当前位置:首页 >> 自然科学 >>

Bioremediation of oil-contaminated soil


Journal of Microbiological Methods 32 (1998) 155–164

Journal of Microbiological Methods

Bioremediation of oil-contaminated soil: microbiological methods for feasibility assessment and eld evaluation
M.T. Balba*, N. Al-Awadhi, R. Al-Daher
Kuwait Institute for Scientic Research, P.O. Box 24885, 13109 Safat, Kuwait

Abstract Bioremediation is emerging as a promising technology for the treatment of soil and groundwater contamination. The technology is very effective particularly in dealing with petroleum hydrocarbon contamination. However, bioremediation is a site-specic process and feasibility studies are required before full-scale remediation can be successfully applied. The type and scale of the feasibility studies that will be needed are specic to the bioremediation approach to be employed during full-scale clean-up operation. In all cases however, these studies have the same goals: to accurately determine if specic hydrocarbon contaminants are amenable to biological treatment and to determine the time and cost required to treat the contaminants of concern according to the regulated clean-up criteria. This contribution provides background information on the chemistry and microbiology of hydrocarbon contamination, discusses the prospective of using biological methods for addressing this problem and describes several microbiological methods which can be used for the feasibility assessment of soil bioremediation. The focus of this chapter is to highlight the needs for the integration of laboratory data to full-scale bioremediation. 1998 Elsevier Science B.V. Keywords: Bioremediation; Petroleum hydrocarbons; Feasibility assessment

1. Introduction There is growing public concern as a wide variety of toxic organic chemicals are being introduced inadvertently or deliberately into the environment. Petroleum hydrocarbons are one common example of these chemicals, which enter the environment frequently and in large volumes through numerous routes. The seepage from natural deposits is one of the major routes by which petroleum oil enters marine environments (National Academy of Science, 1975). Human activities in the production, transportation and storage of petroleum is another route since such activities inevitably involve the risk of
*Corresponding author.

accidental spills that can be only minimized but not eliminated entirely. In recent years, leakage of gasoline from underground storage tanks primarily at automobile service stations and from pipelines has been experienced at an alarming rate. Marine spills are also now becoming a frequent and major source of water and coastal contamination (US Environmental Protection Agency, 1990). The Gulf War in 1991 resulted in the worst man-made environmental disaster, with millions of gallons of crude oil released from the destroyed oil wells into the waters and surrounding land, forming more than 330 oil lakes, covering an area of 49 square km (Salam, 1996). Such releases of large quantities of oil to marine and terrestrial environments present a longterm threat to all forms of life. There is now an increasing need for

0167-7012 / 98 / $19.00 1998 Elsevier Science B.V. All rights reserved. PII: S0167-7012( 98 )00020-7

156

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

cost-effective remediation technologies for hydrocarbon contamination.

2. Composition of petroleum hydrocarbons Petroleum hydrocarbons contain a complex mixture of compounds which can be categorized for simplicity into four fractions: saturates, aromatics, resins (N, S, O) and asphaltene (Shell International Ltd., 1983). The saturates fraction includes straight chain alkanes (normal alkanes), branched alkanes (isoalkanes), and cycloalkanes (naphthenes). The aromatic fraction contains volatile monoaromatic hydrocarbons such as benzene, toluene, xylenes etc., polyaromatic hydrocarbons, naphthenoaromatics and aromatic sulphur compounds such as thiophenes and dibenzothiophenes. It is noteworthy that the polyaromatic hydrocarbon (PAHs) fraction which is associated with oil contamination, includes both suspected and known carcinogens, the most toxic being benzo(a)pyrene. The resins (N, S, O) and asphaltene fractions consist of polar molecules containing nitrogen, sulphur and oxygen. Resins are amorphous solids which are truly dissolved in oil, whereas asphaltenes are large molecules colloidally dispersed in oil. The relative proportions of these fractions are dependent on many factors such as the source, geological history, age, migration and alteration of crude oil.

3. Biodegradation of petroleum hydrocarbons Current evidence suggests that in aquatic and terrestrial environments microorganisms are the chief agents for the biodegradation of molecules of environmental concern, including petroleum hydrocarbons (Alexander et al., 1982; Swanell and Head, 1994). Hydrocarbon-degrading bacteria, yeast and fungi are widely distributed in marine, fresh water and soil habitats. Bacteria and yeast appear to be the dominant degraders in aquatic ecosystems while fungi and bacteria are the main degraders in soil environments (Cooney and Summers, 1976; Hanson et al., 1997). There is a vast amount of literature on the subject of oil breakdown by microorganisms with several major review papers (Atlas, 1977; Higgins

and Gilbert, 1978; Bartha, 1986; Leahy and Colwell, 1990; Haramaya et al., 1997; Colwell and Walker, 1977). It is now generally accepted by the scientic community that no one species of microorganisms will completely degrade any particular oil (Colwell and Walker, 1977). The degradation of both crude and rened oils seems to involve a consortium of microorganisms, including both eukaryotic and prokaryotic forms. The most common genera known to be responsible for oil degradation comprise mainly Nocardia, Pseudomonas, Acinetobacter, Flavobacterium, Micrococcus, Arthrobacter, Corynebacterium, Achromobacter, Rhodococcus, Alcaligenes, Mycobacterium, Bacillus, Aspergillus, Mucor, Fusarium, Penicillium, Rhodotorula, Candida and Sporobolomyces, (Atlas, 1981; Bossert and Bartha, 1984; Atlas and Bartha, 1992; Sarkhoh et al., 1990). Of the various petroleum fractions, n-alkanes of intermediate length (C 10 –C 20 ) are the preferred substrates and tend to be most readily degradable (Singer and Finnerty, 1984), whereas shorter chain compounds are rather more toxic (Klug and Markovetz, 1971). Longer chain alkanes known as waxes (C 20 –C 40 ) are hydrophobic solids and consequently are difcult to degrade due their poor water solubility and bioavailability (Bartha, 1986); branched chain alkanes are also degraded more slowly than the corresponding normal alkanes (Singer and Finnerty, 1984). Many microorganisms are also known to degrade a wide range of aromatic compounds (Cerniglia, 1984; Gibson and Subramanian, 1984; Weissenfels et al., 1990). The degradation of polyaromatic hydrocarbons (PAH) by microorganisms depends to a large extent on their molecular weights among many other factors (Weissenfels et al., 1990; Balba, 1993). One of the concerns which has risen out of a number of in-vitro studies has been the possible production of certain intermediates from PAH degradation, particularly dihydrodiols, which are of greater toxicity than the parent compounds (Cerniglia, 1984). However, studies on PAH degradation in sediments suggested that accumulation of such compounds may not actually occur in the natural environment due to the rapidity with which they are further transformed (Herbes and Schwall, 1978). Most of the literature concerning the microbial transformation of PAHs has centered on the lower-molecular-weight compounds

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

157

such as naphthalene, anthracene and phenanthrene, but more recent studies have described the capabilities of other microorganisms to metabolize higher molecular weight compounds, including the white rot fungus Phanerochaete chrysosporium (Bumpus, 1989; Field et al., 1992; Sack and Gunther, 1993; Barr and Aust, 1994; Yateem et al., 1998; Suga and Lindstrom, 1997). Recently, a soil Mycrobacterium strain was also shown to be able to metabolize pyrene as the sole source of carbon and energy and its rate of metabolism was doubled by the addition of a solvent such as parafn oil (Jimenz and Bartha, 1996). Cycloalkane degradation rates are somewhat variable but tend to be much slower than alkanes and often involve several microbial species (Perry, 1981). Highly condensed aromatic and cycloparafnic structures, tars, bitumen and asphaltic materials have the highest boiling points and exhibit the greatest resistance to biodegradation (Atlas, 1981; Blakebrough, 1978). Asphaltenes are the product of petroleum hydrocarbons in soil that appear to be resistant to microbial degradation (Bossert and Bartha, 1984). It has been proposed that such residual material from oil degradation is analogous to, and could even be regarded as, humic material (Jobson et al., 1972). Due to its inert characteristics, insolubility and similarity to humic materials it is unlikely to be environmentally hazardous.

4. Bioremediation of oil-contaminated soil A variety of technologies are currently available to treat soil contaminated with hazardous materials, including excavation and containment in secured landlls, vapour extraction, stabilization and solidication, soil ushing, soil washing, solvent extraction, thermal desorption, vitrication and incineration (US Environmental Protection Agency, 1988; Russell, 1992). Many of these technologies, however, are either costly or do not result in complete destruction of contamination. On the other hand, biological treatment ‘bioremediation’ appears to be among the most promising methods for dealing with a wide range of organic contaminants, particularly petroleum hydrocarbons. The technology is also environ-

玻璃化

mentally sound, since it simulates natural processes, and since it can result in the complete destruction of hazardous compounds into innocuous products. The use of bioremediation to remove pollutants is typically less expensive than the equivalent physical / chemical methods (Russell, 1992). In situ bioremediation techniques also offer the potential to remediate contaminated soil and groundwater without the need for excavation, which is a major advantage. By using this approach, bioremediation can be implemented below existing buildings, without disturbing normal site operation. While bioremediation has many advantages, it is a site specic process and successful biological treatment of contaminated soils presents a challenge to environmental scientists and engineers for reasons including: (a) heterogeneity of the contaminants, for example, the contaminants can be found as solids, liquid, gases, free or tightly bound to the particulate matter; (b) extreme concentrations of hydrocarbons, for example, the presence of high concentrations of hydrocarbons can be inhibitory or toxic to the microorganisms while extremely low concentrations may not be adequate to support microbial activities; (c) variable site environmental conditions such as soil type and depth and soil microorganisms as well as physical conditions such as pH, temperature, oxygen availability, redox potential, moisture content and substrate bioavailability. These conditions can substantially affect the microbial growth and biodegradation of organic contaminants; and (d) bioremediation is also a slow process and subject to regulatory constraints which inuence its selection as the clean-up technology, particularly with respect to the required clean-up standards and the pressure for immediate site spill or clean-up mandated by public concern, which do not allow enough time for process optimization. The two general approaches to bioremediation are: (a) environmental biostimulation, such as through fertilizer addition, aeration and; (b) addition of adapted microbial hydrocarbon degraders by bioaugmentation. The rst is the most commonly used approach for eld application. The full claims of the effectiveness of soil seeding on enhancing oil degradation has not yet been fully demonstrated in the eld. The main goal of the bioremediation design should

158

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

be the creation of the most favourable conditions for microbial growth and activities. Bioremediation methods fall in two major categories: (a) on-site or above-ground treatment; such methods include landfarming / solid-phase bioremediation, composting, biologically enhanced soil washing and slurry bioreactors, with the rst two being the most commonly used and; (b) in situ bioremediation (in-place), for the remediation of subsurface soil and groundwater. This is usually achieved by the manipulation of the groundwater constituents or the stimulation of air movement, or both. To demonstrate that a bioremediation technology is potentially useful, it is important that the ability to enhance the rate of hydrocarbon biodegradation be demonstrated under controlled conditions. For practical reasons this cannot be easily accomplished in situ and thus must be accomplished in feasibility studies. Such studies are also used to provide information on the estimated cost and duration of treatment. These studies usually involve microbiological laboratory methods to measure the effectiveness of bioremediation under predetermined conditions. The goal of a laboratory feasibility study is to identify limiting factors and recommend ways to mitigate these limitation in the eld. The following section describes several established microbiological methods and approaches which can be used for the feasibility assessment of soil bioremediation.

5. Microbiological methods for bioremediation assessment

bioremediation is a useful tool for following the changes and discerning for microbes active in hydrocarbon degradation. A strong correlation between microbial counts and hydrocarbon degradation has been reported (Al-Awadhi et al., 1996; Song and Bartha, 1990). During landfarming of oil-contaminated soil, the total microbial counts in the form of total colony forming units (TCFU) were increased by four orders of magnitudes (Balba et al., 1998). Bacterial count is usually determined in representative soil composite samples, using the standard serial dilution and nutrient agar plate-counting techniques (Lorch et al., 1995). The plate counts for mesophilic bacteria are typically incubated at 308C for 24 h before bacterial counts are conducted. The counts are usually expressed in the form of CFU’s. The hydrocarbon-utilizing bacteria (HUB) can be assayed similarly, with the exception that solid mineral basal media are used and a suitable hydrocarbon compound such as n-hexadecane provided as the sole source of energy and carbon. In this case, hexadecane is added on a disc of sterilized lter paper which is then placed in the lid of the inverted plate. The plates are incubated at 308C for 72 h before HUB are counted. The use of specic hydrocarbon degrading bacterial counts provide additional information on the hydrocarbon biodegradation potential in a particular soil. The percentage of HUB to the total heterotrophic bacterial counts usually reects the extent of microbial acclimation and hydrocarbon degradation activities in an oil-contaminated sites. The agar plate microbial-counts technique has several limitations particularly when dealing with nonculturable microorganisms.

5.1. Microbial enumeration 5.2. Dehydrogenase activity
Initial soil analyses of the total heterotrophic microbial counts and specic hydrocarbon degrading microbial counts in the contaminated soil can provide useful information on soil biological activities, and the extent to which the indigenous microbial population has acclimated to the site conditions. The results will also indicate whether the soil contains a healthy indigenous microbial population capable of supporting bioremediation. In addition to the initial microbial assessment of the contaminated soil, monitoring microbial populations during the soil Biological oxidation of organic compounds is generally a dehydrogenation process, which is catalyzed by dehydrogenase enzymes (Lenhard, 1956; Paul and Clark, 1989; Page et al., 1982). Therefore, these enzymes play an essential role in the oxidation of organic matter by transferring hydrogen from the organic substrates to the electron acceptor. Many different specic enzyme systems are involved in the dehydrogenase activity of the soils. These systems are an integral part of the soil microorganisms and

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

159

reect to a great extent the soil biochemical activities. The assay of dehydrogenase in contaminated soil can be used as a simple method to examine the possible inhibitory effect of the contaminants on the soil microbial activities (Bartha and Pramer, 1965). For example, toluene and chloroform, if present at elevated concentration, can strongly inhibit soil dehydrogenase, but have little effect at low concentrations (Page et al., 1982). However, because the dehydrogenase activity depends on the total metabolic activities of soil microorganisms, its values in different soils do not always reect the total number of viable microorganisms isolated on a particular medium (Page et al., 1982). The most widely used method for the determination of soil dehydrogenase activities is the colorimetric method, involving the use of 2,3,5triphenyl tetrazolium chloride (TTC) which acts as an electron acceptor for many dehydrogenase enzymes (Page et al., 1982). When this compound is reduced by the catalytic effect of the soil dehydrogenase, it forms triphenyl formazan (TPH) which has a characteristic reddish colour which can be assayed at 485 nm (Page et al., 1982). The intensity of red colour produced from the dehydrogenase assay is a good index for microbial activities within the tested soil. However, several factors may affect the activities of soil dehydrogenase. Also nitrate, nitrite and ferric ions seem to inhibit dehydrogenase activity due to the ions acting as alternative electron acceptors.

5.3. Soil respirometric tests
Mineralization studies involving measurements of total CO 2 production can provide excellent information on the biodegradability potential of hydrocarbons in contaminated soils. The approach, which is considered a preliminary step in the feasibility study, provides rapid, relatively unequivocal timecourse data suitable for testing different biological treatment options, such as the effect of nutrient supplementation, microbial inoculation, etc. The test can be useful also for conrming active hydrocarbon degradation during full scale bioremediation. During the respiration tests, oxygen consumption and / or carbon dioxide evolution rates can be monitored by

using either expensive automated equipment which can handle a large number of samples simultaneously, or by simple respirometric-ask methods, which are commonly used (Bartha and Pramer, 1965; Pritchard et al., 1992). In the latter case, measurement of oxygen consumption can be carried out by the use of a U-tube manometer and barometric control, so when oxygen is consumed by aerobic metabolism, a measurable negative pressure develops within the respirometric ask which is directly related to the oxygen partial pressure. Carbon dioxide which is evolved during the respiration process is simultaneously trapped in a potassium hydroxide (KOH) solution located in a central well or in the side arm attached to the respirometric asks. The amount of carbon dioxide absorbed, is then measured by titrating the residual KOH with a standard solution of hydrochloric acid, after barium chloride is added to precipitate the carbonate ions. The levels of cumulative oxygen consumed and carbon dioxide evolved can then be calculated and plotted in mmol / kg of dry soil as a function of incubation time. In addition to the assessment of the degradation potential of petroleum hydrocarbons, the respirometric tests can also be applied to assess the possible inhibitory effects of heavy metals, toxic compounds, and pH on the soil microbial activities. The Biometer ask, involves the use of small amounts of soil and can be easily adapted to assess mineralization rates, particularly when a large number of samples need to be tested. Examples of the data generated from such a mineralization test are shown in Figs. 1 and 2 which present the respirometric results of routine monitoring of eld trials, involving the remediation of oil-contaminated desert soil by the windrow composting method, in Kuwait (Al-Daher et al., 1995). The soil treatment continued for a period of ten months during which composite soil samples were collected monthly for chemical analyses and respirometric assessment, using biometer asks. The respirometric tests were used to assess the soil microbial activities during the bioremediation program, by measurement of carbon dioxide production. The results presented in Fig. 1 show the relationship between the water content of the treated soil and microbial activities, measured in the form of carbon dioxide. Maximum respiration rate correlated well with the level moisture content of the soil. The

160

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

Fig. 1. Correlation between respiration rate and moisture content during soil bioremediation of petroleum hydrocarbons.

relationship between respiration rate versus total hydrocarbon concentration are presented in Fig. 2. The decrease in the carbon dioxide production rate, towards the end of the treatment, is possibly caused by the exhaustion of the readily degradable organic fraction.

5.4. Biodegradation microcosm test
There are many denitions of ‘microcosm’. A typical one is that of an intact, minimally disturbed piece of an ecosystem brought into the laboratory for study in its natural state (Prichard and Bourquin,

Fig. 2. Correlation between respiration rates and hydrocarbon content during soil bioremediation.

1984). Microcosms can vary in complexity from simple static soil jars of contaminated soil to highly sophisticated systems designed to enable variations in various environmental parameters encountered on site to be more accurately simulated in the laboratory. The microcosm design that closely models real environmental conditions is most likely to produce relevant results. In such experiments, it is important to include appropriate controls, such as sterile treatments, to separate the effects of abiotic weathering of oil from actual biodegradation. Soil microcosm experiments can be a useful tool to assess the biodegradation potential of hydrocarbon contamination and the development of models for predicting the fate of these pollutants. Mathematical equations can then be formulated to describe the kinetics of each of the processes involving transformation of specic hydrocarbon constituents under consideration. Concentrations of the hydrocarbon constituents and their degradation products can subsequently be monitored in various components of the microcosm, to obtain useful kinetic information in relation to their, equilibrium partitioning, biodegradation transformation behaviour, under predetermined environmental conditions. Additionally, the test can be used for screening bioremediation treatments to establish the most appropriate bioremediation strategy for large scale application (Balba et al., 1992; Compeau et al., 1991). Similarly, the biodegradation potential of hydrocarbons can be assessed by using slurry reactors (10–15% soil: water w / v), which offer several advantages over the soil microcosms. Due to more efcient mixing, aeration and improved substrate bioavailabilty, the duration of a treatability study can be signicantly reduced. During the treatability study, microcosms tests are usually monitored regularly for petroleum hydrocarbon degradation, by either sacricing whole microcosm systems or by subsampling techniques. Other parameters which may be monitored, in addition to petroleum hydrocarbons, include microbial counts, pH, nutrient concentration and moisture content. To determine the rate of hydrocarbon biodegradation, accurate and reliable analyses are critical. One of the recommended standard analyses for total petroleum hydrocarbons (TPH) is based on the use

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

161

of infrared (IR) absorption (Standard Methods for the Examination of Water and Wastewater, 1985; Potter, 1993). The method involves the extraction of the soil or the soil slurry with Freon, and the IR absorption of the extract is measured at 2930 cm 21 . This absorption band reects the C-H stretching vibrations of the hydrocarbons and, consequently, the measurement has to be performed in solvents that are free of C-H bonds. The IR method is therefore more sensitive to saturated hydrocarbons than to aromatics. For the best quantitative results, a standard has to be prepared from a balanced mixture of hydrocarbons, with a composition that approximates the hydrocarbon contamination in the samples to be analyzed. The method, however cannot provide information on the fate of individual hydrocarbon constituents. To obtain specic information on the biodegradation of oil constituents, the extract has to be rst fractionated by appropriate chromatographic techniques and the obtained fractions need to be then analyzed by gas chromatography tted with a capillary column and ame ionization detector (GC–FID) or mass spectrometry (GC–MS). Such methods allow detailed information to be obtained on the residual concentration of aliphatics, aromatics and biomarker constituents (Wang et al., 1994). However, the quantication of individual compounds is restricted to those which can be resolved by the gas chromatographic technique.

5.5. Biomarker compounds
The evaluation of hydrocarbon degradation in the eld is much more difcult than in the laboratory due to the heterogeneity of contamination. Polluted sites are often remarkably heterogeneous in their
Table 1 Correlation of C 18 :phytane ratio with TPH degradation Treatment C 18 :phytane ratio T0 Landfarming Control test Windrow piles Control test Static piles Control Test ND5Less than 0.3. 2.4 2.4 2.4 2.4 1.7 1.7 T6 0.3 2.3 0.4 2.4 0.5 1.6 T 12 ND 2.2 ND 2.2 ND 1.4

nature so that the initial analytical data vary from very low to very high concentrations over relatively small areas. In addition, large volumes, generally in the order of thousands of cubic meters of soil, are involved. Under such circumstances, it is very difcult to obtain statistically meaningful data without recourse to analyzing a massive number of samples which would be prohibitively expensive. Because of the difculties in the quantication of hydrocarbons in large scale bioremediation, the ratios of hydrocarbons compounds within the complex hydrocarbon mixture can be used to assess hydrocarbon biodegradation. Hydrocarbon degrading microorganisms usually degrade branched alkanes and isoprenoid compounds such as pristane, phytane and hopane compounds at much slower rates than straight-chain alkanes. Therefore, the ratio of straight-chain alkanes to these highly branched biomarker compounds can reect the extent to which microorganisms have degraded the hydrocarbons in a petroleum mixture (Wang et al., 1994; Prichard and Costa, 1991; Kennicutt, 1988). This ratio concept is based on the assumption that nonbiodegradation processes such as weathering, volatilization and leaching, will not produce differential losses of normal and branched hydrocarbons that have similar gas chromatographic and correspondingly, chemical behaviour (Kennicutt, 1988). An example of the data generated from such analyses is shown, in Table 1. These results were obtained from eld demonstration involving the bioremediation of oil-contaminated soil in Kuwait (2160 m 3 ), using three different methods, namely landfarming, composting piles (480 m 3 ), and static bioventing piles (240 m 3 ). The results summarizes the progressive changes in C 18 :phytane ratio during the course of soil remediation. Hydrocarbon

TPH concentration (mg / kg) T0 39 400 39 400 34 700 35 900 14 400 14 100 T6 14 000 35 500 19 400 39 800 8500 13 600 T 12 7200 31 700 9500 30 600 4600 12 200

TPH reduction (%) 81.7 19.5 72.6 14.8 68.1 13.5

162

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

degradation in the treated soil was accompanied by signicant reduction in the ratio, compared to little or no change in the control tests (Balba et al., 1998). However, the method has some limitations, due to the fact that branched alkanes including phytane biodegrade slowly, which means that the C 18 :phytane ratio underestimates hydrocarbon biodegradation. Also, octadecane (C 18 ) usually degrades rapidly, making the ratio technique useful only during the early stages of oil degradation. Hopane compounds are also molecular fossils that are derived from the biomass that give rise to the crude oil. These molecules are slowly biodegradable and thus become enriched within the residual oil as the oil weathers by evaporation and biodegradation. Some of these compounds have been successfully used as biomarkers for oil degradation assessment (Prince et al., 1994).

crotox test assesses the toxicity of the soil extracts by measuring the reduction in light emission by Photobacterium phosphoreum, while the Ames Test examines the mutagenic effect of the contaminated soil on Salmonella typhimurium (Maron and Ames, 1983).

5.7. Microbial survival test and tracking of GEM
There is a great deal of interest in the use of genetically modied or engineered microorganisms (GEM) to enhance oil degradation, particularly the degradation of high molecular weight polyaromatic compounds and alkane hydrocarbons. Microorganisms with enhanced capabilities to degrade particularly aromatic hydrocarbons and their derivatives, have already been developed (Thomas and Ward, 1994; Krume et al., 1994). Although technologies based on these concepts hold promise for improved bioreactor performance, experience gained from bioaugmentation tests suggest that the use of GEMs will be ineffective without development of techniques to improve their survival in the face of competition from indigenous microbial population (Atlas, 1992). Convenient, economical and effective methods of tracking engineered microorganisms have been developed to enable the examination of their survival, transport and ecological impact, when released in new environments (Veal and Stokes, 1992; O’Donnell and Hopkins, 1993). This topic has been addressed in detail by other authors in this special issue of the Journal of Microbiological Methods.

5.6. Ecological impact and toxicity assessment
In addition to the demonstration of the treatment efcacy, it is necessary to demonstrate that bioremediation does not produce any toxic intermediate products and to avoid undesired environmental and ecological effects (Cerniglia, 1984; Prince et al., 1994). Fertilizers should not be applied at excessive rates and the use of sodium nitrate is discouraged because of the problem of introducing a potential contaminant into groundwater (Russell, 1992). Nitrate has been known to cause the ‘blue baby’ syndrome in small infants and the maximum concentration of nitrate in drinking water supply is 10 mg / l. Necessary engineering measures should also be considered to ensure the containment of the remediation zone and prevent leachate migration outside the treatment zone (Ellis et al., 1990). Background toxicity prior to soil bioremediation and after treatment can also be measured by using appropriate toxicity tests. The tests used for this purpose may include plant, microtox and Ames tests. In the plant tests, the effect of the contaminated soils on the growth and germination of selected monocotyledonous and dicotyledonous plant species and ability of soil to support sustainable growth are assessed (El-Nawawy et al., 1995). In addition to these phytotoxicity tests, acute toxicity of contaminated soil, can be determined by microtox assay and Ames test (Mathew and Hastings, 1987). The mi-

6. Concluding remarks Bioremediation is a cost-effective and environmentally sound remediation technology, particularly for dealing with petroleum hydrocarbon contamination. However, the degradation rates of hydrocarbons are site specic and are limited by the metabolic capabilities of the hydrocarbon-degrading microbial populations and also by a wide range of environmental factors. The effectiveness of the bioremediation depends therefore on the success in identifying the rate-limiting factors and optimizing them in the feasibility studies. In these studies, microbially based methods are usually used to determine site feasibili-

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164

163

ty, the rate and extent of contaminants biodegradation that might be attained during remediation, and to provide design criteria. Feasibility studies are therefore essential and can have an enormous impact on the cost of full-scale remediation. Depending on the circumstances, screening tests such as microbial plate counts and enzyme assessment may be used to determine if existing conditions are favourable for microbial growth and respirometer tests provide conrmation that the microbial population is metabolically active. Treatability studies conducted with soil or slurries can be tested under several conditions, including unmodied microcosm, nutrient-amended and biologically inhibited conditions, to provide useful data on the rate and extent of conversion of contaminants. During bioremediation of hydrocarbons in the eld, the rate of degradation is largely controlled by the rate of supply of nutrients and oxygen, which makes it difcult sometimes, to extrapolate directly the results from laboratory to bioremediation in the eld. References
Al-Awadhi, N., Al-Daher, R., ElNawawy, A., Balba, M.T., 1996. Bioremediation of oil-contaminated soil in Kuwait. I. Landfarming to remediate oil-contaminated soil. J. Soil Contamin. 5 (3), 243–260. Al-Daher, R., El-Nawawy, A., Al-Awadhi, N., 1995. Evaluation of on-site bioremediation of oil-contaminated soil through composting soil piles. Final Report, KISR4625, Kuwait Institute of Scientic Research, Kuwait, 175–180. Alexander, S.K., Schropp, S.J., Schwarz, J.R., 1982. Spatial and seasonal distribution of hydrocarbon-utilizing bacteria of sediment from the northwestern Gulf of Mexico. Contrib. Mar. Sci. 25, 13–19. Atlas, R.M., 1977. Stimulated petroleum biodegradation, CRC Crit. Rev. Microbiol. 5, 371–386. Atlas, R.M., 1981. Microbial degradation of petroleum hydrocarbons: An environmental perspective. Microbiol. Rev. 45, 180–209. Atlas, R.M., 1992. Molecular methods for environmental monitoring and containment of genetically engineered microorganisms. Biodegradation 3, 137–146. Atlas, R.M., Bartha, R., 1992. Hydrocarbon biodegradation and oil-spill bioremediation. In: Marshall, K.C. (Ed.), Advances in Microbial Ecology, vol. 12, Plenum Press, New York, pp. 287–338. Balba, M.T., 1993. Microorganisms and detoxication of industrial waste. In: Gareth Jones, D. (Ed.), Exploitation of Microorganisms, Chapman and Hall, London, pp. 411–409. Balba, M.T., Al-Daher, R., Al-Awadhi, N., 1998. Bioremediation of oil-contaminated soil in Kuwait. Environ. Int. 24 (1), 163– 173.

Balba, M.T., Ying, A.C., McNeice, T.G., 1992. Bioremediation of contaminated soil: bench-scale to eld application. In: The Proceedings of National Research and Development Conference on the Control of Hazardous Materials, HMCRI, Washington, D.C., pp. 145–151. Barr, D.P., Aust, S.D., 1994. Mechanisms white rot fungi use to degrade pollutants. Environ. Sci. Technol. 28 (2), 79A. Bartha, R., 1986. Biotechnology of petroleum pollutant biodegradation. Microbiol Ecol. 12, 155–172. Bartha, R., Pramer, D., 1965. Features of a ask and method for measuring the persistence of and biological effects of pesticides in soil. Soil Sci. 100, 86–170. Blakebrough, N., 1978. Interactions of oil and microorganisms in soil. In: Chater, K.W., Somerville, J.H. (Eds.), The Crude Oil Industry and Microbial Ecosystems, Heydon and Son Ltd., London, pp. 28–40. Bossert, I., Bartha, R., 1984. The fate of petroleum in soil ecosystems. In: Atlas, R.M. (Ed.), Petroleum Microbiology, Macmillan, New York, pp. 435–473. Bumpus, J.A., 1989. Biodegradation of polyaromatic hydrocarbons by Phanerochaete chrysosporium. Appl. Environ. Microbiol. 55, 154–155. Cerniglia, C.E., 1984. Microbial metabolism of polycyclic aromatic hydrocarbons. Adv. Appl. Microbiol. 30, 31–37. Colwell, R.E., Walker, J.D., 1977. Ecological aspects of microbial degradation of petroleum in the marine environment. Crit. Rev. Microbiol. 5, 423–445. Compeau, G.C., Mahaffey, W.D., Patras, L., 1991. Full-scale bioremediation of contaminated soil and water. In: Sayler, G.S. (Ed.), Environmental Biotechnology for Waste Treatment, Plenum Press, New York, pp. 91–109. Cooney, J.J., Summers, R.J., 1976. Hydrocarbon-using microorganisms in three fresh water ecosystems. In: Sharpley, J.M. et al. (Eds.), Proceedings of the Third International Biodegradation Symposium, Applied Sciences, London, pp. 141–156. Ellis, B., Balba, M.T., Theile, P., 1990. Bioremediation of oilcontaminated land. Environ. Technol. 11, 443–455. El-Nawawy, A., Al-Daher, R., Yateem, A., Al-Awadhi, N., 1995. Bioremediation of oil-contaminated soil in Kuwait by landfarming technology. Final Report (KISR) 4596), Kuwait Institute for Scientic Research, Kuwait. Field, J.A., Jong, E.D., Costa, G.F., Bont, J.A., 1992. Biodegradation of polyaromatic hydrocarbons by new isolates of white rot fungi. Applied Environ. Microbiol. 58 (7), 2219–2226. Gibson, D.T., Subramanian, V., 1984. Microbial degradation of aromatic hydrocarbons. In: Gibson, D.T. (Ed.), Microbial Degradation of Organic Compounds, Dekker, New York, pp. 181–282. Hanson, K.G., Nigam, A., Kapadia, M., Desai, A., 1997. Bioremediation of crude oil contamination with Acinetobacter sp. A3. Current Microbiol. 35, 191–193. Haramaya, S., Venkateswaran, K., Toki, H., Komukai, S., Goto, M., Tanaka, H., Ishihara, M., 1997. Degradation of crude oil by marine bacteria. J. Mar. Biotechnol. 3, 239–243. Herbes, S.E., Schwall, L.R., 1978. Microbial transformation of polyaromatic hydrocarbons in pristine and petroleum-contaminated sediments. Appl. Environ. Microbiol. 35, 306–316. Higgins, I.J., Gilbert, P.D., 1978. The biodegradation of hydro-

164

M.T. Balba et al. / Journal of Microbiological Methods 32 (1998) 155 – 164 Miller, R.N. (Eds.), Hydrocarbon Bioremediation, Lewis Publishers, London, pp. 107–124. Pritchard, P.H., Mueller, J.H., Rogers, J.C., Kremer, F.V., Glaser, J.A., 1992. Oil spill bioremediation:experience, lessons and results from the Exxon Valdez oil spill in Alaska. Biodegradation 3, 315–335. Russell, D.L., 1992. Remediation Manual for Petroleum-Contaminated Sites, Technomic Publishing Co., Lancaster, PA., USA. Sack, U., Gunther, T., 1993. Metabolism of PAH by fungi and correlation with extracellular enzymatic activities. J. Basic Microbiol. 33 (4), 269–277. Salam, A.A., 1996. Remediation and rehabilitation of oil-lake beds. In: Al-Awadhi, N., Balba, M.T., Kamizawa, C. (Eds.), Environmental Disaster, Elsevier, Amsterdam, pp. 117–139. Sarkhoh, N.A., Ghannoum, M.A., Ibrahim, A.S., Stretton, R.J., Radwan, S.S., 1990. Crude oil and hydrocarbon degrading strains of Rhodococcus: Rhodococcus strains isolated from soil and marine environments in Kuwait. Environ. Pollut. 65, 1–18. Shell International Ltd., 1983. The chemistry of petroleum. In: The Petroleum Handbook, 6th ed., Elsevier, New York, pp. 223–264. Singer, M.E., Finnerty, W.R., 1984. Microbial metabolism of straight-chain and branched alkanes. In: Atlas, R.M. (Ed.), Petroleum Microbiology, Macmillan Inc., pp. 1–59. Song, H.G., Bartha, R., 1990. Effect of jet fuel spills on the microbial community of soil. Appl. Environ. Microbiol. 56 (3), 646–651. Standard Methods for the Examination of Water and Wastewater, 1985. APHA-AWWA-WPCF, pp. 496–503. Suga, S.F., Lindstrom, J.E., 1997. Braddock, Environmental inuences on microbial degradation of Exxon Valdez oil on the shorelines of Prince William Sound Alaska. Environ. Sci. Technol. 31 (5), 1564–1572. Swanell, R.P.J., Head, I.M., 1994. Bioremediation come of age. Nature 368, 396–397. Thomas, J.M., Ward, C.H., 1994. Introduced organisms for subsurface bioremediation. In: Norris, Hinchee, Brown et al. (Eds.), Handbook of Bioremediation, Lewis Publishers, London, pp. 227–339. US Environmental Protection Agency, 1988. Technology Screening Guide for Treatment of CERCLA soils and sludges. EPA / 540 / 288 / 004, Washington DC. US Environmental Protection Agency, 1990. Bioremediation of Hazardous Wastes, EPA / 600 / 9-90 / 041, Ofce of Research and Development, Washington DC. Veal, D.A., Stokes, H.W., 1992. Genetic exchange in natural microbial communities. In: Marshall, K.C. (Ed.), Advances in Microbial Ecology, vol. 12, Plenum Press, New York, pp. 383–430. Wang, Z., Fingas, M., Li, K., 1994. Fractionation of a light crude oil and identication and quantication of aliphatic, aromatic, and biomarker compounds by GC–FID and GC–MS Part I. J. Chromatogr. Sci. 32, 361–366. Weissenfels, W.D., Beyer, M., Klein, J., 1990. Degradation of phenanthrene, uorene, uoranthene by pure bacterial cultures. Appl. Microbiol. Biotechnol. 32, 479–484. Yateem, A., Balba, M.T., Al-Awadhi, N., El-Nawawy, A.S., 1998. White rot fungi and their role in remediating oil contaminated soil. Environ. Int. 24 (1), 181–187.

carbons. In: Chater, K.W., Somerville, H.J. (Eds.), The Oil Industry and Microbial Ecosystems, Hedon and Son Ltd., London, pp. 80–117. Jimenz, I., Bartha, R., 1996. Solvent-augmented mineralization of pyrene by a Mycobacterium sp. Appl. Environ. Microbiol. 62 (7), 2311–2316. Jobson, A., Cook, F.D., Wastelake, D.W.S., 1972. Microbial utilization of crude oil. Appl. Microbiol. 32, 1082–1089. Kennicutt, M.C., 1988. The effect of biodegradation on crude oil bulk and molecular composition. Oil Chem. Pollut. 4, 89–112. Klug, M.J., Markovetz, K., 1971. Utilization of aliphatic hydrocarbons by microorganisms. Adv. Microbial Physiol. 5, 1–39. Krume, M.L., Smith, R.L., Egestorff, J. et al., 1994. Behaviour of pollutant-degrading microorganisms in aquifers: prediction of genetically engineered organisms. Environ. Sci. Technol. 28, 1134–1138. Leahy, J.G., Colwell, R.R., 1990. Microbial degradation of hydrocarbons in the environment. Microbiol. Rev. 54, 305–315. Lenhard, G., 1956. Die dehydrogenaseaktivitat des Bodens als Mass fur die Mikroorganismentatigkeit im Boden. Z. Panzenernaehr. Dueng. Bodenkd. 73, 1–11. Lorch, H.J., Benckieser, G., Ottow, J.C.G., 1995. Basic methods for counting microorganisms in soil and water. In: Alef, K., Nannipieri, P. (Eds.), Methods in Applied Soil Microbiology and Biochemistry, Academic Press, New York, pp. 146–161. Maron, D.M., Ames, B.N., 1983. Revised methods for the Salmonella mutagenicity test. Mutat. Res. 113, 173–215. Mathew, E., Hastings, L., 1987. Evaluation of toxicity test procedure for screening treatability potential of waste in soil. Toxic. Assessment 2, 265–281. National Academy of Science, 1975. Petroleum in the Marine Environment, National Academy Press, Washington, D.C., p. 107. O’Donnell, A.G., Hopkins, D.W., 1993. Extraction, detection and identication of genetically engineered microorganisms from soils. In: Edward, C. (Ed.), Monitoring Genetically Manipulated Microorganisms in the Environment, Wiley Publishers, pp. 111–131. Page, A.L., Miller, R.H., Keeney, D.R., 1982. Methods of Soil Analysis, part II, 2nd ed., ASA-SSSA, Wisconsin, pp. 937–970. Paul, E.A., Clark, F.E., 1989. Soil Microbilogy and Biochemistry, Academic Press, New York, pp. 46–48. Perry, J.J., 1981. Microbial metabolism of cyclic alkanes. In: Atlas, R.M. (Ed.), Petroleum Microbiology, Macmillan Inc., pp. 62–128. Potter, L., 1993. Analysis of petroleum contaminated soil and water: An Overview. In: Calabrese, E., Kostecki, P. (Eds.), Principle and Practices for Petroleum Contaminated Soils, Lewis Publishers, London, pp. 1-24. Prichard, P.H., Bourquin, A.W., 1984. The use of microcosms for evaluation of interactions between pollutants and microorganisms. In: Marshall, K.C. (Ed.), Advances in Microbial Ecology, vol. 7, Plenum Press, New York, pp. 133–215. Prichard, H.P., Costa, C.F., 1991. EPA’s Alaska oil spill bioremediation report. Environ. Science Technol. 25, 372–379. Prince, R.C., Clark, J.R., Lindstrom, J.E. et al., 1994. Bioremediation of the Exxon Valdez oil spill: Monitoring safety and efcacy. In: Hinchee, R.E., Alleman, B.C., Hoeppel, R.E.,


相关文章:
更多相关标签: