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Contaminated land is widely recognised as a potential threat to human health.The release of petroleum hydrocarbon through improper handling, storage and waste management is a major issue in environmental remediation. Despite regulatory steps that have been implemented to reduce, remove or eliminate production and release of these chemicals into the environment, significant environmental contamination has occurred in the past and will probably continue to occur in the future. Several studies have shown that microbial degradation of crude oil is an important factor contributing to the elimination of oil spills from the environment. Based on frequency of isolation, hydrocarbon degrading bacteria and fungi genera in soils and aquatic environment are mainly the Bacillus, Micrococcus, Achromobacter, Nocardia, Pseudomonas, Trichoderma, Aspergillus, Mortierella and Penicillum species In the bioremediation of organically
polluted sites, microbial extracellular enzyme activities are important and key steps, since only compounds with molecular mass lower than 600 daltons can pass through cell pores. The three common extracellular enzymes are lipase, amylase and protease. In this study, petroleum hydrocarbon degrading bacteria were isolated from soil collected from the Mechanic Village in Nsukka. They were screened for lipase producing potential and the isolate with the best potential was selected for the study.
Through morphological and biochemical tests, it was identified as Pseudomonas aeruginosa. Lipase was produced through submerge fermentation in a stationary phase. Microbial utilisation of petroleum hydrocarbon was monitored using the
microorganism and the lipase. Results showed that the lipase was most active in the pH range 6.0 – 8.0 with an optimal reaction temperature of 40oC. Statistical analysis revealed that there was significant difference (P<0.05) between the controls and the test




Contamination of soil and groundwater through the release of petroleum hydrocarbons has been recognized as one of the principal issues in environmental remediation (Shah and Bhatt, 2006). Improper handling, storage and waste management practices have resulted in contamination of soil and aquifers; constituting threat to drinking water

supplies, adversely affecting the people and many other areas of the environment (Fetter, 1993; Shah and Bhatt, 2006). In many countries, the problems associated with contaminated sites are assuming an increasing prominence, and as such have become a worldwide problem (Vidali, 2001). In the oil producing areas of Nigeria, crude oil exploration has resulted in serious problems of oil contamination of the environment such as reduction in farming and fishing activities (Oluyege and Oluyemi, 2005). In the United States, it is estimated that most underground water tables are leaking (Kovalick, 1991).

Despite regulatory steps that have been implemented to reduce, remove or eliminate production and release of these chemicals into the environment, significant environmental contamination has occurred in the past and will possibly continue to occur in the future (Baker and Diana, 1994). The release of these pollutants in some cases such as industrial emissions are deliberate and well regulated; while in other cases such as chemical spills, they are accidental and largely unavoidable (Akpofure et al., 2000). Also, the activities of individuals at various mechanic workshops further contaminate the environment with petroleum products such as diesel, engine oil and petrol (De and Bello, 2002).


Contaminated land is widely recognised as a potential threat to human health (Vidali, 2001). Although the components of petroleum derived products can be biodegraded easily in contrast to man-made compounds, they are also dangerous (Wyszkowska et al, 2006). They display potential carcinogenic and mutagenic activities (Krahl et al.,

2002). This recognition has resulted in international efforts to remedy contaminated sites, either as a response to the risk of adverse health or environmental effects caused by contamination or to enable such sites to be developed for use (Vidali, 2001).

Several studies have shown that microbial degradation of crude oil is an important factor contributing to the elimination of oil spills from the environment (Okpokwasili and Amanchukwu, 1988; Fuentes et al., 1998; De and Bello, 2002). Based on frequency of isolation, hydrocarbon degrading bacteria and fungi genera in soils and aquatic environment are mainly the Bacillus, Micrococcus, Achromobacter, Nocardia,

Pseudomonas, Trichoderma, Aspergillus, Mortierella and Penicillum species (Atlas, 1981;De and Bello, 2002). In the bioremediation of organically polluted sites, microbial extracellular enzyme activities are important (Munster and De Haan,

1998), and key steps since only compounds with molecular mass lower than 600 daltons can pass through cell pores (Hoppe, 1991). The three groups of common extracellular enzymes are lipase, amylase and protease (Shah and Bhatt, 2006).


Lipases are glycerol ester hydrolases that catalyse the hydrolysis of ester linkages of glycerides at water-oil interface (Kashmiri et al., 2006).

Even though this enzyme occurs widely in nature, only the microbial lipases are commercially significant (Sharma et al., 2001). They are amongst the most important biocatalysts carrying out novel reactions in both aqueous and non-aqueous media. This is as a result of their ability to utilise a wide range of substrates, high stability towards extremes of temperature, pH and organic solvents (Saxena et al., 1999).

They catalyse the hydrolysis of triacylglycerols to glycerol and free fatty acids, and are activated only when adsorbed to an oil-water interface (Martinelle et al., 1995). However, the amount of oil available at the interface determines the activity of the lipase. This interface area can be increased substantially to its saturation limit by the use of emulsifiers as well as by agitation (Saxena et al., 1999). Lipases are important in a large number of industrial applications such as organic chemical processing, detergent formulations, synthesis of biosurfactants, the oleochemical industries, the dairy industries and the agrochemical industries (Table 1).

They can also be used to accelerate the degradation of fatty waste (Masse et al., 2001).

Table 1: Industrial Application of Microbial Lipases (Vulfson, 1994)

Industry Action Product Or Application Detergents Hydrolysis of fats Removal of oil stains from fabrics Dairy foods Hydrolysis of milk fat, cheese ripening, modification of butter Development of flavoring agents in milk, cheese, and Bakery foods Flavor improvement Shelf-life prolongation Beverages Improved aroma Beverages Food dressings Quality improvement Mayonnaise, dressings, and Whippings Health foods Transesterification Health foods Meat and fish Flavor development Meat and fish products; fat Removal Fats and oils

Transesterification; hydrolysis cocoa butter, margarine, fatty acids, glycerol, mono-,and diglycerides Chemicals Enantioselectivity, synthesis Chiral building blocks, Chemicals Pharmaceuticals Transesterification, hydrolysis Specialty lipids, digestive Cosmetics Synthesis Emulsifiers, moisturizers Leather Hydrolysis Leather products Paper Hydrolysis Paper with improved quality Cleaning Hydrolysis Removal of fats The development of lipase-based technologies for the synthesis of novel compounds is rapidly expanding the uses of these enzymes (Liese et al., 2000).

The most widely used enzymes in organic synthesis are lipases and more than 20% biotransformations are performed with lipases (Gitlesen et al., 1997). Currently, lipases are therefore, the enzymes of choice for biochemists, organic chemists, pharmacists, biophysicists, biochemical and process engineers, biotechnologists and microbiologists (Saxena et al., 1999).

So many lipase purification schemes have been described in literature, focusing on purifying small amounts of the enzyme to homogeneity for the purpose of characterisation. Such schemes include ethanol precipitation, ammonium sulphate precipitation, chromatographic techniques. (Mase et al., 1995; Takahashi et al., 1998). Highly purified lipases are usually not required for commercial applications due to the expensive nature of excessive purification, which also reduces overall recovery of the enzyme (Chisti, 1998).


Micro-organisms that produce lipase have been found in diverse habitats such as industrial wastes, vegetable oil processing factories, dairies, soils contaminated with oil, decaying oil seeds, and decaying foods (Sztajer et al., 1988); compost heaps, coal tips and hot springs (Wang et al., 1995). These micro organisms include bacteria, fungi, yeasts and actinomyces (Sharma et al., 2001).

Bacterial lipases are glycoproteins, though some extracellular bacterial lipases are lipoproteins (Saxena et al., 1999). Among bacteria, the following: Achromobacter sp., Alcaligenes sp., Arthrobacter sp. Pseudomonas sp. Staphylococcus sp. and Chromobacterium sp have been exploited for lipase production (Saxena et al., 1999). The chief fungal producers of lipase include Aspergillus niger, Candida cylindracea, Mucor miehei and Rhizopus arrhizus (Saxena et al., 1999).


Submerged culture is mainly used for the production of microbial lipases (Ito et al., 2001), though solid-state fermentation methods can also be used (Chisti, 1999). The majority of lipases are secreted extracellularly. This is because lipids are insoluble in water and to function as nutrients in the cell, they need to be broken down

extracellularly into their more polar components to facilitate absorption.

Generally, lipids stimulate the production of lipase. Lipase production is usually coordinated with, and dependant on the availability of triacylglycerol. However, free fatty acids, hydrolysable esters, bile salts and glycerol can also stimulate lipase production

(Saxena et al., 1999).

The type and concentration of carbon and nitrogen sources, the culture pH, the growth temperature, and the dissolved oxygen concentration influence microbial lipase production (Elibol and Ozer, 2001). Several carbohydrate and lipid sources have been used as carbon sources for the production of lipase. These include fructose, glucose, olive oil, soybean oil, sunflower oil, sesame oil, cotton seed oil, corn oil, peanut oil and palm oil (Sharma et al., 2001; Papaparaskevas et al., 1992).

Also, different nitrogen sources are used for the production of microbial lipase (Kashmiri et al., 2006; Sharma et al., 2001). Such nitrogen sources are peptone, corn steep liquor, soybean meal, yeast extract and ammonium phosphate. Generally, organic nitrogen sources are said to provide high yield of microbial lipase (Sharma et al., 2001)

The effect of aeration on lipase production by different microorganisms varies. While vigorous aeration greatly reduced lipase production by Pseudomonas fragi, high aeration was needed for high lipase activity by A.wentii (Saxena et al., 1999).


Bioremediation has been defined by several authors in different ways. According to Mueller et al., 1992, bioremediation is the process whereby organic wastes are biologically degraded under controlled conditions to an innocuous state, or to levels below concentration limits established by regulatory authorities. Vidali (2001) defined bioremediation as the use of living organisms, primarily microorganisms, to degrade environmental contaminants into less toxic forms. As an environmental remediation option, bioremediation uses relatively lowcosts, low- technology techniques and can often be carried out on site (Vidali, 2001).

Although bioremediation does not require complex technical methodologies, considerable experience and expertise are needed to design and implement a successful bioremediation programme, as it is important to thoroughly assess a site for suitability and to optimize conditions necessary for achieving a satisfactory result (Vidali, 2001). The following are some basic steps that may be necessary for any bioremediation project: Compliance analysis, site characterisation, method selection (feasibility studies), remediation proper and end of project analysis (Bonaventura et al., 1995).

There is a rapid increase in research in this field, because bioremediation presents a good alternative to conventional clean-up technologies. Bioremediation has been used at a number of sites worldwide, including Europe, with varying degrees of success. With increasing knowledge and experience, there are improvements on bioremediation techniques (Vidali, 2001).

Bioremediation uses naturally occurring bacteria and fungi or plants to degrade or detoxify substances that are hazardous to human health and/or the environment. Such bacteria and fungi may be indigenous to the contaminated site or may be isolated from other places and brought to the contaminated area (Vidali, 2001)


Many factors influence the control and optimisation of bioremediation processes. These factors include the following:

  1. Microbial population: Microorganisms can be isolated from diverse environmental conditions. However, the existence of a microbial population capable of degrading the pollutants is considered the key component of bioremediation (Okoh and Trejo-Hernandez, 2006; Vidali, 2001). Due to their adaptability, microbes can be used to degrade or remediate environmental hazards (Vidali, 2001).

For these microbes to effectively carry out biodegradation, they must be present in appropriate densities with the capability to degrade the target compound(s) (Okoh and Trejo -Hernandez, 2006).

  1. Availability of contaminants to microorganisms: It is of extreme importance that microorganisms and the contaminants should be in contact (Vidali, 2001). Such contaminants must be accessible to the organisms in a form that they can utilise energy and carbon source (Okoh and Trejo-Hernandez, 2006).
  2. Type of contaminant: The structure and metabolic abilities of a microbial community may be affected by different types of waste (Okoh and Trejo -Hernandez, 2006). It is therefore very’ important to consider the type of contaminant in a bioremediation study.

Also, for proper evaluation of the feasibility of a bioremediation technology, adequate information about site characterization is important (Okoh and Trejo -Hernandez, 2006).

  1. Environmental factors: Favourable environmental conditions that support the growth and enzymatic activity of the microorganisms that will degrade contaminants are very important. Such environmental conditions should limit the growth of competitive organisms in favour of those conducting the desired reactions (Vidali, 2001; Graham et al., 1999). These conditions include temperature, pH, moisture, oxygen, nutrients and soils structure.

Temperature has a great effect on biochemical reaction rates; above a given temperature, the cells die. While water is essential for all the living organisms, the amount of available oxygen will determine whether the system is aerobic or anaerobic (Vidali, 2001). Nutrients are the basic building blocks of life and allow microbes to produce the enzymes necessary for the breaking down of contaminants (Vidali, 2001). The effective delivery of air, water and nutrients is controlled by the soil structure.

It has also been reported that soil fauna redistributes microbes and or help re-introduce them from less contaminated soil layers, and as such they are useful in bioremediation (Okoh and Trejo – Hernandez, 2006).

Worms of various sizes for example, mix the soil and make it more porous thereby improving aeration which is necessary for effective bioremediation (Romantschuk et al., 2001).

Some environmental conditions affecting degradation include soil pH, oxygen and nutrient contents, temperature, heavy metals, soil type as shown in Table 2.

Table 2: Environmental Conditions Affecting Degradation

Parameters Condition required for microbial activity

Optimum Value for an oil Degradation

Soil moisture 25-28% of water holding capacity ‘ 30-90%

Soil pH 5.5-8.8 6.5-8.0

Oxygen content Aerobic, minimum air-filled pore space of 10%10-40%

Nutrient content

N and P for microbial growth C:N:P=100:10:1 Temperature(°C)

15-45 20-30

Contaminants Not too toxic

Hydrocarbon 5-10% of dry weight of soil

Heavy metals Total content 2000 ppm 700ppm

Type of soil Low clay or silt content Source: Vidali . (2001)


Different bioremediation strategies and procedures are currently being used. The various techniques are either in stitu or ex situ. In situ techniques are those bioremediation activities that are carried out at the site with minimal disturbance, while ex situ techniques take place at a location different from the site of contamination (Vidali, 2001).

The following strategies are used for bioremediation activities (vide

Table 3 for their summary)

  1. Bioaugmentation: This is the addition of microorganisms indigenous or exogenous to a contaminated site in order to supplement the indigenous microbial population and speed up degradation (Smith, 2004; Vidali, 2001). It involves the removal of microbial samples from a polluted site, enrich the useful microbes, scale-up the population and re-inoculate large quantities of the microbes into the contaminated site (Smith, 2004).
  2. Biosimulation: This is the addition of essential growth nutrients such as nitrogen and phosphorous to a contaminated site to stimulate microbial growth. This will lead to an increase in the breakdown of pollutants (Smith, 2004).

iii. Bioventing: This is an in situ treatment that involves supplying of air and nutrients through wells to contaminated soil to stimulate the indigenous microorganisms. Only the amount of oxygen necessary for the biodegradation is provided through low air flow 

rates. There is minimal volatilisation and release of contaminants to the atmosphere (Vidali, 2001).

  1. Biosparging: This is the process of injecting air under pressure below the water table in order to increase groundwater oxygen concentrations, thereby enhancing the rate of biodegradation of contaminants by naturally occurring microorganisms (Vidali,


  1. Landfarming: This involves the excavation of contaminated soil and spreading it over a prepared bed. It is then periodically tilled until the pollutants are degraded. It is aimed at stimulating indigenous

biodegradative microorganisms and to facilitate their aerobic degradation of contaminants (Vidali, 2001).

  1. Composting: This is the mixing of contaminated soil with nonhazardous organic amendants such as manure or agricultural wastes. The development of a rich microbial population is supported by the organic materials (Vidali, 2001).

vii. Biopiles: Biopiles are a hybrid of landfarming and composting. They

are basically used for the treatment of surface contamination with petroleum hydrocarbons (Vidali, 2001).

viii. Bioreactors: A slurry bioreactor is a containment vessel used to create a three-phase (solid, liquid and gas) mixing condition to increase the bioremediation rate of soil-bound and water-soluble pollutants. It allows for aeration, adequate mixing and control of many of the factors affecting biodegradation (Okoh and Trejo -Hernandez, 2006; Vidali, 2001).

Table 3: Summary of Bioremediation Strategies

Technology Examples Benefits Limitation Factors to ConsiderIn situ In situ

Bioremediation Biosparaging

Bioventing Bioaugmentation Most cost efficient. Noninvasive.Relatively passive. Natural Attenuation process. Treats soil and water. Environmental constraints. Extended treatment time. Monitoring difficulties. Biodegradative abilities of indigenous microorganisms. Presence of metals and other inorganics. Environmental parameters. Biodegradability of pollutants. Chemical solubility. Geological Factors. Distribution of pollutants.

Ex situ Landfarming Composting Biopiles Cost efficient Low cost Can be done on site Space requirements. Extended treatment time. Need to control abiotic loss. Mass transfer problem. Bioavailabilty limitation. See above Bioreactors Slurry reactors Aqueous reactors Rapid degradation kinetic. Optimized environmental parameters. Enhances mass transfer. Effective use of inoculants and surfactants. Soil requires excavation. Relatively high cost capital. Relatively high operating cost See above Bioaugmentation. Toxicity of amendments. Toxic concentration of contaminants.

Source: Vidali (2001)


The growth and metabolism of microorganisms in natural environments are always limited by the availability of electron donors, electron acceptors, or other essential nutrients (Atlas, 1981). In petroleum spills however, there is no lack of electron donors (petroleum hydrocarbons), microbial metabolism therefore, is generally limited by the availability of electron acceptors (such as oxygen, nitrate, sulphate,

Fe (III) or by the availability of essential nutrients


such as nitrogen, potassium and phosphate (Chapelle, 1999).

Although bioremediation can be used to treat oil-contamination, formulating treatment strategies that will produce a specified outcome in terms of degradation rate and residual contaminant concentration is an important limitation of the technology (Head, 1998). This is partly due to the empirical development of bioremediation. The amount of nutrient (principally N and P) applied to spilled oil for example, may be based on

consideration of the amount of N and P required to convert a given amount of hydrocarbon to carbon dioxide, water and microbial biomass under oxic conditions, or from the concentration of nutrients shown to support maximal growth rates of alkane –degraders in culture (Head and Swannell, 1999). Also, it has been suggested that the amount of slow release fertiliser applied to a beach should be as much as possible without exceeding toxic concentrations of ammonia and /or nitrate (Pritchard et al., 1992); or that nutrient addition should be sufficient to cause a detectable increase in the N and P content of water, thus ensuring that N and P are not limiting the microbial population (Swannell et al., 1996). These approaches however, almost inevitably result in the addition of more nutrients than is strictly necessary.

Due to the complexity of bioremediation systems, and limited knowledge of how indigenous microbial populations respond to environmental perturbations; most treatment strategies are ad hoc formulations and rely on data from laboratory or microcosm biodegradation feasibility studies (Head and Swannell, 1999). Therefore, a theoretical basis to explain the behaviour of microorganisms in the environment in response to specific stimuli has not been well developed (Head and Swannell, 1999).

However, resource-ratio theory that has been applied to hydrocarbon biodegradation (Smith et al., 1998) relates the structure and function of biological communities to competition for growth – limiting resources. When the quantitative requirements for limiting resources (i.e concentration of a limiting resource that supports zero net growth) and the growth and death rate of different competing organisms are known, resource –ratio theory offers the possibility to predict the outcome of such interactions (Smith et al., 1998).

Using the context of resource –ratio theory, it may be possible to devise bioremediation treatments objectively, by imposing conditions that select for the microorganisms most fit to remove the contaminants of greatest environmental concern, at the optimal rate with minimum intervention. Different organisms for example, have different requirements for N and P. The provision of these nutrients at different concentrations will select for the organisms most able to utilise the nutrients at the levels provided in the contaminated habitat (Head and Swannell, 1999).

A consideration of resource-ratio theory in bioremediation could thus provide a more sophisticated way of manipulating the indigenous microbiota to bring about rapid removal of toxic organic contaminants from many types of polluted habitat (Head and Swannell, 1999).

Furthermore, identification of the nutrient limitation status of the indigenous microbial population is an important factor in determining if nutrients must be added to stimulate biodegradation (Head and Swannell, 1999). Using wet chemicals or instrumental methods to measure nutrient concentrations may not necessarily reflect the bioavailability of nutrients; however, it is possible to identify nutrient limitation by the analysis of genes expressed in response to nutrient starvation (Head and Swannell, 1999).

For example, specific proteins expressed in response to phosphate limitation have been identified in Pseudomonas fluorescens, Cyanobacteria and Thiobacillus ferrooxidans; and immunological approaches have been used to detect their expression in individual cells (Varela et al., 1998). A possible means for monitoring changes in nutrient limitation status in response to bioremediation treatments is the detection of such proteins or their mRNA in natural populations of bacteria at the site of oil spills (Fleming et al., 1993).


The process of quantifying bioremediation is not an easy one.

Bioremediation is often monitored by following target contaminant concentrations, reductions of which are not always indicative of decreased soil toxicity (Philips et al., 2000). There could be incomplete degradation resulting in the formation of toxic intermediary constituents (Heitzer and Sayler, 1993). Philips et al., 2000 recommends a battery of chemical analysis for target contaminant levels, and toxicity testing for measuring soil toxicity. Such chemical analysis and test include:

(1) The use of bioassays to monitor changes in soil toxicity during bioremediation to complement chemical analysis of contaminants, particularly where complex mixtures of soil contaminants are present and possible biodegradation products have not been characterized (Knoke et al., 1999)

(2) Monitoring biological responses of aquatic protozoan in order to test effluent toxicity. An indicator organism of choice due to its rapid behavioural and physiological response to possible hazardous substances is Daphnia similis (Burton, 1998).

(3) Microbial respiratory activity: The use of respirometry for the measurement of CO2 production rates arising from microbial activity in hydrocarbon contaminated soils is highly recommended (Zucchi et al., 2003). Since CO2 production is proportional to microbial activity, measuring its release can provide accurate data and proof that biodegradation is occurring (Zucchi et al., 2003).

(4) Resting-cell assay which is a common technique performed routinely for the quantification of polycyclic aromatic hydrocarbons degradation (Stringfellow and Aitken, 1995)

(5) Chemical analysis of contaminant compounds using gas chromatography-mass spectrometer (GC-MS) and flame ionization detector (FID) (Bach et al., 2005)




With recent developments and applications of state of the art molecular technique, the process of hydrocarbon catabolisms has substantially advanced. Many novel catalytic mechanisms have been understood and characterised such as the cellular and other physiological adaptations of microbes to the presence of hydrocarbons, as well as the biochemical mechanisms involved in hydrocarbon accession and uptake (van Hamme et al., 2003) Chakrabarty et al. (1973) combined the degradative capabilities of two different strains of Pseudomonas into a single organism using the techniques of genetic engineering.

One of the strains, Pseudomonas aeruginosa, designated strain AC 59, was capable of degrading C6 , C8, and C10 compounds but could not degrade longer aliphatic compounds . The other strain designated strain AC 63, could degrade C12 and C14 molecules but could not degrade shorter hydrocarbons .

The new organism was capable of degrading C6 through C18 hydrocarbons. Furthermore, the marriage of 


modern recombinant DNA technology and the petroleum industries may produce new strains capable of not only broad hydrocarbon metabolism, but also adaptability to contaminated environments (van Hamme et al., 2003).


Extracellular enzyme activity is a key step in degradation and utilisation of organic polymers. This is because only compounds with molecular mass lower than 600 Daltons can pass through cell pores (Hoppe, 1991; Meyer Reil, 1991). For effective biodegradation, the presence of an inducer that will cause the synthesis of specific enzymes for the target compound(s) is an important requirement (Okoh and Trejo- Hernandez, 2006).

The rate of petroleum hydrocarbon biodegradation for example, is affected by any factor which influences the rate of microbial growth and enzymatic activities (Atlas, 1981). Microorganisms are the key components in bioremediation. However, it is the enzymes they produce that are involved in the degradative reactions that eventually lead to the elimination or detoxification of the chemical pollutants (Okoh and Trejo-Hernandez, 2006).

After screening extracelluar enzymes from isolates of hydrocarbon contaminated soil, Shah and Bhatt (2006) found that lipase production had good correlation with hydrocarbon degradation. The generalised scheme of catabolic pathways for aromatic compounds, has revealed that the initial conversion steps are carried out by different enzymes leading to intermediates of central metabolic routes, such as the tricarboxylic acid cycle (Okoh, 2006; Chaudry and Chapalamadugu, 1991).

This suggests that microorganisms have extended their substrate range by developing peripheral enzymes, which are able to transform initial substrates into one of the central intermediates (Van der Meer et al., 1992).


  1. The advantages of bioremediation according to Vidali (2001) include:

• Many compounds that are largely considered to be hazardous can be transformed to harmless products.

• Complete destruction of target pollutant is possible.

• Less expensive than other technologies that are used for the cleanup of hazardous waste.

• Can be carried out on site thereby eliminating the need to transport quantities of waste off site with its potential threats to human health and environment.

  1. The disadvantages of bioremediation according to Vidali (2001)


• Limited to compounds that are biodegradable.

• There are some concerns that the products of biodegradation may be more persistent or toxic than the parent compound.

• Extrapolation from bench and pilot-scale studies to full – scale field operations is difficult.

• It often takes longer than other treatment options

• There are no acceptable endpoints for bioremediation treatments.


Petroleum was first obtained in pre-Christian times by the Chinese and has been known to occur in the surface seepage for several years (Okoh, 2006). It is an extremely 

complex mixture of hydrocarbons, and several classes based on related structures can be recognised from the hundreds of individual components (Atlas, 1981). The hydrocarbons in crude petroleum are classified on a structural basis as alkanes (normal and iso), cycloalkanes and aromatics. The unsaturated analogues of alkanes (alkenes) are rarely found in crude petroleum but occur in many refined petroleum products as a result of the cracking process (Bartha, 1986). In addition to the large number of hydrocarbons that occur in crude petroleum, small amounts of oxygen- (phenols, naphthenic acids), nitrogen – (pyridine. pyrrole, indole) and sulphur – (alkylthiol, thiophene) containing compounds, collectively known as resins, and a partially oxygenated, highly condensed asphaltic fraction are found; but they are not present in refined petroleum (Bartha, 1986). On the basis of their respective distillation residues, crude oils could be classified as paraffins, naphthenes or aromatics and based on the relative proportions of the heavy molecular weight constituents as light, medium or heavy (Okoh, 2006). Fig 1 shows structural classification of some crude oil components.

Fig 1. Structural Classification of Some Crude Oil Components (Okoh, 2006)

The biodegradability of these individual components is a function of their chemical structure. However, the physical state and toxicity of the compounds also has a strong influence on biodegradability (Okoh, 2006; Bartha, 1986). The n-alkanes as a structural group for example, are generally considered the most readily degraded components in a petroleum mixture, however, the C5 –C10 homologues have been shown to be inhibitory to the majority of hydrocarbon degraders (Okoh, 2006; Bartha, 1986). Being solvents, they tend to disrupt lipid membrane structures of microorganisms (Okoh 2006). Waxes i.e. alkanes in the C20- C40 range are hydrophobic solids at physiological temperatures and this physical state is said to interfere strongly with their biodegradation (Bartha and Atlas, 1977).

The biodegradation pathway of intact hydrocarbons proceeds by the action of oxygenases and therefore, requires the presence of free oxygen. In case of alkanes, it is a monoterminal attack by monooxygenase resulting in the formation of a primary alcohol (Bartha, 1986; Atlas, 1981). The alcohol product is then oxidised to an aldehyde and a monocarboxylic acid (Atlas, 1981). The carboxylic acid is further degraded by beta -oxidation to produce a two – carbon -unit shorter fatty acids and acetyl coenzyme A, with eventual liberation of C02 (Atlas,1981). Extensive methyl branching interferes with the beta-oxidation process and necessitates diterminal attack or other bypass mechanisms with the formation of a secondary alcohol and subsequent ketone (Bartha, 1986; Atlas, 1981).

Petroleum hydrocarbons occur in complex mixtures in crude petroleum as well as in refined products, and they influence each other’s biodegradation. The effects could be negative or positive (Bartha, 1986). Some iso-alkanes for example, are spared as long as n-alkanes are available as substrates, while some condensed aromatics are metabolized only in the presence of more easily utilisable petroleum hydrocarbons, a process referred to as co-metabolism (Wackett, 1996).

1.4.1 DIESEL

Diesel is a mixture of hydrocarbons obtained from the fractional distillation of crude oil between 250°C and 350°C at atmospheric pressure. It is composed of about 75% saturated hydrocarbons and 25% unsaturated hydrocarbons. Diesel oil contains primary hydrocarbons formed by 9-23 carbon atoms per molecule. The percentage composition of aromatic hydrocarbons in diesel is about 45% (Smith, 1990).

The biological balance of soil contaminated with diesel oil is usually distorted, altering the succession of microorganism directly associated with the activity of soil enzymes (Bundy et al., 2002; Wyszkowska et al., 2002; Kaplan and Kitts, 2004). Under this condition, there is an increase in the counts of microorganisms and enzyme activity without leading to any improvement in soil fertility (Wyszkowska and Kucharski, 2005; Delille et al., 2004). Due to the destructive influence of diesel oil on the soil structure as well as soil air, water and chemical properties, there is a negative response of plants to diesel oil contamination (Caravaca and Rodan, 2003).

Wyszkowska et al., 2006 observed that diesel oil contamination had adverse influence on the growth and development of oat maize.


The term total petroleum hydrocarbon refers to a large family of several hundreds of chemical compounds that originally come from crude oil (ATSDR, 2008). Due to the presence of so many different chemicals in crude oil and in other petroleum products, it is not practical to measure each one separately; and as such, the total petroleum hydrocarbon is measured (ATSDR, 2008). Thus total petroleum hydrocarbon can be referred to as the measurable amount of petroleum-based hydrocarbons in the environment.

Total petroleum hydrocarbon is divided into groups of petroleum hydrocarbons that act alike in soil or water called petroleum hydrocarbon fractions. Each fraction contains so many individual chemicals (ATSDR, 2008). Total petroleum hydrocarbon may contain some of these chemicals: hexane, jet fuels, mineral oils, benzene, toluene, xylenes, naphthalene and fluorene. Samples of total petroleum hydrocarbon however, are likely to contain only some or a mixture of these chemicals (Draggan. 2008)

Exposure to total petroleum hydrocarbon (TPH) either through breathing air at petrol stations, using chemicals or certain pesticides at home, drinking water contaminated with TPH, working in places that use petroleum products, living in an area near a spill or leak of petroleum products or touching soil contaminated with TPH could lead to several health hazards. Some of these health hazards include headache, dizziness, nerve disorders especially the peripheral neuropathy consisting of numbness in the feet and legs and cancer. (Draggan, 2008).

Animal studies also have shown that some TPH compounds affect reproduction and the development of the foetus in some animals (Draggan, 2008).


Of great importance in the microbial degradation of petroleum hydrocarbons is the composition of the pollutants. The qualitative hydrocarbon content of the petroleum mixture influences the degradability of individual hydrocarbon components (Atlas, 1981).

Westlake et al., (1978), reported that the ability of mixed microbial populations to utilise the hydrocarbons in four crude oils as the sole carbon source was found to depend not only on the composition of the unsaturated fraction but also on that of the asphaltic fraction. A petroleum hydrocarbon mixture with its large number of potential primary substrates, provides an excellent chemical environment in which co-oxidation can occur 

(Atlas, 1981). Co-oxidation is a phenomenon in which compounds which otherwise would not be degraded can be enzymatically attacked within the petroleum mixture due to the abilities of the individual microorganisms to grow on other hydrocarbons within the oil (Horvath, 1972).

Another important factor in the biodegradation of petroleum hydrocarbons is the physical state of the oil pollutants. The physical state of the petroleum hydrocarbon pollutant determines the initial surface area for the initiation of biodegradation (Bartha, 1986). At very low molecular weight, hydrocarbons are soluble in water, but most oil spill incidents release petroleum hydrocarbons in concentrations far in excess of the solubility limits thereby leading to the spreading of the oil (Boylan and Tripp, 1971). In aquatic environment, the degree of this spreading partly determines the surface area of oil available for microbial colonization by hydrocarbon-degrading microorganisms (Atlas, 1981). At low temperatures, the degree of spreading is reduced due to the viscosity of the oil (Atlas, 1981). The spreading of oil is limited in soils due to absorption of petroleum hydrocarbons by plant matter and soil particles (Atlas, 1981). In their study, Wodzinsky and LaRocca (1977) reported that diphenylmethane which is a liquid at 30°C was degraded, but at 20°C it could not be degraded due to its solid form. Also, they found that naphthalene could not be utilised in the solid form but when dissolved in a liquid hydrocarbon, it could be utilised by microorganisms.

The growth of hydrocarbon utilising microorganisms in water and soils is often limited by inadequate mineral nutrients, especially nitrogen and phosphorus (Bartha, 1986). Some of these nutrients could become limiting depending on the nature of the impacted environment, thus affecting the process of biodegradation (Okoh, 2006). In the event of a major oil spill in marine and fresh water environments, carbon supply is increased dramatically and the availability of nitrogen and phosphorous generally becomes the limiting factor for oil degradation (Atlas, 1984).

However, biodegradation activity could be inhibited by excessive nutrient concentrations. For example, the negative effect of a high NPK levels on biodegradation of hydrocarbons has been reported (Chaineau et al., 2005).

In hydrocarbon polluted coastal areas of the Gulf of Marseille,

France, foams were frequently observed at the sea surface (Goutx et al.,

1987). These foams have the ability to emulsify crude oil (Rambeloarisoa et al., 1984). Emulsifying agents are responsible for this foam formation, and since detergents (synthetic emulsifiers) were not detected, it was believed that the foams are of biological origin (biosurfactants) (Goutx et al., 1987). Rambeloarisoa et al. (1984) reported that hydrocarbons and hydrocarbonoclastic bacteria accumulated extensively in these foams 107 to 108 bacterial ml-1). Hydrocarbon- degrading microorganisms have been reported to produce emulsifying agents (Rosenberg, 1986). The production of surface – active molecules by bacteria could therefore be an important factor involved in the production of foam (Goutx et al, 1987).

These surface active agents include low molecular weight compounds such as fatty acids, triacylglycerols and phospholipids, as well as the heavier glycolipids (Cirigliano and Carman, 1984). Many authors have considered the emulsification of hydrocarbons by surface active agents to be an essential step in hydrocarbon biodegradation, particularly in the marine environment (Bartha, 1986; Floodgate, 1984).

Many hydrocarbon-degrading microorganisms produce cell wall – associated or extracellular surface-active agents, in order to facilitate the uptake of hydrocarbon through the hydrophilic outer membrane (Haferburg et al., 1986). Goutx et al. (1987) reported that there is a good relationship between lipid production and emulsifying activity. They found that increase in lipid production leads to an increase in emulsifying activity One of the primary mechanisms through which petroleum hydrocarbon pollutants are eliminated from the environment, is biodegradation of hydrocarbons by natural populations of microorganisms (Leahy and Colwell, 1990). The major areas of interest in the microbial degradation of hydrocarbons have been the effects of environmental parameters, elucidation of metabolic pathways and genetic bases for hydrocarbon dissimilation by microorganisms, and the effects of hydrocarbon contamination on microorganisms and microbial communities (Leahy and Colwell, 1990).

By its effect on the physical nature and chemical compositions ofoil, rate of hydrocarbon metabolism by microorganisms and composition of the microbial community, temperature influences petroleum hydrocarbon degradation (Atlas, 1988). The onset of biodegradation is delayed at low temperatures due to increases in the viscosity of the oil and reduction in the volatilisation of toxic short-chain alkanes (Atlas and Bartha, 1972). Also, the rate of degradation decreases with temperature as a result of decreased rates of enzymatic activity (Atlas and Bartha, 1972). The rates of hydrocarbon metabolism increases at higher temperatures to a maximum, usually in the range of 30-40oC, above which the membrane toxicity of hydrocarbons is increased (Bossert and Bartha, 1984). Climate and season therefore, would select for different populations of hydrocarbon-utilising microorganisms which are adapted to ambient temperatures (Leahy and Colwell, 1990).

The initial step in the catabolism of hydrocarbons by bacteria and fungi involves the oxidation of the substrate by oxygenases, which requires molecular oxygen (Singer and Finnerty, 1984). Therefore, aerobic conditions are important for microbial oxidation of hydrocarbons in the environment through this route. However, some studies have shown that anaerobic degradation of petroleum hydrocarbons by microorganisms occurs, though at negligible rates; and its ecological significance has been generally considered to be minor (Atlas, 1988; Bailey et al., 1973; Bossert and Bartha, 1984). Also, the microbial degradation of aromatic and halogenated aromatic compounds such as benzoates, halobenzoates, chlorophenols and polychlorinated biphenyls has been shown to occur under anaerobic conditions (Boyd and Shelton, 1984; Suflita et al., 1982). The ability of microbial populations to degrade hydrocarbons is influenced by extremes of pH (Leahy and Colwell, 1990).

Assemblages of mixed populations with overall broad enzymatic capacities, are required to degrade complex mixtures of hydrocarbons such as crude oil in soil and marine environments. This is because individual organisms can metabolise only a limited range of hydrocarbon substrates (Leahy and Colwell, 1990). Some common factors affecting petroleum hydrocarbon biodegradation are listed in Table 4 below.

Table 4: Some Common Factors Affecting Petroleum Hydrocarbon

Biodegradation Limiting factor Explanation or examples

Petroleum hydrocarbon composition (PHC) Structure, amount, toxicity, Physical state Aggregation, spreading, dispersion, Adsorption Weathering Evaporation, photo oxidation Water activity Osmotic and matrix forces, exclusion of water from hydrophobic aggregates Oxidant O2 required to initiate oxidation,NO-3 or SO4 2- to sustain PHC biodegradation. Mineral nutrients N,P, Fe may be limiting

Reaction Low pH maybe limiting Microorganisms PHC degraders may be absent or low in Numbers Temperature Influence on evaporation and degradation rates

Source: Bartha. 1986



Polycycilc aromatic hydrocarbons are fused benzene ring compounds that are structurally complex. Fig 2 below show the chemical structure of some polycyclic aromatic hydrocarbons.

Source: NASA Ames Research Centre, 2005


Pericondensed Catacondensed

Pyrene C16H10 Coronene C24H12 Naphthalen C10H8 Phenanthrene C14H10 Perylene C20H12 Benzo[ghl]perylene C22H12 Tetraphene C18H12 Chrysene C18H12 Antanthrene C22H12 Ovalene C32H14 Pentaphene C22H14 Pentacene C22H14

They are unique environmental contaminants because they are generated continuously by the inadvertently incomplete combustion of organic matter, such as in forest fires, home heating, traffic and waste incineration (Johnsen et al.,2005). Under normal conditions, they are highly recalcitrant due to their strong molecular bonds and they have high affinity for soil material and particulate matter. Many of the consistuents of polycyclic aromatic hydrocarbons are not only carcinogenic and mutagenic, they are also potent immunotoxicants (Bach et al., 2005). Polycyclic aromatic hydrocarbons possess such physical properties as low aqueous solubility and high solid –water distribution ratios, which make them not to be readily available for microbial utilisation and promote their accumulation in the solid phases of the terrestrial environment. Table 5 shows the characteristics of a typical polycyclic aromatic hydrocarbon contaminated soil.

Source: Piskonen and Itavaara, (2004)

Table 5: Characteristics of a Typical PAH Contaminated Soil

Although polycyclic aromatic hydrocarbons with lower molecular weight (two to four ringed compounds) such as naphthalene, acenaphtlyene and fluorene are relatively easy to degrade; the rate of degradation in general is rather slow in the environment, particularly in the aquatic systems (Han et al., 2003).

Microbial bioremediation of polycyclic aromatic hydrocarbons depends largely on the following factors:

(i) Physical characteristics of the polycyclic aromatic hydrocarbon

constituents: As the molecular size increases and there is prolonged exposure to soil particles, bioavailability is greatly reduced and biodegradation rates become slower. (Piskonen and Itavaara, 2004)

(ii) The choice of microbial consortium: Many microbial strains are capable of degarading only specific hydrocarbon compounds. A single bacterial specie has only limited capacity to degrade all the fractions of hydrocarbons present (Loser et al., 1998).

Therefore, a mixture of outside bacterial armies that can degrade a broad range of the hydrocarbon constituents should be employed.

(iii) Factors affecting the biodegradation mechanism: Several physical, chemical and biological factors ultimately determine the effectiveness of strategies of choice for microbial bioremediation. Biosurfactants for example, are important agents in the effective uptake of polycyclic aromatic hydrocarbons by bacteria and fungi (Leahy and Colwell, 1990).

Also, pH and availability of nutrients such as nitrogen and phosphorus are very important. Others include salinity, oxygen, temperature, pressure and moisture content (Leahy and Colwell, 1990; Boyd and Shelton, 1984; van Hamme et al., 2003). Genetic compatibility and readiness are also an important factor determining the success of microbial catabolism of polycyclic aromatic hydrocarbons (van Hamme et al., 2003)

Bacterial degradation of polycyclic aromatic hydrocarbons is initiated by the action of intracellular deoxygenases. The polycyclic aromatic hydrocarbons must therefore be taken up by the cells before degradation can take place (Johnsen et al., 2005). Bacteria most oftenNoxidise polycyclic aromatic hydrocarbons to cis-dihydrodiols by incorporation of both atoms of an oxygen molecule. The cis-hydrodiols are further oxidised first to the aromatic dihydroxy compounds (catechols) and then channeled through the ortho- or meta-cleavage pathways (Smith, 1990). However, anaerobic degradation with nitrate and sulphate as terminal acceptors has been reported (Johnsen et al., 2005).

The biological degradation of polycyclic aromatic hydrocarbons serves the following functions (Johnsen et al., 2005)

(i) Assimilative biodegradation that yields carbon and energy for the degrading organism and goes along with the mineralization of the compound or part of it.

(ii) Intracellular detoxification processes in order to make the polycyclic aromatic hydrocarbon water soluble as a prerequisite for excretion of the compounds.

(iii) The degradation of polycyclic aromatic hydrocarbons without generation of energy and carbon for the cell metabolism, known as co-metabolism.

Despite concerted efforts, only a very limited number of bacteria that can grow in pure cultures of polycyclic aromatic hydrocarbons with five or more aromatic rings (high molecular weight) have been isolated (Johnsen. et al.,2005).


The objectives of this study were:

  1. To isolate microorganism from diesel contaminated soil and use it to produce lipase.
  2. To monitor the effect of the lipase so produced on hydrocarbon utilisation by the microorganism.

iii. To study some of the properties of the enzyme


Considering the fact that microbial degradation of petroleum hydrocarbons is associated with the production of lipids and other surface active agents, it is possible therefore that introducing a lipase could accelerate the process of hydrocarbon biodegradation.

This will contribute to the growing research efforts towards finding ways of containing the environmental problems of petroleum hydrocarbon contamination. Moreso, the lipase produced could be exploited for the production of other value added products.


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