et al. 1997). BaP has been the subject of comprehensive studies on biodegradation due to its carcinogenicity (Kanaly and Harayama 2000).

7.2.1 Degradation of Four Ring PAHs

Of the four ring PAHs, degradation of fluoranthene, pyrene, chrysene, and benzo(a)anthracene has been investigated to varying degree. Fluoranthene, a non-alternant PAH containing a five carbon aromatic ring, is observed to be metabolized by a variety of bacteria, and pathways describing its biodegradation have been proposed.

7.2.1.1 Degradation of Fluoranthene by Bacteria

In 1990, two independent groups of researcher reported the isolation of a single organism capable of utilizing fluoranthene as a sole source of carbon and energy. Weissenfels et al. (1990) demonstrated the isolation of the soil microorganism Alcaligenes denitrificans strain WW1, which biodegraded fluoranthene at a rate of 0.3 mg/ml per day for an initial concentration of 1 mg/ml. It could also utilize pyrene and benzo(a)anthracene by cometabolism. A dioxygenase pathway was involved as confirmed by identification of some metabolites. Three metabolites of fluoranthene biodegradation by strain WW1, 7-hydroxyacenaphthylene, 7-acenaphthenone, and 3-hydroxymethyl-4,5-benzo-coumarine, were identified by UV, mass spectroscopy and NMR spectroscopic methods (Weissenfels et al. 1991).

The bacterium Sphingomonas paucimobilis EPA505 reported by Mueller et al. (1990) was isolated from a seven member bacterial community. This culture was capable of utilizing fluoranthene as a sole source of carbon and energy. Use of a non-ionic surfactant Tween 80 (200 mg/L) increased the aqueous solubility of fluoranthene from 1 to 28 ?M and decreased the generation time of the strain from 21 to 11 h. A resting cell suspension of EPA505 grown on fluoranthene was also capable of utilizing other four ring and five ring PAHs, such as, benzo(b)fluorene, benzo(a)anthracene, chrysene and pyrene.

In batch cultures with 1.66 mg/L fluoranthene, Kelley and Cerniglia (1991) reported that Mycobacterium sp. strain PYR-1 was capable of degrading greater than 78% fluoranthene within 5 days. When added to soil and water microcosms, Mycobacterium strain caused seven-fold increase in mineralization of fluoranthene over the indigenous microorganisms. Although mineralization of fluoranthene to CO2 was rapid, some metabolites were identified: 8-hydroxy-7-methoxyfluoranthene, 9-hydroxyfluorene, 9-fluorenone, 9-fluorenone-1-carboxylic acid, 1-acenaphthenone, 9-hydroxy-1-fluorene carboxylic acid, phthalic acid, 2-carboxybenzaldehyde, benzoic acid, phenyl acetic acid, and adipic acid.

7.2.1.2 Degradation of Pyrene by Bacteria

The first report of pyrene degradation as sole source of carbon and energy was demonstrated for a Rhodococcus sp. strain UW1 (Walter et al. 1991). Within 2 weeks, 72% (initial concentration 500 mg/L) was mineralized to CO2. A metabolite of molecular formula C16H10O4 was identified. It was formed as a result of recyclization of the direct meta- fission product of pyrene. This strain could also utilize chrysene as sole source of carbon and energy.

Various Mycobacterium species have been reported for their pyrene degrading ability and the pathways have also been hypothesized based on identification of metabolites. Heitkamp and Cerniglia (1988) isolated a Mycobacterium sp. from sediments near a hydrocarbon source. This culture could mineralize pyrene when grown in mineral salts medium supplemented with organic nutrients. Pyrene induced culture mineralized over 60% of radiolabeled pyrene within 96 h. Enzymes responsible for pyrene catabolism seemed to be inducible, since a long lag phase in pyrene mineralization was observed in cultures grown in the absence of pyrene. In contrast, no pyrene mineralization was observed in non-induced culture. Seven metabolites of pyrene metabolism were detected by high-performance liquid chromatography. This included three ring oxidation products, cis-4,5-pyrene dihydrodiol, trans-4,5-pyrene dihydrodiol, and pyrenol. Four ring fission products were also observed, i.e., 4-hydroxyperinaphthenone, 4-phenanthroic acid, phthalic acid, and cinnamic acid. 4-phenanthroic acid was the major metabolite. Pyrenol was possibly formed by non-enzymatic dehydration of pyrene dihydrodiols or by oxidative metabolism of pyrene by the Mycobacterium. Multiple pathways for initial oxidative attack on pyrene was suggested since both cis- and trans-4,5-dihydrodiols were identified as metabolites. Further studies confirmed that both dioxygenase and monooxygenase enzymes were secreted by this microorganism. Sediment microcosms from where this strain was isolated showed enhanced mineralization of various PAHs, including pyrene and BaP when inoculated with the Mycobacterium strain (Heitkamp and Cerniglia 1989). Although this strain could utilize pyrene as primary substrate, pyrene degradation was inhibited by increasing the organic nutrients in the microcosm. Nutrient depletion was possibly caused by the utilization of organic nutrients by the rapidly growing indigenous microorganisms.

Later, other Mycobacterium species, i.e., Mycobacterium sp. strain RJGII-145 (Grosser et al. 1991), Mycobacterium flavescens (Dean-Ross and Cerniglia 1996) and Mycobacterium sp. strain KR2 (Rehmann et al. 1998) have been reported to utilize pyrene as a sole source of carbon and energy. Mycobacterium sp. strain RJGII-135 was isolated by Grosser et al. (1991) from soil near a coal gasification plant. It utilized pyrene as sole source of carbon and energy. When the pyrene induced culture was reintroduced in the soil containing pyrene, enhanced mineralization of pyrene was observed. Pyrene mineralization reached 55% within 2 days, compared to 1% for the indigenous population. Metabolites of pyrene degradation were also identified through further studies (Schneider et al. 1996). Three stable intermediates, 4-phenanthrene carboxylic acid, 4,5-pyrene dihydrodiol, 4,5-phenanthrene dicarboxylic acid were formed within 4–8 h after the start of the experiment.

Mycobacterium sp. strain BB1 was isolated from a former coal gasification site (Boldrin et al. 1993). Fluoranthene and pyrene were used as sole source of carbon and energy by this culture. It exhibited a maximum growth rate of 1.2 mg/L/h for large crystals of pyrene and 5.6 mg/L/h for small crystals during exponential growth in a 1.6 L fermenter with an initial concentration of 0.5 g/L of pyrene. This strain was used to examine the effects of various culture conditions, such as degradation of pyrene at low dissolved oxygen concentrations (Fritzsche 1994), utilization of PAHs in mixtures (Tiehm and Fritzsche 1995) and degradation of pyrene in presence of non-ionic surfactants. Jimenez and Bartha (1996) used a Mycobacterium sp. in solvent-augmented mineralization of pyrene. The cells, which physically adhered to solvent droplets, showed 8.5 times faster rate of pyrene degradation compared to cells in suspension. Another Mycobacterium sp. strain CH1 isolated from PAH contaminated freshwater sediments could mineralize pyrene and fluoranthene as the sole carbon and energy source. This culture was also capable of growth on a wide range of branched alkanes and n-alkanes (Churchill et al. 1999). Gordona sp. strain BP9 and Mycobacterium sp. strain VF1 were isolated from hydrocarbon-contaminated soil and each was capable of utilizing fluoranthene and pyrene as sole carbon and energy sources (Kastner et al. 1994). Reintroduction of BP9 into soil after growth on pyrene (200 mg/L) as sole source of carbon and energy exhibited six-fold increase in pyrene metabolism compared to native un-inoculated soil microorganisms (Kastner et al. 1998). The rate and extent of pyrene degradation by various pure cultures are listed in Table 7.2.

Table 7.2 Rate and extent of pyrene degradation by pure cultures
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7.2.2 Limitations in Degradation of PAHs in Soil

Many microorganisms are metabolically capable of mineralizing or partially transforming PAHs in the environment. Some fungi can produce extracellular enzymes, but for most of the bacteria, the pollutants must cross the cell membrane to get direct access to the enzymes or at least attach to the membrane bound enzymes. With the exception of some bacterial species that can attach directly to the interface between the organic and aqueous phase, in general for degradation to take place, the contaminants must be solubilized in the aqueous phase. The two major mass transfer limitations are caused by: (a) compound availability and (b) crossing of the membrane.

7.2.2.1 Compound Availability

Bioavailability of a compound depends on its physico-chemical properties. Aqueous solubility plays the most important role in determining the potential for bioremediation. Solubility of PAHs is decreased with increase in the number of rings. Very low aqueous solubility and high hydrophobicity (high Kow) are the two governing factors responsible for the recalcitrance of four and greater than four ring PAHs. Higher ring PAHs often exist in a separate phase, i.e., as components of non-aqueous phase liquids (NAPLs) or remain strongly sorbed on to soil. Mukherji and Weber (1998, 2001) demonstrated that mass transfer of naphthalene from a NAPL affected its biodegradation rate. However, some cultures are reported to overcome bioavailability limitations by virtue of enhanced cell surface hydrophobicity (CSH) as discussed later.

7.2.2.2 Crossing of Membrane

Another challenge is for PAHs to cross the cell membrane of the microorganism for gaining access to the enzymes present inside the cell. Crossing the membrane is required even for membrane bound enzymes since they are located on the inner side of the cell membrane. Both sides of the membrane are considered fairly polar due to the presence of polar head groups on the phospholipids that comprise the membrane while the center region is non-polar in nature due to the fatty acid moieties of the phospholipids. In addition to the polarity obstacles, transport through the cell membrane is also affected by the membrane associated proteins and peptidoglycan. In gram-negative bacteria, an additional outer membrane needs to be crossed. For a contaminant to overcome all the cell barriers and access the enzymes inside the cell, it must either be selectively taken into the cell, or it must have the right combination of polarity, size, and functional groups to get across the cell membrane. Bressler and Gray (2003) surveyed the literature and determined the maximum biodegradation rate of various PAHs under aerobic condition. The maximum biodegradation rate was correlated with log Kow. PAHs with log Kow values above three demonstrated reduced biodegradation. This limitation is mainly due to the poor aqueous solubility and adsorption properties of these compounds, however, the repulsion of these highly non-polar compounds by the polar regions of the cell membrane cannot be ignored. Compounds with log Kow values below one also demonstrated reduced rates of biodegradation possibly due to reduced ability of these extremely polar compounds to diffuse across the non-polar regions of the cell membrane.

Bressler and Gray (2003) derived an expression for maximum flux through the membrane (Fmax) as a function of the octanol water partition coefficient (Kow), molecular weight (MW) and aqueous solubility (Caq) of a compound (Eq. 7.1) and computed the values for some known environmental contaminants. They assumed the membrane crossing step as the rate limiting step in biodegradation assuming that the intracellular concentration of a compound is maintained as zero by the active enzyme and demonstrated a linear relationship between the maximum reported biodegradation rate and the maximum membrane flux for various environmental contaminants and fitted a regression line using the data set.

$$ {\text{F}}_{\hbox{max}} = 0. 0 0 3 { }\left( {{\text{K}}_{ow} {\text{C}}_{aq} / \, \hbox{MW}^{ 0. 5} } \right) $$
(7.1)

They concluded that compounds with biodegradation rate falling significantly above the regression line (such as, toluene and phenol) are not transported by diffusion through the lipid bilayer. In contrast, those compounds having biodegradation rate falling significantly below the regression line are limited by the rate of enzymatic conversion rather than by membrane permeation. Biodegradation of fluorene, chrysene and several other PAHs falling on the regression line are limited by the membrane permeation. In contrast, the maximum biodegradation rate of the four ring PAHs, fluoranthene and pyrene was well below the predicted line. Thus, biodegradation of pyrene and fluoranthene is limited by the synthesis of key enzymes.

7.2.3 Surfactant Enhanced Biodegradation of HMW PAHs

Contradictory and inconclusive results have been observed regarding the effects of surfactants on the biodegradation of PAHs at surfactant concentration below and above the critical micelle concentration (CMC) (Margesin and Schinner 1999). Some studies showed an enhanced effect in which not only LMW PAHs were degraded successfully, but even compounds containing four and more than four rings were also degraded effectively. In contrast, in other studies, degradation was found to be inhibited. While some researchers believe that PAHs solubilized within surfactant micelles are completely bioavailable, yet others believe that they are only partially bioavailable or not at all bioavailable (Guha and Jaffe 1996; Luning Prak and Pritchard 2002 ). Volkering et al. (1995) reported a strong inhibition in biodegradation of polynuclear aromatic hydrocarbons (PAHs) in the presence of surfactants above their CMC. Surfactants may also directly interact with the bacterial cell surfaces, the interaction being dependent both on the type of surfactant and the type of microorganisms. Efroymson and Alexander (1991) reported that Triton X-100 prevented the adherence of cells to an organic-aqueous interface, however, this had a beneficial effect on naphthalene biodegradation by Arthrobacter sp. Since naphthalene has relatively high aqueous solubility, it was available both from the aqueous phase as well as from the organic-water interface. Hindered adherence may have adversely affected the uptake of HMW PAHs with very low aqueous solubility. However, Triton X-100 is also reported to promote the attachment of Burkholderia cultures to the NAPL-water interface (Mohanty 2010). Some surfactants are toxic to bacterial cultures while enhanced solubilization of toxic components from a multi-component NAPL phase such as petroleum caused by the surfactants, may adversely affect bacterial cultures.

Thibault et al. (1996) used two pyrene degrading soil Pseudomonas species in soil reinoculation experiments to test the effects of four surfactants, namely, Simple Green, Biosolve, Witconol SN70 and Sodium dodecyl sulphate (SDS). Witconol SN70 caused maximum enhancement in solubilization and mineralization of pyrene under unsaturated conditions. However, when pyrene degraders were inoculated into soil slurries, degradation of pyrene was highest in the absence of surfactants. It was concluded that increased solubilization of pyrene in the soil slurry, caused toxicity to the microorganisms.

Boonchan et al. (1998) evaluated various surfactants for their impact in improving the biodegradation of four, five and seven ring PAHs by Strenotrophomonas maltophilia VUN 10,010, isolated from manufactured gas plant site soil. All the cationic and anionic surfactants tested were found to be highly toxic for the bacterial strains while the Tween series was utilized as growth substrate. Five non-ionic surfactants (Brij 35, Igepal CA-630, Triton X-100, Tergitol NP-10, and Tyloxapol) increased the apparent solubility of fluoranthene, pyrene and B(a)P at least 250 fold at 10 g/L surfactant concentration. They were less toxic and were not used as growth substrates. Pyrene and also the five and seven ring PAHs were utilized by the culture as sole source of carbon and energy. Not only the solubility, but also the rate of pyrene degradation by strain VUN 10,010 was enhanced by the addition of the four non-ionic surfactants (5–10 g/L). However, an inhibition in degradation rate was observed with Igepal CA-630 (5 g/L). The specific growth rate of VUN 10,010 on pyrene was increased by 67% in the presence of 10 g/L Brij 35 or Tergitol NP-10. The addition of Brij 35 and Tergitol NP-10 to media containing a single HMW PAH (four and five benzene rings) as the sole carbon source increased the maximum specific PAH degradation rate and decreased the lag period normally seen for PAH degradation. The addition of Tergitol NP-10 to VUN 10,010 cultures which contained a PAH mixture (three to seven benzene rings) substantially improved the overall degradation rate of each PAH and increased the specific growth rate of VUN 10,010 by 30%. Their results suggested that besides the improvement in dissolution rate, surfactants may also facilitate the transport of PAHs through the cell membrane.

Doong and Lei (2003) observed a correlation between the polyoxyethylene chain length and the solubilization efficiency of surfactants. Four non-ionic surfactants and one anionic surfactant were found to enhance the bioavailability of naphthalene, phenanthrene and pyrene with efficiencies ranging from 21.1 to 60.6, 33.3 to 62.8 and 26.8 to 70.9%, respectively. The trend in degradation efficiency was as follows: Brij 30 > Triton X-100 > Tween 80 > Brij 35. The HLB values of Brij 30, Triton X-100, Tween 80 and Brij 35 were 9.7, 13.5, 15.0 and 16.9, respectively. With increase in HLB value, the degradation potential of the surfactant was decreased. Triton X-100 and Brij 30 were utilized as sole source of carbon and energy by Pseudomonas putida. Although the dissolution of PAHs was increased, Brij 35 and Tween 80 inhibited the growth of P. putida. Increase in surfactant concentration also decreased the mineralization rate of pyrene possibly due to mass transfer limitation from the micelle. Mass transfer processes are proportional to concentration gradient of PAHs inside the micelle so that when the PAHs are diluted in a larger micellar mass, transfer to the cells may decrease. Another possibility may be catabolic repression due to preferential utilization of the surfactant.

Willumsen and Arvin (1999) studied the impact of Triton X-100 on fluoranthene degradation by S. paucimobilis EPA505. Although the surfactant enhanced mineralization of fluoranthene, it was much lower compared to the increase in solubilization. A model which described the degradation kinetics of Triton X-100 solubilized fluoranthene by S. paucimobilis EPA505 cells was also developed after accounting for micellar solubilization, metabolite accumulation and its bioavailability considerations. In the presence of Triton X-100 and calcium, the mineralization rate of fluoranthene by S. paucimobilis EPA505 was almost doubled (Willumsen and Karlson 1998). In contrast, absence of calcium inhibited fluoranthene mineralization possibly due to the adverse effect of Triton X-100 on the cytoplasmic membrane.

In a study involving the Tween series surfactants (Tween 20, Tween 40, Tween 60 and Tween 80), Kim and Weber (2003) reported enhanced solubilization of phenanthrene at different doses of surfactant above their CMC. However, the micelle solubilized phenanthrene was not bioavailable to S. paucimobilis EPA 505. Preferential microbial uptake of the hydrophobic part of the chemical surfactant caused destabilization of the micelles due to an imbalance in the amphiphilic molecule. Thus, phenanthrene was released from the micellar core such that its biodegradation was increased.

7.3 Bacterial Mechanisms for Overcoming Bioavailability Limitation

It is generally believed that most bacteria can access and degrade pollutants only when they are dissolved in water. However, some PAH-degrading bacteria are reported to overcome the bioavailability limitations for enhancing uptake through (1) direct interfacial uptake and (2) uptake facilitated by secretion of biosurfactants.

7.3.1 Direct Interfacial Uptake

In direct interfacial uptake, bacteria attach itself to the surface of the NAPL droplets which are larger in size compared to the bacterial cell. Substrate uptake takes place at the point of contact through diffusion. CSH plays an important role in direct uptake. Microorganisms capable of direct uptake can modify their cell surface outer membrane to enhance CSH. In this case, substrate uptake does not depend on the dissolution rate. Some microorganisms have surface structures such as, fimbriae and capsules which help them to attach to the NAPL-water interface such that they can directly uptake substrate from a separate phase.

CSH is an important parameter affecting the adhesion behavior of bacterial cells to surfaces and interfaces. CSH is commonly measured by determining the contact angle (Busscher et al. 1984) of a uniform layer of cells. Some other assays for hydrophobicity include hydrophobic interaction chromatography (Dillon et al. 1986), salt aggregation test (Lindahl et al. 1981), polystyrene microsphere attachment (Rosenberg 1981) and microbial attachment to microsphere (Zita and Hermansson 1997). The bacterial adhesion to hydrocarbons (BATH) assay (Rosenberg et al. 1980) is also used for determining the potential of a culture for direct interfacial uptake.

BATH assay is a simple and quantitative method for determining the degree of adherence of bacterial cells to a specific liquid hydrocarbon, i.e., n-hexadecane when an aqueous suspension of bacterial cells is mixed with the hydrocarbon. Adherence is directly related to change in absorbance of the aqueous suspension with respect to the control. However, BATH assay results are affected by solution phase interaction between the NAPL and bacterial culture, hence it is not a true measure of CSH. Bacterial cell surface charge is reported to vary with pH, growth medium, culture age and ionic strength. At a certain pH, the maximum adherence occurs when either the bacterial culture or the NAPL droplet is uncharged. Busscher et al. (1995) reported the zeta potential of various solvents such as, n-hexadecane, toluene and xylene ranges from ?63 to ?18 mV at pH 7. Chakraborty et al. (2010) demonstrated the effect of NAPLs of varying composition serving as growth substrate on adherence measured in BATH assay and surface charge of three bacterial strains Burkholderia cepacia (ES1) and Burkholderia multivorans (NG1 and HN1). All the cultures used depicted growth on the NAPLs composed of n-hexadecane, naphthalene, phenanthrene and pyrene in varying proportion. Change in NAPL used as growth substrate affected both the zeta potential and CSH of the cultures (based on contact angle measurements), which in turn affected their adherence to the NAPL-water interface determined using the BATH assay. Zeta potential was found to be close to zero at pH 2 for all the cultures. For each culture a strong inverse correlation was observed between zeta potential and adherence to n-hexadecane. Thus, BATH assay results are the outcome of both hydrophobic and electrostatic interaction between the bacterial culture and NAPL phase.

Some bacterial cultures are capable of inducing high CSH when only hydrocarbon/oil is provided as sole substrate. This enhancement in hydrophobicity is reported to facilitate the direct interfacial uptake of hydrocarbons from NAPLs and the uptake of solid hydrocarbons. Most PAHs other than naphthalene and substituted naphthalenes have low aqueous solubility of the order of 1 mg/L or less so that direct interfacial uptake may provide a distinct advantage to the microorganism utilizing it as a substrate. Induction of CSH has been demonstrated by various researchers by growing cultures both in the presence of a readily degradable soluble substrate and in the presence of hydrocarbons/oil. Subsequently, the CSH is determined through contact angle measurements or BATH assay. Pseudomonas sp. strain PP2 grown on phenanthrene depicted greater adherence to aliphatic (hexane and heptane) and aromatic hydrocarbons (benzene and xylene) compared to the dextrose grown cells (Prabhu and Phale 2003). These changes were attributed to growth associated with production of an extracellular biosurfactant. Mycobacterium sp. LB501T is reported to form a biofilm on solid anthracene provided as sole substrate (Wick et al. 2002). Such a biofilm formation was not observed when glucose was provided as a co-substrate along with anthracene. This was attributed to surface modifications that caused enhancement in hydrophobicity of anthracene grown cells. In adhesion experiments, anthracene grown cells demonstrated 1.5–8.0 fold greater adherence to Teflon and up to 70 times greater adherence to anthracene surfaces compared to glucose grown cells (Wick et al. 2002). This activity was linked to specific changes in the phospholipid fatty acid (PLFA) and glycolipid fatty acid (GLFA) patterns (Wick et al. 2003). In another study, mineralization of fluoranthene by a Mycobacterium strain was facilitated by direct attachment to fluoranthene provided as sole substrate (Willumsen and Karlson 2001). High CSH and adherence to emulsified solvent droplets containing dissolved pyrene facilitated its mass transfer to the degrading bacteria, during solvent-augmented biodegradation by a Mycobacterium sp. (Jimenez and Bartha 1996).

Similar enhancement in hydrophobicity and attachment to NAPLs has been reported for uptake of aliphatic hydrocarbons from NAPLs, such as n-hexadecane, diesel and petroleum (Amin et al. 1996; Mohanty and Mukherji 2008). Thus, cultures can actively regulate their CSH as a strategy for optimizing the uptake of substrates characterized by low aqueous solubility. In addition to extracellular biosurfactants, the extracellular polymeric substances (EPS) associated with bacterial cells may play a role in bioavailability enhancement and enhanced attachment to NAPLs. The role of EPS in affecting attachment to NAPLs was implicated for an alkane degrading Rhodococcus strain Q15 (Whyte et al. 1999). For Rhodococcus sp. strain 094, a good correlation was observed between CSH, alkane oxidation and EPS activity (Berdholt et al. 2002). EPS release in the stationary phase was correlated with decrease in hydrophobicity and emulsification of the culture broth. Membrane bound vesicles with some emulsifying activity were responsible for CSH in Acinetobacter venetianus RAG-1 and Acinetobacter sp. strain H01-N (Leahy et al. 2003). Mohanty and Mukherji (2008) demonstrated the effect of growth substrates by conducting the BATH assay using two oil degrading bacterial cultures B. cepacia ES1 and Exiguobacterium aurantiacum AS1 with two different NAPL types, i.e., diesel and n-hexadecane. For both the cultures, adherence to a NAPL phase was highest when the corresponding NAPL was used for culture growth. For dextrose grown cultures of B. cepacia and E. aurantiacum, negligible adherence to NAPL phase was observed irrespective of the NAPL phase used in the assay.

7.3.2 Production of Biosurfactants and PAH Uptake

Production of biosurfactant is another mechanism adopted by certain bacterial species to degrade complex petroleum hydrocarbons. Like chemical surfactants, biosurfactants are amphiphilic molecules having both hydrophilic and hydrophobic regions which help them in attachment to different surfaces. Biosurfactants are categorized into groups based on their MW (Ron and Rosenberg 2002). LMW surfactants are typically glycolipids where the carbohydrate group is attached to long chain alkyl acids or lipoproteins. Rhamnolipids, trehalolipids and sophorolipids are a few known biosurfactants that come under this category. Several species of Pseudomonas are reported to produce rhamnolipids. External addition of rhamnolipids was reported to have varying impacts on CSH of bacterial cultures degrading phenanthrene, i.e., Bacillus subtilis BUM and P. aeruginosa P-CG3, which, in turn, was found to alter the relative contribution of the two cultures in phenanthrene biodegradation (Zhao et al. 2011). External addition of a sophorolipid mixture produced by Candida bombicola has been shown to enhance phenanthrene biodegradation by Pseudomonas yanoikuyae through enhanced solubilization (Schippers et al. 2000).

HMW biosurfactants are composed of polysaccharides, proteins, lipopolysaccharides, lipoproteins or a mixture of these polymers. They are less effective in reducing interfacial tension compared to LMW biosurfactants. However, they can create a hydrophilic coat around the oil droplets which inhibits coalescence. The most studied are surfactants produced by different species of the genera Acienetobacter. Emulsan produced by Acinetobacter RAG-1 and alasan produced by Acinetobacter radioresistens are a complex mixture of heteropolysaccharides and proteins. One of the protein moieties of 45 kDa in alasan has been studied extensively and reported to be the most important structural component determining the activity of the complex. It contributed to the formation of stable oil–water emulsion (Navon-Venezia et al. 1998; Toren et al. 2001). Only 500 ?g/ml of alasan increased the aqueous solubility of phenanthrene, fluoranthene and pyrene by 6–25 times. The presence of alasan more than doubled the mineralization rate of the PAHs by S. paucimobilis EPA 505 (Barkay et al. 1999).

Biosurfactants are usually produced during the stationary growth phase and this is induced by molecular signals involved in quorum sensing (Ron and Rosenberg 2002). Emulsification of oil by biosurfactants increases bioavailability by increasing the surface area. Biosurfactants are also reported to cause adhesion/detachment of bacteria by regulating their CSH. Biosurfactants may be cell bound or may be released into the medium. Cell bound biosurfactants may be bound to the hydrophobic region in the outer layer of the cell surface thereby exposing the hydrophilic part to the outside and thus the bacteria will be able to interact with hydrophilic surfaces but not hydrophobic surfaces (Nue 1996). The reverse orientation is also possible, so that the surfactant is bound through the hydrophilic part to the cell surface thereby exposing the hydrophobic part to the environment. This would allow the bacteria to interact with hydrophobic surfaces only. CSH of Pseudomonas aeruginosa was increased by the presence of cell bound biosurfactant rhamnolipids whereas, for Acinetobacter strain, reduced CSH has been reported due to presence of cell bound bioemulsifier (Nue 1996).

Deziel et al. (1996) first reported the production of rhamnolipid biosurfactant by P. aeruginosa 19 SJ for facilitating growth on the PAHs, naphthalene and phenanthrene. Maximum extracellular biosurfactant production was observed at the onset of the stationary phase when high cell density limited the availability of the substrate. Production of the biosurfactant enhanced solubilization of the substrate. Prabhu and Phale (2003) also indicated the role of extracellular biosurfactants in phenanthrene biodegradation. Bordoloi and Konwar (2009) reported biosurfactant production by various strains of P. aeruginosa isolated from petroleum contaminated soil in Assam, India. All the strains were capable of degrading the PAHs, fluorene, phenanthrene and pyrene as sole substrate. The biosurfactants produced by the various strains lowered the surface tension up to 30–32 mN/m and was characterized by CMC in the range 100–110 mg/L. Differences were observed in the nature of the various biosurfactants. The lipopeptide biosurfactant produced by strains MTCC 7815 and 7812 depicted the best pyrene solubilizing activity. Strain MTCC 8163 produced a biosurfactant that was proteino-starchy in nature, while strains MTCC 8165 and MTCC 7814 produced structurally complex biosurfactants comprising of protein, carbohydrates and lipids. Enhanced solubilization by the biosurfactants was linked to sustained growth on the PAHs. B. subtilis DM04 and a muccoidal (M) and non-muccoidal (NM) strain of P. aeruginosa isolated from petroleum contaminated soil in north-eastern India were reported to degrade pyrene as sole carbon and energy source with the help of biosurfactants that enhanced solubilization (Das and Mukherjee 2007).

7.4 Degradation of HMW PAHs and Substrate Interaction Effects

In the environment, PAHs exist as complex mixtures along with a multitude of other compounds, such as, in creosote and coal tar. In this scenario, one compound may affect the rate and extent of degradation of another compound, e.g., through increasing the biomass growth, competitive inhibition and co metabolism. These effects are referred as substrate interaction effects. As observed by various researchers, the net result could be an increase in the rate of biodegradation as a result of enhanced biomass growth (Beckles et al. 1998), a decrease in biodegradation rate due to competitive inhibition or hindrance in enzyme induction (Guha et al. 1999; Molina et al. 1999), or no noticeable effect because the opposing effects cancel out (Beckles et al. 1998; Guha et al. 1999). Degradation of a difficult to degrade substance may be enhanced due to enhancement in the biomass population caused by an easily degradable substrate. The interactions and effects encountered in a multi-substrate system are a function of the microbial community, the type of culture (mixed versus pure) and the physiological state of the community. In general, the biodegradation rates of the more degradable and abundant compounds are reduced due to competitive inhibition, while enhanced biodegradation occurs for the recalcitrant PAHs due to simultaneous increase in biomass growth on multiple substrates. Studies by Guha et al. (1999), using naphthalene, phenanthrene and pyrene as sole substrate, and in binary and ternary mixtures revealed that in ternary mixture, biodegradation of naphthalene was inhibited while the biodegradation rates of phenanthrene and pyrene were enhanced. Knightes and Peters (2006) studied the interactions in binary mixtures and in a complex mixture using nine PAHs, i.e., naphthalene, 1-methylnaphthalene, 2-methylnaphthalene, 2-ethylnaphthalene, phenanthrene, anthracene, pyrene, fluorene and fluoranthene. They reported both positive and negative interactions between the substrates in these multi-substrate systems.

Competitive inhibition, i.e., retardation in rates of degradation of a substrate may be caused as a result of competition for the same enzyme by other substrates. For example, cis-naphthalene dihydrodiol dehydrogenase has a broad substrate specificity and oxidizes cis-dihydrodiols of many PAHs beside naphthalene so that a competitive inhibition is observed in naphthalene biodegradation. Such negative effects due to competitive inhibition may be observed where at least one step in the multistep transformation of a PAH occurs through a common enzyme. Biodegradation pathways for a wide range of PAHs are similar, and a dioxygenase enzyme is involved in the initial step of aerobic biodegradation. Thus, competitive inhibition is quite prevalent. The degradation of phenanthrene by Pseudomonas stutzeri P-16 and P. saccharophila P-15 isolated from a creosote contaminated soil was inhibited by naphthalene, 1-methylnaphthalene, 2-methylnaphthalene and fluorene. These cultures could also utilize naphthalene, 1-methylnaphthalene and 2-methylnaphthalene as sole substrate but were unable to utilize fluorene (Stringfellow and Aitken 1995). Bouchez et al. (1995) observed that a Pseudomonas sp. strain S Ant Mu5 degraded anthracene to a greater extent when it is supplied alone compared to when it is supplied in a mixture with fluorene.

Positive effects and enhanced degradation of substrate may occur as a result of proliferation of biomass on multiple substrates. Since degradation rate is a function of the biomass concentration, the biodegradation rate for each compound in a mixture may be increased if common bacterial cultures are responsible for their utilization. The presence of a suitable PAH substrate or pre-exposure to suitable PAH can facilitate the degradation of other PAHs through enzyme induction. Bauer and Capone (1988) observed enhanced degradation of PAHs as a result of pre-exposure to other aromatic hydrocarbons. Naphthalene and phenanthrene enhanced the degradation of various other PAHs through cross acclimation (Bauer and Capone 1988). Exposure to certain compounds may block enzyme induction. Such malefic association has been reported for a pyrene degrading Rhodococcus strain S Pyr Na 1 in the presence of either phenanthrene or fluoranthene (Bouchez et al. 1995). In another study, phenanthrene degradation by several strains was found to decrease due to lack of pre-exposure to specific PAHs (Molina et al. 1999). McLellan et al. (2002) reported that the addition of phenanthrene as a second substrate could increase benzo[a]pyrene metabolism by Mycobacterium sp. strain RGJII-135 up to six-fold. Pyrene and phenanthrene were potent inducers of benzo[a]pyrene degradation activity.

Co-metabolism of a non-growth substrate can occur in the presence of a growth substrate, however, the products of co-metabolic transformation may hinder degradation of the growth substrate. Partial co-metabolic transformation of HMW PAHs can yield more water soluble intermediates which may be degraded by other bacterial strains. PAH-degrading bacteria, Burkholderia sp. VUN10013 was found to utilize and degrade low MW PAHs, such as phenanthrene and anthracene, but not the HMW PAHs, pyrene, fluoranthene, chrysene and benzo[a]pyrene (Somtrakoon et al. 2008). However, when present in a mixture with phenanthrene or anthracene, the HMW PAHs, pyrene and fluoranthene could be co metabolically degraded by this strain. Phenanthrene was better at enhancing the transformation of the non-growth substrates compared to anthracene. Presence of the HMW PAHs adversely affected the degradation of anthracene, but not that of phenanthrene. Bouchez et al. (1995) investigated the degradation of PAHs in binary mixtures using six bacterial strains, including two Rhodococcus spp. capable of growth on pyrene and fluoranthene. All individual strains were capable of transforming the PAHs co-metabolically and both inhibition and synergistic interactions were observed. The ability of Pseudomonas sp. strain S Phe Na l to degrade phenanthrene was partially inhibited by the transformation products formed by fluorene co metabolism. Moreover, the presence of fluoranthene as a second PAH slightly decreased (92%) the degradation of phenanthrene by Pseudomonas sp. strain S Phe Na l. Inhibition was most commonly observed when the added PAH was more water soluble than the PAH added originally. It was also observed that mineralization yields were higher and biomass yields were lower for HMW PAH-degrading bacteria than for LMW PAH-degrading bacteria.

Luning Prak and Pritchard (2002) reported the degradation of mixtures of PAHs, pyrene, fluoranthene and phenanthrene by S. paucimobilis EPA 505 in the presence of the surfactant, Tween 80. This strain could utilize both phenanthrene and fluoranthene as growth substrates, but could not utilize pyrene. In mixtures, phenanthrene was preferentially degraded over fluoranthene, and preference for pyrene was the least. Competing substrates hindered degradation, however, after degradation of the competing substrate, the other PAHs were degraded at rate higher than or comparable to the single-substrate system.

7.5 Degradation of PAHs in Presence of Other Hydrocarbons

The effect of the presence of aliphatics on the rate and extent of degradation of PAHs has not been studied extensively. Ghosh et al. (2010) demonstrated negligible degradation of pyrene present in a NAPL along with other constituents, i.e., n-alkanes and LMW PAHs although the Sphingomonas sp. used was grown on pyrene provided as sole source of carbon and energy. Zytner et al. (2006) used a synthetic diesel containing the aliphatics: dodecane, tetradecane, heptadecane, pristane, octadecane and eicosane together contributing to 90%. The remaining 10% was composed of naphthalene and phenanthrene. Biodegradation rate was studied in 1-L bioreactors containing soil spiked with the synthetic diesel/individual compounds. Degradation by heterotrophic microorganisms indigenous to the Elora silt loam soil was studied and the fraction of contaminants remaining was analyzed over time by GC-FID over a period of 30 days. The degradation trend observed for the synthetic diesel fuel was: C12 > naphthalene > C17 > C19 > C20 > phenanthrene. The order changed when the compounds were degraded individually: C17, C19, C20, C12 > phenanthrene > naphthalene. The amount degraded was also changed in the mixture compared to individual compound. For example, phenanthrene degradation decreased to 9% in the synthetic diesel mixture compared to 26% when present individually. This difference may be due to the competitive inhibition.

7.6 Conclusions

Bacterial degradation of HMW PAHs with more than four rings has been conclusively demonstrated in-spite of their structural complexity and low aqueous solubility. Pyrene, fluoranthene and other HMW PAHs can serve as sole source of carbon and energy for naturally occurring microorganisms. The degradation pathways have been inferred for various organisms through identification of metabolites. The degradation rates are often found to be low due to either bioavailability limitations or membrane flux limitations. In engineered systems, the bioavailability limitations may be overcome by the addition of surfactants, however, chemical surfactants are also reported to cause adverse effects on biodegradation. Bacteria degrading HMW PAHs are often found to induce high CSH that promotes adherence to solid PAHs, NAPLs or sorbed PAHs. Alternatively, some bacteria gain a selective advantage by secretion of biosurfactants. Biosurfactants are reported to increase uptake through enhancement in solubilization, emulsification and also by inducing changes in CSH. The presence of HMW PAHs along with other easily degradable PAHs and other hydrocarbons (i.e., n-alkanes) may adversely affect their degradation due to competitive inhibition. However, beneficial effects such as, cometabolic degradation of HMW PAHs are also reported. Bioavailability limitations together with adverse substrate interaction effects may be largely responsible for the persistence of HMW PAHs in the environment.