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Bioremediation of Petroleum oil Contaminated Soil and Water



P.K. Jain, V.K. Gupta, R.K. Gaur, M. Lowry, D.P. Jaroli and U.K. Chauhan
 
ABSTRACT

Environmental pollution with petroleum and petrochemical products (complex mixture of hydrocarbon) has been recognized as one of the most important serious current problem. People working in garage etc. are always exposed with oily sludge which are potent immunotoxicants and carcinogenic. Accidental leakages from petroleum carrying ships lead to oily layers over the water surface, possessing great threat to the existing flora and fauna. The currently used physical and chemical treatments are effective for the degradation of petroleum products but they lag behind in the desired properties, apart they frequently produce many hazardous compounds which are potent immunotoxicants and carcinogenic for living beings. In contrast, bioremediation is effective treatment in terms of efficacy, safety on long terms use, cost and simplicity of administration. However, for the foreseeable future, long term tolerance studies are needed before being recommended for large scale use.

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P.K. Jain, V.K. Gupta, R.K. Gaur, M. Lowry, D.P. Jaroli and U.K. Chauhan, 2011. Bioremediation of Petroleum oil Contaminated Soil and Water. Research Journal of Environmental Toxicology, 5: 1-26.

DOI: 10.3923/rjet.2011.1.26

URL: https://scialert.net/abstract/?doi=rjet.2011.1.26
 
Received: July 23, 2010; Accepted: August 02, 2010; Published: November 12, 2010

INTRODUCTION

The world depends on oil. Vast amount is used, transported, processed and stored around the world. In 2003, the total world consumption of petroleum was over 13.1 billion liters per day. The United States Energy Information Administration projects (as of 2006) world consumption of oil to increase to 98.3 million barrels per day (15.63x106 m3 day-1) in 2015 and 118 million barrels day-1 (18.8x106 m3 day-1) in 2030 (EIA, 2006). With such a high consumption, oil spills are inevitable.

The most notable oil spills at sea involve large tankers, such as Exxon Valdez, which spilled thousands of tones oil (Paine et al., 1996; Albaiges et al., 2006). These oil spills can cause severe damage to sea and shoreline organisms (Whitfield, 2003). Most responsible for the contamination are service stations, garages, scrap yards, waste treatment plants, sawmills and wood impregnation plants. Thereafter, several studies have been examined the fate of petroleum in various ecosystem (Boehm et al., 1995; Whittaker et al., 1999).

The development of petroleum industry into new frontiers, the apparent inevitable spillages that occur during routine operations and records of acute accidents during transportation has called for more studies into oil pollution problems (Timmis et al., 1998), which has been recognized as the most significant contamination problem (Snape et al., 2001).

Oil is a complex mixture of hydrocarbons and other organic compounds, including some organometallic constituents (Butler and Mason, 1997). It contain hundreds or thousands of aliphatic, branched and aromatic hydrocarbons (Prince, 1993; Wang et al., 1998), most of which are toxic to living organisms (ATSDR, 1995).

This engine oil renders the environment unsightly and constitutes a potential threat to humans, animals and vegetation (ATSDR, 1997; Edewor et al., 2004). Fat soluble components may accumulate in the organs of animals and may be enriched in the food chain, even up to humans (Mackay and Fraser, 2000). Prolonged exposure and high oil concentration may cause the development of liver or kidney disease, possible damage to the bone marrow and an increased risk of cancer (Propset et al., 1999; Lloyd and Cackette, 2001; Mishra et al., 2001). To animals, even a slight staining by oil can be fatal (Wells, 2001). Oil stains can decrease the insulation of a bird's feathers causing the bird to freeze to death. In addition, oil diminishes the ability of birds to fly, dive and swim, which leads to starvation. Birds also swallow oil when cleaning their feathers, which causes intoxication (EPA, 1999). In the long term, toxic and carcinogenic compounds can cause intoxication, diseases, cell damage, developmental disorders and reproduction problems (ATSDR, 1995). In addition to toxic effects, oil products can affect plant and animals physically. A thick layer of oil inhibits the metabolism of plants and suffocates them. Destruction of plants affects the whole food web and decreases the natural habitats of numerous species (HELCOM, 2003).

Hydrocarbon compounds such as petroleum are essential for life. Since, they do not naturally occur in the forms most useful to humans, they can be hazardous. Fuel and lubricating oil spills have become a major environmental hazard to-date. The contamination of the environment with petroleum hydrocarbons provides serious problems for many countries. Man has dealt with the clean up of petroleum products contamination since the first day oil was discovered.

REMEDIATION OF HYDROCARBON-CONTAMINATED ENVIRONMENT

Remediation refers to removing, degrading or transforming contaminants to harmless or less harmful substances. Additionally, it include methods that reduce mobility and migration of the contaminants, preventing their spreading to uncontaminated areas; toxicity of the contaminants remains unaltered, but the risk they pose to the environment is reduced (US. DOD, 1994). For the treatment of contaminated soil physiochemical and biological treatment is used. The physicochemical treatments are incineration, thermal desorption, coker, cement kiln, solvent extraction and land filling etc.

Incineration is a very effective treatment method, but it is costly and after burning, the soil has lost most of its nutritional value and structure. Land filling does remove the contaminants but only relocates the problem (Lageman et al., 2005). Furthermore, in Europe, legislation requires reduction of the number of landfills. In 2004, the number of landfills untreated contaminated material in England and Wales was reduced from over 200 to only eleven (Environment Agency, 2006). In Finland, the number of landfills has decreased from 232 in 1999 to 80 in 2005 (FEI, 2006). As a result, the cost of dumping contaminated soils into landfills has risen considerably. It is therefore evident that new, innovative methods are needed to treat contaminated soils.

Claude E. ZoBell had as far back as 1946, recognized that many microorganisms have the ability to utilize hydrocarbons as the sole source of carbon and energy. He further recognized that the microbial utilization of hydrocarbons were highly dependent on the chemical nature of the components in the petroleum mixture and environmental determinants (Atlas, 1981). Microbial biodegradation of pollutants has intensified in recent years as mankind strives to find sustainable ways to cleanup contaminated environments (Diaz, 2008). Biodegradation of hydrocarbons by natural populations of microorganisms represent one of the primary mechanism by which petroleum and other hydrocarbon pollutants are eliminated from the environment. The effects of environmental parameters on microbial degradation of hydrocarbons, the elucidation of metabolic pathways, genetic basis for hydrocarbon dissimilation by microorganisms and the effects of hydrocarbon contamination on microbial communities have been the areas of intense interest and the subjects of several reviews (Atlas, 1981, 1994).

Microbial remediation of a hydrocarbon-contaminated site is accomplished with the help of a diverse group of microorganisms, particularly the indigenous bacteria present in soil. These microorganisms can degrade a wide range of target constituents present in oily sludge (Eriksson et al., 1999; Barathi and Vasudevan, 2001; Mishra et al., 2001). A large number of Pseudomonas strains capable of degrading PAHs have been isolated from soil and aquifers (Kiyohara et al., 1992: Johnson et al., 1996; Pathak et al., 2008). Other petroleum hydrocarbon degraders include Yokenella sp., Alcaligenes sp., Roseomonas sp., Stenotrophomonas sp., Acinetobacter sp., Flavobacter sp., Corynebacterium sp., Streptococcus sp., Providencia sp., Sphingobacterium sp., Capnocy-tophaga sp., Moraxella sp., Bacillus sp. and Enterobacter sp., (Rusansky et al., 1987; Antai, 1990; Bhattacharya et al., 2002; Jain et al., 2010). Other organisms such as fungi are also capable of degrading the hydrocarbons of engine oil to a certain extent. However, they take longer periods of time to grow as compared to their bacterial counterparts (Prenafeta-Boldu et al., 2001).

Hydrocarbon degrading bacteria and fungi are widely distributed in marine, freshwater and soil habitats. Similarly, hydrocarbon degrading cyanobacteria have been reported (Lliros et al., 2003; Chaillan et al., 2004), although, contrasting reports indicated that growth of mats built by cyanobacteria in the Saudi coast led to preservation of oil residues (Barth, 2003). Typical bacterial groups already known for their capacity to degrade hydrocarbons include Pseudomonas sp., Marinobacter sp., Alcanivorax sp., Microbulbifer sp., Sphingomonas sp., Micrococcus sp., Cellulomonas sp., Dietzia sp. and Gordonia sp. (Brito et al., 2006). Molds belonging to the genera Aspergillus sp., Penicillium sp., Fusarium sp., Amorphoteca sp., Neosartorya sp., Paecilomyces sp., Talaromyces sp., Graphium sp. and the yeasts Candida sp., Yarrowia sp. and Pichia sp. have been implicated in hydrocarbon degradation (Chaillan et al., 2004).

Petroleum oil degradation by bacteria can occur under both aerobic and anaerobic conditions (Zengler et al., 1999). In general aerobic metabolism of hydrocarbons requires oxygenase enzymes which incorporate molecular oxygen into the reduced substrate. Typically with aliphatic hydrocarbons alcohols are initially produced; these are oxidized sequentially via dehydrogenases to carboxylic acids, which then undergo β-oxidation. In the case of aromatic substrates as well as Polycyclic Aromatic Hydrocarbons (PAHs) hydroxylation of a ring occurs via mono- or dioxygenase enzymes in eukaryotes and prokaryotes (Cerniglia, 1984). After diol formation the ring is cleaved then further degraded.

Bacteria have developed two general strategies for enhancing contact with water-insoluble hydrocarbons: specific adhesion mechanisms and production of extracellular emulsifying agents. Many hydrocarbon-degrading microorganisms produce extracellular emulsifying agents. In some cases, emulsifier production is induced by growth on hydrocarbons (Hisatsuka et al., 1971).

Several reviews have been published on the microbial metabolism of and aromatic hydrocarbons (Cerniglia, 1984; Perry, 1984). It has been established that the first step in the aerobic degradation of hydrocarbons by bacteria is usually the introduction of molecular oxygen into the hydrocarbon. Prokaryotes convert aromatic hydrocarbons by an initial dioxygenase attack to trans-dihydrodiols that are further oxidised to dihydroxy products example, catechol in the case of benzene (Atlas and Bartha, 1998) (Fig. 1).


Fig. 1: Aerobic degradation of the BTEX compounds

Eukaryotic microorganisms use mono-oxygenases, producing benzene 1, 2-oxide from benzene, followed by the addition of water, yielding dihydroxydihydrobenzene (cis-dihydrodiol). This is oxidised in turn to catechol, a key intermediate in biodegradation of aromatics, which is then opened by ortho or meta cleavage, yielding muconic acid or hydroxymuconic semialdehyde, respectively.

In the case of alkanes, mono-oxygenase attack results in the production of alcohol. Most microorganisms attack alkanes terminally whereas some perform sub-terminal oxidation. The alcohol product is oxidized finally into an aldehyde and finally to a fatty acid. The latter is degraded further by beta-oxidation. Different microorganisms exhibit different group specificities. For example, some grow on alkanes of six to ten carbons in chain length, whereas, others grow on long chain alkanes. Some of the oxygenases are encoded on plasmids and others on chromosomal genes. Subterminal oxidation apparently occurs in some bacterial species (Markovetz and Kallio, 1971).

Anaerobic degradation: In the subsurface, oil biodegradation occurs primarily under anoxic conditions, mediated by sulfate reducing bacteria (Holba et al., 1996) or other anaerobes using a variety of other electron acceptors as the oxidant. A number of reports (Ward and Brock, 1978; Berry et al., 1987; Grabic-Galic and Vogel, 1987; Aeckersberg et al., 1991) have demonstrated that toluene, benzene and a variety of alkanes can be degraded under the strictest anaerobic conditions by sulfidogenic and methanogenic cultures. Hydrocarbon biodegradation under anaerobic, denitrifying conditions also follows an oxidative strategy. In the presence of nitrate hydrocarbon substrates e.g., toluene are metabolized to oxidized intermediates prior to further biodegradation (Evans et al., 1992; Flybjerg et al., 1993). Hutchins et al. (1992) and Barbara et al. (1992) reported that the substituted aromatics toluene ethylbenzene and xylenes were biologically removed from the soil under denitrifying conditions. Anaerobic degradation of petroleum hydrocarbons in natural environments by microorganisms has been shown in some other studies to occur only at negligible rates and its ecological significance has been generally considered to be minor (Atlas, 1981; Cooney, 1984). However, the microbial degradation of oxidized aromatic compounds such as benzoate and halogenated aromatic compounds such as the halobenzoates, chlorophenols and polychlorinated biphenyls (Chen et al., 1988) has been shown to occur under anaerobic conditions.

Complete genomes were also determined for bacteria capable of anaerobic degradation of halogenated hydrocarbons by halorespiration. The ~1.4 Mb genomes of Dehalococcoides ethenogenes strain 195, Dehalococcoides strain CBDB1 and ~5.7 Mb genome of Desulfitobacterium hafniense strain Y51 have been sequenced. Characteristic of all these bacteria is the presence of multiple paralogous genes for reductive dehalogenases, implicating a wider dehalogenating spectrum of the organisms than previously known. Moreover, genome sequences provide unprecedented insights into the evolution of reductive dehalogenation and differing strategies for niche adaptation (Heider and Rabus, 2008).

PHYSICAL, CHEMICAL AND ENVIRONMENTAL FACTORS AFFECTING THE BIODEGRADATION OF HYDROCARBONS

Successful application of bioremediation technology to contaminated systems requires knowledge of the characteristics of the site and the parameters that affect the microbial biodegradation of pollutants (Sabate et al., 2004). The overall degradation rate of hydrocarbons biodegradation in soils are strictly limited by a variety of parameters (Rockne et al., 2002). It is therefore necessary to understand the factors limiting microbial degradation in order to adopt appropriate methodology to optimize the process of degradation. Documented reports on the various conditions and factors, which determine the rate of degradation of PAHs (McGill, 1980; Atlas, 1981; Sandrik et al., 1986; Block et al., 1993; Oleszczuk and Baran, 2003).

Temperature: Temperature plays very important role in biodegradation of petroleum hydrocarbons, firstly by its direct effect on the chemistry of the pollutants and secondly on its effect on the physiology and diversity of the microorganisms. Ambient temperature of an environment affects both the property of spilled oil and the activity of microorganisms (Venosa and Zhu, 2003). At low temperatures, the viscosity of the oil increases, while the volatility of toxic low-molecular weight hydrocarbons is reduced, delaying the onset of biodegradation (Atlas, 1981).

Temperature affects the solubility of hydrocarbons (Foght et al., 1996). Although, hydrocarbon biodegradation can occur over a wide ranges of temperature. The rate of biodegradation generally decreases with decreasing temperature. Highest degradation rates generally occur in the range of 30-40°C in soil environments, 20-30°C in some fresh water environments and 15-20°C in marine environments (Bossert and Bartha, 1984; Cooney, 1984). The biodegradation of hydrocarbons in psychrophilic environments have been also reported (Yumoto et al., 2002; Delille et al., 2004; Pelletier et al., 2004).

Cold-adapted microorganisms are able to grow and multiply even at 0°C and below. Their minimum, optimum and maximum temperature for growth are respectively 0-5, >15 and >20°C for psychrotolerants and <0, <15 and <20°C, respectively for psychrophiles. Among psychrotolerants, many have been proven to be hydrocarbon degraders, example, alkane degraders (Whyte et al., 1998; Bej et al., 2000; Ruberto et al., 2005), aromatic degraders (Aislabie et al., 2000; Baraniecki et al., 2002; Grishchenkov et al., 2003; Margesin et al., 2005) and chlorophenol degraders (Jarvinen et al., 1994; Mannisto et al., 2001; Tiirola et al., 2002). Cold-adapted microorganisms are widely distributed in nature with Gram negative bacteria are dominant (Margesin and Schinner, 1999). Cold-adapted strategies include the molecular adaptation of membrane lipid composition, enzyme activity and protein synthesis (Gounot and Russell, 1999; Margesin and Schinner, 1999).

Hydrocarbon degraders are ubiquitous in most ecosystems. They comprise less than 0.1% of the microbial community in unpolluted environments but can constitute up to 100% of the cultivable microorganisms in hydrocarbon-polluted ecosystems (Atlas, 1981). The characteristics of many hydrocarbon degrading bacteria are being examined by the scientists for their possible use in bioremediation. Deeb and Alvarez-Cohen (1999) studied the temperature effects on a consortium of toluene-degrading bacteria. They found that their consortium grew best at 35°C. Richmond et al. (2001) studied the temperature effects on bacteria from benzene-contaminated aquifer, Alaska using glutamate as a carbon source, they found that the overall microbial metabolic rates were higher at 25°C than 10°C.

Nutrients: The nutrient status of soil has direct impacts on microbial activity and biodegradation. To grow heterotrophic bacteria require in addition to an organic compound that serves as a source of carbon, electron donor and a group of other nutrient elements. Many bacteria and fungi also require low concentrations of one or more amino acids and vitamins. Nitrogen and phosphorus are necessary for cellular metabolism and can be found in low concentrations in many soils, including Arctic soils (Pritchard and Charles, 1991; Braddock et al., 1997; Mohn and Stewart, 2000).

McMillen et al. (1995) reported that the C/N ratio approximately 2 to 200 is optimum for fertilizer applications. In principle, the problem can be solved by using nitrogen and phosphorus compounds with low C:N and C:P ratios. It was found that a combination of paraffinized urea and octyl phosphate was able to replace nitrate and inorganic phosphate, respectively (Atlas and Bartha, 1998). Urea and diammonium phosphate were used as fertilizers in both the field and microcosm experiments described in project at a C:N:P ratio of 100:3.25:0.75 as it was recommended by past experiments dealing with CFS Alert hydrocarbon contaminated soil (Allen, 1999).

Biodegradability is inherently influenced by the composition of the oil pollutant. For example, kerosene (consists of almost exclusively medium chain alkanes) is totally biodegradable. Similarly, crude oil is also biodegradable quantitatively, but heavy asphaltic-naphthenic crude oil, only about 11 % biodegradable within a reasonable time period, even if the conditions are favourable. Okoh (2002) reported that between 8.8 to 29% of the heavy crude oil Maya was biodegraded in soil microcosm by mixed bacterial consortium in 15 days, although major peak components of the oil was reduced by between 6.5 to 70% (Okoh, 2003). Also, about 89% of the same crude oil was biodegraded by axenic culture of Burkholderia cepacia RQ1 in shake flask (Okoh et al., 2001) within similar time period.

The petroleum biodegradation has been reported to be mostly enhanced in presence of a consortium of bacteria species compared to monospecies activities (Ghazali et al., 2004). Rahman et al. (2003) reported that the percentage of degradation by the mixed bacterial consortium decreased from 78 to 52% as the concentration of crude oil was increased from 1 to 10%. There has been a reported case of lack of correlation between degradation rates, specific growth rates and concentration of the starter oil (Thousand et al., 1999), in such a case, it would appear that biomass was required only to a particular threshold enough to produce the appropriate enzyme system that carry through the degradation process even when biomass production had ceased (Okoh, 2002), a phenomenon completely at systems (Pitter and Chudoba, 1990), where production of variance with the theory of microbial growth in batch cells is totally dependent on the consumed carbon source.

Various reports on the effect of sunlight irradiation so far published have focused on the physico-chemical changes on intact crude oil other than to biodegraded crude oil (Jacquot et al., 1996; Nicodem et al., 1998). Maki et al. (2005) reported that photo-oxidation increases the biodegradability of petroleum hydrocarbons by increasing its bioavailability and thus, enhancing microbial activities. In a related study, Trindade et al. (2005) assessed the bioremediation efficiency of a weathered and freshly contaminated soil. The additions of nutrients are necessary to enhance the biodegradation of oil pollutants (Choi et al., 2002; Kim et al., 2005). Pelletier et al. (2004) assessed the effectiveness of fertilizers for crude oil bioremediation in sub-Antarctic intertidal sediments over a one year and observed that chemical, microbial and toxicological parameters demonstrated the effectiveness of various fertilizers in a pristine environment. In study using poultry manure as organic fertilizer in contaminated soil increased biodegradation was reported but the extent of biodegradation was influenced by the incorporation of alternate carbon substrates or surfactants (Okolo et al., 2005).

Chaillan et al. (2006) reported that excessive nutrient concentration can inhibit the biodegradation activity and several authors have also reported the negative effect of a high NPK levels on the hydrocarbons biodegradation (Oudot et al., 1998; Chayneau et al., 2005) and more especially on the aromatics (Carmichael and Pfaender, 1997).

Effect of chemical composition of petroleum hydrocarbons: Petroleum hydrocarbons can be divided into four classes: saturates, aromatics, asphaltenes (phenols, fatty acids, ketones, esters and porphyrins) and resins (pyridines, quinolines, carbazoles, sulfoxides and arnides). Hydrocarbons differ in their susceptibility to microbial attack and ranked in the following order of decreasing susceptibility: n-alkanes > branched alkanes> low molecular weight aromatics> cyclic alkanes (Perry, 1984).

Solubility: Biodegradability of hydrocarbons in soil has been demonstrated to correlate to their water solubility because bacteria in the unsaturated soil occur mainly in the interstitial water of soil. Therefore, solubility of the chemical will determine its concentration in soil.

Bioavailability: Bioavailability is the amount of a substance that is physiochemically accessible to microorganisms. It is a key factor in the efficient biodegradation of pollutants. Chemotaxis or the directed movement of motile organisms towards or away from chemicals in the environment is an important physiological response that may contribute to effective catabolism of molecules in the environment. In addition, mechanisms for the intracellular accumulation of aromatic molecules via various transport mechanisms are also important (Parales, 2008). Volkering et al. (1993) demonstrated that growth on crystalline substrates (naphthalene) results in linear growth rates indicating that partitioning, example, solubilization of the substrate is rate limiting to biodegradation. Britton (1984) reported that uptake of hydrocarbons most likely occurs by attachment then incorporation into the cytoplasmic membrane. Alternately transport occurs by passive or facilitated diffusion in the presence of solubilizing agents; intracellular transport is probably coordinated with enzymatic oxidation.

Introduction of external nonionic surfactants, e.g., the main components of oil spill dispersants, influence the alkane degradation rate (Bruheim and Eimhjelle, 1998; Rahman et al., 2003). Experience so far indicates that the use of surfactants in situations of oil contamination may have a stimulatory, inhibitory or neutral effect on the bacterial degradation of the oil components (Liu et al., 1995). The need to accurately characterize the role of chemical and biological surfactants has been proposed in order that performance in biological systems may be predicted (Rocha and Infante, 1997; Lindstrom and Braddock, 2002). However, in contrast to chemical dispersants, which caused ecological damage after application for abatement of spilled oil in marine ecosystems (Smith, 1968), biosurfactants from soil or freshwater microorganisms are less toxic and partially biodegradable (Poremba et al., 1991). Commercially available surfactants both ionic and nonionic in nature (Laha and Luthy, 1992; Thai and Maier, 1992; Pennell et al., 1993) as well as biosurfactants and biosurfactant-producing bacteria have been investigated for their ability to increase bioavailability (Van Dyke et al., 1991; Volkering et al., 1993; Miller, 1994). Other methods for increasing bioavailability may also enhance the biodegradation of contaminants in a soil. For example, physical disruption of soil aggregates using sonication has been reported to increase biodegradation rates effectively in a landfarm experiment. Weissenfels et al. (1992) demonstrated that soil constituents have significant impact on the bioavailability of contaminants.

Manilal and Alexander (1991) reported that mineralization rate of contaminants are lower in soils with a high organic matter content, which readily absorbs hydrophobic compounds. Soluble humic substances in particular humic and fulvic acids appear to be major binding sites. Their binding potential can be attenuated by mineral soil components, as well as pH and salt concentrations (Schlautman and Morgan, 1993). Weathering or the age of contamination may also affect bioavailability by physically trapping, hindering and/or slowing desorption of contaminants from the soil (Connaughton et al., 1993).

Physiochemical properties of the soil: Soils vary widely with regard to geology, hydrology, climate, fertility and other physical attributes. The most important physical and chemical property of the soil is determined by the composition of soil, organic matter and fraction of soil (Owabor and Ogunbor, 2007).

Oxygen: Oxygen is another important parameter because it determines the bacterial pattern of dissimilatory and energy yielding process. Microbial utilization of aliphatic (Singer and Finnerty, 1984), cyclic (Perry, 1984) and aromatic (Cerniglia, 1997) hydrocarbons by bacteria and fungi required electron sink. In the initial attack molecular oxygen used as electron sink. In the subsequent steps too, oxygen is the most common electron sink. In the absence of molecular oxygen, further biodegradation of partially oxygenated intermediates may be supported by nitrate or sulphate reduction. Little or no hydrocarbon metabolism occurs in strictly anoxic sediments (Dibble and Bartha, 1976).

Soil moisture: Soil moisture is another important parameter in determining the rate of biodegradation of petroleum compounds. Microbes live in the interstitial water of soil pores and the lower amount of water, the smaller the number of microbes and thus, slow removal rate through biodegradation (Dibble and Bartha, 1976).

Acidity or alkalinity: The acidity (pH) of the soil is an important soil parameter. Soil pH can be highly variable, ranging from 2.5 in mine spoils to 11 in alkaline deserts. Most heterotrophic bacteria favour a pH 7.0 but fungi being more tolerant to acidic conditions. Therefore, extremes pH of soils would be have a negative influence on the ability of microbial populations to degrade hydrocarbons. Verstraete et al. (1975) reported that a doubling rate of biodegradation of gasoline in an acidic (pH 4.5) soil by adjusting the pH to 7.4. Rates dropped significantly, however, when the pH was further raised to 8.5. Similarly, Dibble and Bartha (1976) observed an optimal pH of 7.8, in the range 5.0 to 7.8 for the mineralization of oily sludge in soil.

Calculation of degradation rates of PAHs by bacteria in laboratory is often useful. In a batch laboratory culture, four phases of growth are found in bacteria: lag, log, stationary and death phase. Optical density tests are performed using spectrophotometer that determine the percentage of light absorbed or transmitted by the sample. When the optical density increases linearly, the bacteria are in log phase. Graphs of cell numbers versus time can also be used in a similar fashion to determine log phase (Tortora et al., 1995).

Prior to the degradation of many organic compounds, a period is noted in which no destruction of the chemical is evident. This time interval is designated an acclimation period or sometimes as adaptation or lag period. The duration of the acclimation period may vary enormously. It may be less than one hour or can take many months (Alexander, 1999). The duration varies among chemicals and environments. It is also depends on the concentration of the compound and a number of environmental conditions. A large number of reports reviewed by Colwell et al. (1978), Atlas (1981) and Cooney (1984) have shown that the numbers of hydrocarbon utilizing microorganisms and their proportion in the heterotrophic community increase upon exposure to petroleum or other hydrocarbon pollutants. The levels of hydrocarbon-utilizing microorganisms generally reflect the degree of contamination of the ecosystem.

Plasmid DNA plays an important role in genetic adaptation as it represents a highly mobile form of DNA which can be transferred via conjugation or transformation and can impart novel phenotypes including hydrocarbon-oxidizing ability to recipient organisms. The pathways for the metabolism of naphthalene, salicylate, camphor, octane, xylene and toluene have been shown to be encoded on plasmids in Pseudomonas sp. (Chakrabarty, 1976). The emerging trend is that the large gene repertoires of potent pollutant degraders such as LB400 and RHA1 have evolved principally through more ancient processes. That this is true in such phylogenetically diverse species is remarkable and further suggests the ancient origin of this catabolic capacity (McLeod and Eltis, 2008).

The approach to inoculation must be calculated and prudent. If there is an indigenous micro flora capable of carrying out the degradative reaction conditions that favour its multiplication and rapid destruction of the pollutant are not essential and additions of inocula are not needed. Such environments contain bacteria able to grow on and destroy a variety of hydrocarbons. Persistence of the components of oil is not a consequence of the absence of organisms but rather the absence of full set of conditions necessary for the indigenous species to function rapidly (Atlas, 1981, 1994; Margesin and Schinner, 1999).

Petroleum hydrocarbons are typically a complex mixture of aliphatic and aromatic organic compounds. They can be fractionated by distillation into saturates, aromatics, asphaltenes and resins (Solomon, 1996). The saturates include n-alkanes, branched alkanes and cycloalkanes. Polycyclic Aromatic Hydrocarbons (PAHs) are organic molecules with two or more benzene rings in which the number and arrangement of the rings result in diverse physical and chemical properties (Leahy and Colwell, 1990).

PAHs are formed during combustion reactions. Contribution of the petroleum hydrocarbons to the environment results from burning of fossil fuels and subsequent atmospheric deposition, primarily from industrialized countries (Sims and Overcash, 1983). Internal combustion engines produce large amounts of petroleum hydrocarbon byproducts. In addition, many industrial activities associated with processing, production and disposal of petroleum hydrocarbons contribute to the overall environmental load. Some of these activities include liquefaction, heat and power generation using fossil fuels, catalytic cracking procedures, coal tar production, refining, distillation of crude oil, wood-treatment processes, storage, open burning of tires and accidents resulting from handling etc. (Bumpus et al., 1985; Wilson and Jones, 1993).

Soil contaminated with petroleum hydrocarbons are remediated using a diverse set of physicochemical and biological (biodegradation and bioremediation) treatment. In physico-chemical treatment incineration, thermal desorbtion, coker, cement kiln and solvent extraction are used but they have some disadvantages.

Bioremediation processes have been shown to be an effective method that stimulates the biodegradation in contaminated soils (McLaughlin, 2001; Swanell et al., 1996). Biodegradation of hydrocarbon-contaminated soils (exploits the ability of microorganisms to degrade and/or detoxify organic contamination) has been established as an efficient, economic, versatile and environmentally sound treatment. In 1946, Claude E. ZoBell reviewed the action of microorganisms on hydrocarbons. He recognized that many microorganisms have the ability to utilize hydrocarbons as sole sources of energy and carbon and that such microorganism are widely distributed in nature. Bioremediation involves the use of indigenous or introduced microorganisms to degrade environmental contaminants (Margesin and Schinner, 1997). Leahy and Colwell (1990) reported the biodegradation of petroleum oil by Achromobacter, Arthrobacter, Acinetobacter, Alcaligenes, Bacillus, Flavobacterium, Nocardia, Pseudomonas and Rhodococcus. Biodegradation of oil by fungi Rhodotorula, Sporobolomyces, Aspergillus and Penicillium has been also studied (Head and Swannell, 1999). Similarly bioremediation of malathion pesticide for pollution control reported by Adhikari (2010).

Metabolic studies were implemented on the aerobic pathways for alkane, cycloalkane and Polycyclic Aromatic Hydrocarbon (PAH) biodegradation (Cerniglia, 1992; Kampfer et al., 1993; Juhasz and Naidu, 2000) for transformations of nitrogen and sulfur compounds (Kaiser et al., 1996; Bressler et al., 1998; Beller and Spormann, 1999; Bressler and Fedorak, 2000) and the microbial mechanisms of anaerobic hydrocarbon catabolism (Harwood and Gibson, 1997; Heider et al., 1998; Lovely, 2000).

In situ remediation: In this remediation, the soil is not excavated but treated as it is. It usually leads to considerable savings because excavation costs and costs of transport to treatment facilities. According to the calculations of the federal remediation technologies roundtable (Federal Remediation Technologies Roundtable, 2000) the costs of biological in -situ treatments were approximately 8-80 m-3.

Selection of the best treatment method depends on the type and quantity of the contaminants, treatment costs, the soil type and the environmental conditions on the site, among other things (US. DOD, 1994; Khan et al., 2004). Before selection of the best available method or methods, it is necessary to conduct a thorough investigation of the properties of the soil. This is especially true for in-situ remediation, because performance of the remediation process is more difficult to monitor and control during treatment than traditional ex-situ remediation processes (Morgan and Watkinson, 1992).

In situ remediation is a good option on many occasions. Excavating the soil may be too expensive or even impossible, if the area to be treated is very large or the contaminated zone lies deep in soil. It is costly to excavate the whole mass and transport it for remediation. If the contaminated site has buildings or is otherwise in active use, excavation is not an option (Carberry and Wik, 2001), however, challenges are exist with in-situ treatment. A major challenge with all in-situ treatment methods is to achieve uniform remediation throughout the treatment area. As contaminated soils are usually heterogeneous, a uniform delivery of chemicals to the whole area is very difficult to achieve (Federal Remediation Technologies Roundtable, 2000).

In situ biological treatment technologies include bioventing, land treatment, phytoremediation and natural attenuation.

Bioventing: It is also called as soil vaccum extraction. It is used for removal of oily phase contaminants above the water-table. A well is bored near the point of contamination but above the water-table. Vaccum is applied from this extraction well and volatile emissions are safely vented. Oxygenated air comes in contact with the undissolved contaminated subsurface material which gets biodegraded. Appearance of CO2 in extraction well is indication of biodegradation activity. Bioventing is cheaper than use of nitrates or hydrogen peroxide as the source of electron acceptors.

Land treatment: It is a full scale bioremediation technology in which contaminated soils, sediments or sludges are periodically turned over (tilled) and allowed to interact with the soil and climate at the site. One advantage of land treatment is that only conventional farming equipments are required.

Biosparging: It involves the injection of air under pressure below the water-table to increase groundwater oxygen concentrations and enhance the rate of biological degradation of contaminants by naturally occurring bacteria. Biosparging increases the mixing in the saturated zone and thereby increases the contact between soil and groundwater. The ease and low cost of installing small-diameter air injection points allows considerable flexibility in the design and construction of the system.

Electroremediation: Electrokinetic remediation has been used to remove metals and polar organic compounds from soils, sludges and sediments (Page and Page, 2002; Virkutyte et al., 2002). Electrokinetic remediation methods use electrodes with a low-level direct current electric field (usually <10 V cm-1 or mA cm-2) installed into the contaminated soil (Acar and Alshawabkeh, 1993; Pamucku, 1997; Van-Cauwenberghe, 1997). The current mobilizes and transports charged chemicals in the soils liquid phase towards the electrodes. Negatively charged anions and organic compounds will move to the anode, whereas positively charged chemicals, such as metals, will move towards the cathode (Van-Cauwenberghe, 1997).

Contaminants can be removed from the electrodes with several methods including electroplating, adsorption onto the electrodes, pumping near the electrodes, precipitation or co-precipitation at the electrodes, complexing and capturing the contaminants in reactive permeable barriers (Acar and Alshawabkeh, 1993; Pamucku, 1997; Van-Cauwenberghe, 1997; Virkutyte et al., 2002).

The main phenomena affecting the movement of contaminants in electrokinetic remediation are electro-migration and electro-osmosis. In electro-migration, the chemicals move towards the electrodes according to their charges, whereas in electro-osmosis, they are transported by water to the cathode. Water can also transport uncharged organic and inorganic compounds. Electro-migration is independent of the pore size of the soil, therefore, applicable to both coarse- and fine-grained soils. However, electro-osmosis is ineffective in coarse-grained soils (Acar and Alshawabkeh, 1993; Probstein and Hicks, 1993; Pamucku, 1997; Van-Cauwenberghe, 1997; Virkutyte et al., 2002).

Metals are usually present in soil as cations, therefore migrating to the cathode (Page and Page, 2002). This migration is further enhanced by the electro-osmotic flow. A number of studies have presented the feasibility of removing metals from soil with electro-kinetics. The most common contaminant metals studied are lead (Reed et al., 1995; Acar and Alshawabkeh, 1996; Kim et al., 2001), cadmium (Reddy and Chinthamreddy, 1999; Kim et al., 2001; Vengris et al., 2001; O’Connor et al., 2003), copper (Maini et al., 2000a, b) and chromium (Reddy and Chinthamreddy, 1999; Pamukcu et al., 2004).

Although, electroremediation of organics is not as widespread as treating metals, positive results have encouraged its implementation (Acar et al., 1992; Shapiro and Probstein, 1993; Ribeiro et al., 2005). The contaminants studied include for example phenol (Acar et al., 1992; Shapiro and Probstein, 1993; Yang and Long, 1999), acetic acid (Shapiro and Probstein, 1993), atrazine (Ribeiro et al., 2005), PAHs (Maini et al., 2000b) and 2, 4-dichlorephenoxyacetic acid (Jackman et al., 2001).

Phytoremediation: Phytoremediation is an emerging technology that uses plants to remove contaminants from soil and water (USEPA, 1999, 2000).

We found following types of phytoremediation technique, classified based on the contaminant fate: phytoextraction, phytotransformation, phytostabilization, phytodegradation, rhizofiltration, even if, a combination of these can be found in nature.

The sorbents used presently, can be classified as polymers, natural materials or treated cellulosic materials (Deschamps et al., 2003). Several studies of various natural, synthetic and mineral sorbents exist (Reynold et al., 2001; Teas et al., 2001; Deschamps et al., 2003; Toyoda and Inagaki, 2003). A number of natural sorbents have been studied for use in oil spill cleanup: e.g., cotton wool, bark, milkweed, kapok, kenaf, rice straw, even dried biomass of an aquaphyte. Most of them have higher absorption capacities than synthetic ones but they often absorb water well, which is a disadvantage when used in marine environments (Wei et al., 2003). Occasionally their hydrophobicity can be enhanced by various treatments; however, that increases the overall costs of the sorbents.

Specific challenges of in situ bioremediation: Bioremediation techniques have been tested in laboratories with good results but the transfer to the field has been often quite difficult (Sturman et al., 1995; Allard and Neilson, 1997). In laboratories, the tests are often performed under optimal conditions, which do not correlate well with conditions in the field (Sturman et al., 1995; Mandelbaum et al., 1997; Goltz et al., 2001; Davis et al., 2003). Application of data from laboratory experiments to the field has often lead to unsatisfactory results, example, the half-lives of compounds in the field tend to be 4-10 times longer than in the laboratory (Schirmer et al., 2000).

Ex situ: This treatment generally requires shorter time period than in situ treatment and provides more uniformity of treatment because of the ability to homogenizing, screening and monitoring. However, ex situ treatment requires excavation, leading to increase cost and engineering for equipment. Available ex situ biological treatment technologies include composting, landfarming, slurry phase biological treatments, bioreactors and biopiles etc.

Composting: It is a controlled biological process by which organic contaminants are converted by microorganisms to innocuous and stabilized by products. Typically, thermophilic conditions (54 to 65°C) must be maintained to properly compost soil contaminated with hazardous organic contaminants. The increased temperatures result from heat produced by microorganisms during the degradation of organic material.

Landfarming: It is a full-scale bioremediation technology that usually incorporates liners and other methods to control leaching of contaminants and requires excavation and placement of contaminated soils. Contaminated media are applied into lined beds and periodically turned over or tilled to aerate. Soil conditions are often controlled to optimize the rate of contaminant degradation. Conditions normally controlled include the following:

Moisture content usually by irrigation or spraying

Aeration by tilling the soil at a predetermined frequency

pH (buffered near neutral pH by adding crushed limestone or agricultural lime)

Other amendments (soil bulking agents, nutrients etc.)

Contaminated media are usually treated in lifts that are up to 0.46 m thick. When the desired level of treatment is achieved, the lift is removed and a new lift is constructed. It may be desirable to remove only the top of the remediated lift and then construct the new lift by adding more contaminated media to the remaining material and mixing. This serves to inoculate the freshly added material with an actively degrading microbial culture, which can reduce treatment times. Landfarming is a medium to long term technology.

Bioreactor: It is a vessel, ranging from small (1-5 m3) portable units to large plants built specifically on polluted site. When contaminated soil is inaccessible (in prime location beneath the building or when buried pipes, cables prevent or when groundwater is to be treated), this method is desirable.

Biopiles: Biopiles are hybrid of landfarming and composting and used for treatment of surface contamination with petroleum hydrocarbons. These are refined version of landfarming that tend to control physical losses of the contaminants by leaching and volatilization. Biopiles provide a favorable environment for indigenous aerobic and anaerobic microorganisms.

The analysis of microbial communities that are involved in hydrocarbon degradation are still a challenge to microbiologists and microbial ecologists. It is estimated that ≈90-99% of the species making up heterotrophic communities do not form colonies when current laboratory techniques are applied (Roszak and Colwell, 1987).

Our understanding of bacterial biogeography and community assembly is correspondingly vague, anecdotal and controversial (Curtis et al., 2002). Traditional approaches to the study of microbial diversity have relied on laboratory cultivation of isolates from natural environments and identification by classical techniques, including analysis of morphology, physiological characteristics and biochemical properties. These approaches provide information on fine scale but suffer from bias by media and cultivation conditions (Prosser, 2002). Estimated genotypic diversity in bacterial communities based on DNA renaturation experiments suggests that there are 4x103 to 7x103 different genome equivalents per g of soil (Torsvik et al.,1990), which if extrapolated to species diversity, suggests that there are perhaps 103 or even more species per g of soil.


Table 1: Ribosomal RNAs in prokaryotes
*The name is based on the rates that are molecule sediments. Woese et al. (1990)

Molecular phylogeny is a portion of a gene. Analysis of the nucleotide base sequence in certain genes of bacterial DNA makes it possible to deduce broader phylogenetic relationship among bacteria. These genes are conserved through the billions of years of evolutionary divergence and have changed more slowly during evolution. The bulk of genome and by allowing them to be used as molecular/evolutionary chronometers and these define the ribosomal RNAs (rRNAs). Ribosomal RNA proposed as one of the best molecule and has been used by Woese et al. (1990) for studies on bacterial evolution. Major properties of rRNA are (1) rRNA are old molecules present in the ribosomes, (2) they are functionally constant, (3) have a wide distribution, (4) are well conserved over large phylogenetic distances and (5) they occur in large number of cells (1000-100000/cells).

Most bacteria have 5S, 16S and 23S rRNA with different chain length and sedimentation rate in several copies with in each cell (Table 1).

The 5S has been extensively studied, but its size is usually too small for reliable phytogenetic inference and specificity. It is used to distinguish major phytogenetic groups. The 16S and 23S rRNA is sufficiently large to be quite useful and reliable for phytogenetic identification. The 23S rRNA is excellent for phylogenetic studies, but so far few studies are available. 16S rRNA has been given most attention. These have been widely used for this purpose and useful pressure in molecular biology.

According to variability map, over 10% of bases in the 16S rRNA gene are totally conserved (within a sample of 500 bacterial sequences). However, majority of the conserved bases are not adjacent to each other and thus, form no continuous conserved regions for universal priming. The longest string of totally conserved bases is between positions 788 and 798, but in most areas of the gene absolutely conserved bases are found in strings of less than 4. Thus, no primer of sufficient length can be designed that is a 100% match to all bacterial, let alone, all prokaryotic 16S rRNA gene sequences. Furthermore, sequences of recently discovered taxa are not adequately represented in the variability map (Watanabe et al., 2001).

Out of many molecular methods that are currently available, Polymerase Chain Reaction (PCR) based techniques have become most powerful (Innis et al., 1990; McPherson and Taylor, 1992). This approach was first time applied by Giovannoni et al. (1990). The PCR is a technique to amplify specific DNA sequences with the help of a DNA polymerase enzyme and specific primers. Over the past 25 years, a large number of primer sequences for amplification and sequencing of rRNA genes has been published (Watanabe et al., 2001; Kolganova et al., 2002). Some of these primers have been designed as taxa specific, whilst others have been designed to amplify all prokaryotic rRNA genes and are referred to as universal (Hugenhoitz et al., 1998; Nielsen et al., 1999). But several authors have emphasized how notoriously incomplete most PCR libraries are for describing microbial diversity on the basis of only a few hundred 16S rRNA amplicons at best (Acinas et al., 2004; Kemp and Aller, 2004; Hong et al., 2006).

The ability to identify and quantify specific microbial species within environmental settings would be valuable to numerous fields of research. Changes in microbial community structure induced by natural or anthropogenic factors could be monitored. However, the techniques available for identifying and quantifying specific microbial species have been of limited usefulness for in-situ real world settings due to the necessity of prior culturing of bacteria, inability to quantify microorganisms or lack of sensitivity (Welsh and McClelland, 1990). Archaea have several mismatches in their 16S rRNA sequences to the commonly used broad-specificity primers and were therefore not present in the 16S rRNA sequences to the commonly used broad-specificity primers and not present in the 16S rRNA libraries (Baker et al., 2006).

To perform a correct assessment, it is necessary to consider various microorganisms having a variety of genomes and expressed transcripts and proteins. However, several high-throughput techniques originally developed for medical studies can be applied to assess biotreatment in confined environments (Watanabe and Kasai, 2008). Our increasing capabilities in adapting the catalysts to specific reactions, process requirements by rational and random mutagenesis broad the scope for application in the fine chemical industry, but also in the field of biodegradation. In many cases, these catalysts need to be exploited in whole cell bioconversions or in fermentations, calling for system-wide approaches to understanding strain physiology, metabolism and rational approaches to the engineering of whole cells as they are increasingly put forward in the area of systems biotechnology and synthetic biology (Meyer and Panke, 2008).

CONCLUSION

The quality of life on earth is linked inextricably to the overall quality of the environment. For centuries, we believe that atmospheric, terrestrial and aquatic systems were sufficient to absorb and breakdown wastage from population centre, industry and farming. However, today the resources in the world show greater or lesser degree due to our carelessness and negligence in using them. The problems associated with petroleum contaminated sites assume an increasing prominence in many countries. These pollution problems often result in huge disturbances of both the biotic and abiotic components of the ecosystem. The currently accepted disposal methods are incineration or burial in secure landfills can become prohibitively expensive when the amounts of contaminants are large. Microbial remediation of an oil contaminated site is accomplished with the help of a diverse group of microorganisms, particularly the indigenous bacteria present in soil. This process is known as bioremediation. It uses relatively low-cost, low technique, which generally have a high public acceptance and can often be carried out on site.

ACKNOWLEDGMENT

Authors are grateful to Prof. Shakti Baijal, Faculty of Arts, Science and Commerce, MITS University for providing technical assistance and critical review.

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