Escuela Técnica Superior De Ingenieros De Minas

ESCUELA TÉCNICA SUPERIOR DE INGENIEROS DE MINAS

PROYECTO FIN DE CARRERA

Departamento de Ingeniería Química y Combustibles

BIOREMEDATION OF OIL SPILLS

YOLANDA SANCHEZ-PALENCIA GONZALEZ OCTUBRE 2011

ÍNDICE

Abstract ...... vi

1 AIM AND SCOPE ...... 8

2 BACKGROUND ...... 9

2.1 Petroleum as source of hydrocarbons in the environment: pollution of marine environments...... 9

3 DEFINITIONS ...... 10

4 PETROLEUM AND MICROBIAL DEGRADATION OF HIDROCARBONS

IN MARINE ENVIRONMENTS ...... 12

4.1 Microbial degradation of hydrocarbons ...... 12

4.2 Microbial hydrocarbon-degrading communities...... 13

4.3 Obligate hydrocarbonoclastic marine bacteria...... 14

5 BIOAUGMENTATION AND BIOSTIMULATION ...... 16

5.1.1 Bacteria consortia ...... 19

5.2 Biostimulation ...... 33

5.2.1 Using guano as fertilizer ...... 35

5.3 Bioaugmentation versus biostimulation...... 38

5.4 Commercial products: ongoing scenario ...... 39

5.5 Confined systems and real-case studies: bridging the gap...... 40

6 PRODUCTION OF LIPIDS AND STORAGE COMPOUNDS BY HCB

BACTERIA...... 43

6.1 Production of lipids by hydrocarbonoclastic marine bacteria...... 43

6.2 Storage compound accumulation in hydrocarbonoclastic bacteria...... 44

6.3 Production of biosurfactants by oil-degrading bacteria ...... 44

6.4 Production and export of lipids by Alcanivorax strains ...... 45

6.5 Concluding remarks and future perspectives ...... 45

7 DISPERSANTS AND OIL MINERAL AGREGATES ...... 47

7.1 Use of dispersants ...... 47

7.2 Oil–mineral interaction in particle image velocimetry (PIV) ...... 48

8 COLD ENVIRONMENTS ...... 58

8.1 Properties of spilled oil ...... 59

8.2 Fate of oil spilled in the cold environments ...... 59

8.3 Cold-adapted, oil-degrading microorganisms...... 60

8.4 Factors affecting biodegradation...... 61

8.5 Techniques, successful cases, challenges and constraints ...... 65

9 CONCLUSIONS ...... 69

10 REFERENCES ...... 70

iii

ÍNDICE DE FIGURAS

Figure 1: A. borkumensis ...... 15

Figure 2: CO2 Evolution during degradation of phenanthrene (NaCl: 30 g/l) ...... 25

Figure 3: Degradation of different concentrations phenanthrene ...... 26

Figure 4: Degradation of different concentrations fluorene ...... 26

Figure 5: Degradation of different concentrations pyrene...... 27

Figure 6: Degradation of different concentrations benzo(e)pyrene...... 27

Figure 7: Gas chromatography analysis for residual aliphatics through biodegradation (DMCB) ...... 31

Figure 8: Residual concentrations in independent assays (DMC B) ...... 32

Figure 9: Oil degradation using guano or ammonium sulfate as the nitrogen source ...... 38

Figure 10: ANS oil droplet with different types of minerals ...... 51

Figure 11: Mineral–oil interactions (green area= crude oil and red spots = minerals)...... 51

Figure 12: Effect of dispersants ...... 52

Figure 13: Effect of dispersants ...... 52

Figure 14: Vector field at the surface of an oil droplet of ANS ...... 53

Figure 15: Vector field around a moving oil droplet at different positions ...... 54

Figure 16: Vector field of modified Kaolin and a static oil droplet (MESA)...... 55

Figure 17: Comparison of interaction of hydrophilic minerals with different oil types...... 56

Figure 18: Effect of hydrophobic property on oil–mineral interaction ...... 56

ÍNDICE DE TABLAS

Table 1: Biodegradation of crude petroleum oil in soil by bacterial consortium...... 21

Table 2: Nutrient (NPK) and pH status of the bioremediated and normal soils ...... 22

Table 3: Factors influencing bioaugmentation and biostimulation processes ...... 41

Table 4: Bioaugmentation and biostimulation: field and real case studies ...... 42

Table 5: Composition of oils ...... 50

Table 6: Physical properties of minerals ...... 50

Abstract

Due to the increasing demand of petroleum everywhere, and the great amount of spills, accidents and disasters, there is an urgent need to find an effective, non-cost and harmless method to clean up the affected areas. There are microorganisms in nature (bacteria and fungi, mainly) that feed on hydrocarbons and transform them into others harmless chemical substances. These bacteria produce enzymes that degrade oil very effectively. This natural process can be accelerated by adding more bacteria or providing nutrients and oxygen to facilitate their growth, which is called ―bioaugmentation and biostimulation‖. Through this project we discover that these processes can be affected by different factors making difficult the biodegradation execution and opening a gap between the laboratory experiments and the real cases. Therefore, there is much remain to be done and a lot of study ahead to make this technique available in a great scale.

BIOREMEDIATION OF OIL SPILLS

1 AIM AND SCOPE

This Project is a review of the last advances and discovers related to biodegradation of petroleum hydrocarbons both in water and in soil during last years. It is included temperate and cold environments, bioaugmentation (focusing on consortia rather than a single strain) and biostimulation, the use of dispersants and oil mineral aggregates and the production of lipids and storage compounds by hydrocarbonoclastic bacteria as a way of feeding.

2 BACKGROUND

2.1 Petroleum as source of hydrocarbons in the environment: pollution of marine environments

The world demand for oil in 2008 was 85.62 million barrels per day (OPEC 2009). Petroleum-based products are the major source of energy for industry as well as daily life and represent the raw material for many chemical processes. The global transport and use of both petroleum and its derivatives have made petroleum hydrocarbons (PHCs) major contaminants in both prevalence and quantity in the environment.

Along with transport, processing of crude oil represent an enormous potential for oil spills caused by tanker accidents exemplified by the one of the tanker Amoco Cadiz, which polluted the French coast of Brittany in 1978 or the Exxon Valdez tanker, which severely affected the Alaska coast in 1989 (Syutsubo et al., 2001).

It was estimated that about 2 to 10 millions of tons of crude petroleum oil enter marine environments annually, the majority of which is generated from anthropogenic sources but also from natural sources like plants, algae, and bacteria (Tyagi, 2011).

Another major part of oil enters the environment from municipal and industrial wastes and runoffs (Head and Swanell, 1999). In the case of petrol stations, refineries and storage tanks, leakage may be small, but continuous and prolonged in time. In this case, the probability of contamination of the groundwater is higher, thus making efficient remediation desirable and needed (Tyagi, 2011)

3 DEFINITIONS

Microcosms: are artificial, simplified ecosystems that are used to simulate and predict the behavior of natural ecosystems under controlled conditions. Open or closed microcosms provide an experimental area for ecologists to study natural ecological processes. Microcosm studies can be very useful to study the effects of disturbance or to determine the ecological role of key species (Wikipedia, 2011).

Mesocosm: is an experimental tool that brings a small part of the natural environment under controlled conditions. In this way mesocosms provide a link between observational field studies that take place in natural environments, but without replication, and controlled laboratory experiments that may take place under somewhat unnatural conditions. When pursuing a laboratory experiment, the experimenter cannot account for every possible factor that would normally occur in the original aquatic environment. Mesocosms circumvent this problem as the experiment is performed in the natural environment, but in an enclosure that is small enough that key variables can be brought under control (Wikipedia, 2011).

Agar: A gelatinous material obtained from the marine algae, used as a bacterial culture medium, in electrophoresis and as a food additive (Wiktionary, 2011).

Agarose: A polymeric cross-linked polysaccharide extracted from the seaweed agar; used to make gels that are used in electrophoresis (Wordnik,2011).

Bioemulsifiers: High molecular weight biosurfactants (bio-emulsifiers) produce stable emulsions, which allow bacteria to adhere strongly to hydrophobic surfaces and then degrade large biological complexes (Nerurkar et al., 2009).

Metabolites: are the intermediates and products of metabolism. The term metabolite is usually restricted to small molecules. A primary metabolite is directly involved in normal growth, development, and reproduction. A secondary metabolite is not directly involved in those processes, but usually has an important ecological function. Examples include antibiotics and pigments (Wikipedia, 2011).

Electrophoresis: also called cataphoresis, is the motion of dispersed particles relative to a fluid under the influence of a spatially uniform electric field (Wikipedia, 2011).

Denaturation: a structural change in macromolecules caused by extreme conditions (Wikipedia, 2010).

Temperature Gradient Gel Electrophoresis (TGGE) and Denaturing Gradient Gel Electrophoresis (DGGE): are forms of electrophoresis which use either a temperature or chemical gradient to denature the sample as it moves across an acrylamide gel. TGGE and DGGE can be applied to nucleic acids such as DNA and RNA, and (less commonly) proteins. TGGE relies on temperature dependent changes in structure to separate nucleic acids. DGGE was the original technique, and TGGE a refinement of it (Wikipedia 2008).

Amphiphile: is a term describing a chemical compound possessing both hydrophilic (water-loving, polar) and lipophilic (fat-loving) properties (Wikipedia 2011)

Gas chromatography (GC): is a common type of chromatography used in analytical chemistry for separating and analysing compounds that can be vaporized without decomposition (Wikipedia 2011).

4 PETROLEUM AND MICROBIAL DEGRADATION OF HIDROCARBONS IN MARINE ENVIRONMENTS

4.1 Microbial degradation of hydrocarbons

The impact of oil spill incidents on marine environments is enormous, and it has triggered research on cost-effective, environmentally benign cleanup strategies. Physical and chemical methods remove rapidly the majority of beached oil, but a significant part of the contaminants remains at the site. In natural degradative processes that remove the remaining oil, bacteria are the predominant agents of hydrocarbon degradation in the environment (Röling et al., 2002). Petroleum oil is highly toxic to the majority of living organism. Its biodegradation in natural ecosystems is complex and depends on the nature of the oil, on the composition and physiological status of the indigenous microbial community, and on a variety of environmental factors which influence microbial activities, e.g., temperature, physical state of oil pollutants, nutrients, etc. (Atlas 1981).

A recent review pointed out that there are 79 bacterial genera that can use hydrocarbons as a sole source of carbon and energy, while nine cyanobacterial genera, 103 fungal genera, and 14 algal genera are known to also comprise members capable of degrading or transform hydrocarbons (Head et al., 2006). In the same way, 56 yeasts species are able to utilize hydrocarbons (Atlas, 1981).

Crude petroleum is mainly composed of linear and branched-chain alkanes, cycloalkanes and aromatics, but it also contains small amounts of oxygen-, nitrogen- and sulfur- containing compounds, such as phenol, indole and thiophene, respectively. Alkenes are present in several refined petroleum products but usually not in crude petroleum, being mainly obtained by catalytic cracking at high temperatures, whilst the heavy asphaltenes are present in crude but not in refined fractions (Matar, 1992). Although n-alkanes are the most biodegradable petroleum hydrocarbons (PHCs), those with 5–10 carbon atoms are inhibitory to most hydrocarbon degraders because they can disrupt the lipid membrane of microorganisms (Bartha, 1986). On the other hand, PHCs with C20–C40, (waxes), are hydrophobic solids at physiological temperatures. This probably explains their low

biodegradability (Bartha and Atlas, 1977). Interestingly, some bacteria also produce waxes when they degrade crude oil (Ishige et al., 2003; Walker and Colwell, 1976).

In general, aliphatic paraffins are more readily degraded than aromatic hydrocarbons. Saturated compounds are degraded more readily than unsaturated compounds, and branched chains are decomposed less readily than straight chain compounds. However, low molecular weight aromatics, such as benzene, toluene and xylene, which are among the toxic compounds in petroleum, are also very readily degraded by marine microorganisms (Atlas, 1995).

The biodegradability of the oil components generally decreases in the following order: n- alkanes>branched-chain alkanes>branched alkenes>lowmolecular- weight n-alkyl aromatics>monoaromatics>cyclic alkanes>polycyclic aromatic hydrocarbons (PAHs) >> asphaltenes (van Hamme et al., 2003).

Some strains able to degrade petroleum hydrocarbons and aromatic hydrocarbons such as benzene, toluene, ethylbenzene and xylene are Pseudomonas, Rhodococcus and Ralstonia, as well as polyaromatic hydrocarbons such as naphthalene: Pseudomonas, phenanthrene : Pseudomonas and Haemophilus, anthracene : Rhodococcus, pyrene:Haemophilus and Mycobacterium, and the highly carcinogenic benzo[a]pyrene:Rhodococcus and Mycobacterium. (Tyagi et al., 2011)

4.2 Microbial hydrocarbon-degrading communities

Of the diverse range of oil-degrading bacteria isolated to date, less than a quarter was isolated from marine environments (Yakimov et al., 2007). Biodegradation of petroleum in marine environments is principally carried out by diverse bacterial populations, including various Pseudomonas species (Atlas 1995), which are widely distributed in these environments. These strains had been isolated from diverse sites, indicating that they are distributed ubiquitously (Atlas, 1995; Head et al., 2006). In pristine waters, hidrocarbondegrading bacteria (HDB) comprise less than 1% of the total bacterial population (Atlas 1981), but they multiply and grow rapidly in oil-polluted waters, where

they can constitute 80% to 90% of the microbial community (Harayama et al., 1999; Kasai et al., 2002a; Syutsubo et al., 2001).

The presence of hydrocarbons in the environment frequently brings about an in situ selective enrichment of hydrocarbon-utilizing microorganisms. There are two essential characteristics that define hydrocarbonutilizing microorganisms: (1) membrane-bound groupspecific oxygenases; and (2) mechanisms for optimizing contact between the microorganisms and the waterinsoluble hydrocarbon (Rosenberg et al., 1992).

After an oil spill, a remarkable increase of the relative abundance of oil degraders is usually observed. Several culture-independent studies of oil-impacted marine environments have shown that some bacteria are rapidly and strongly selected when hydrocarbon degradation is stimulated by the addition of nutrients (Head et al., 2006). Alcanivorax spp. represent a good example of these bacteria, because they were shown to increase from being undetectable to constituting 70–90% of the bacterial population in oil- treated sea water within 1–2 weeks of nutrient amendment (Syutsubo et al., 2001). Cycloclasticus pugetii constitutes another example of ―bacterial selection‖ after an oil spill accident (Maruyama et al., 2003).

4.3 Obligate hydrocarbonoclastic marine bacteria

An unusual group of marine HDB has been recognized and described over the past few years and has been shown to play an important role in the biological removal of petroleum hydrocarbons from contaminated sites. This group of marine bacteria, the obligate hydrocarbonoclastic marine bacteria (OHCB), exhibits a narrow substrate spectrum (obligate hydrocarbon utilization). Only a few species are able to metabolize substrates other than hydrocarbons (Yakimov et al., 2007). Within this group of γ proteobacteria, a number of new genera and families have recently been discovered, comprising the genera Alcanivorax (Yakimov et al., 1998), Cycloclasticus (Dyksterhouse et al., 1995), Oleispira (Yakimov et al., 2003), Oleiphilus (Golyshin et al., 2002), and Thalassolituus (Yakimov et al., 2004). Together with species of the genus Neptunomonas (Hedlund et al., 1999) and Marinobacter (Gauthier et al., 1992), these bacteria constitute an important group of marine hydrocarbon-degrading bacteria.

From these, Alcanivorax borkumensis (SK2) became a model organism to study the OHCB because of its predominance and pivotal role in the degradation of hydrocarbons.

Borkumensis strain SK2, is a slowly growing, alkanedegrading Gram-negative bacterium, which is present only at low titers in pristine waters. After an oil spill, its population can increase dramatically, and it can become the most abundant organism in oil-polluted waters (Harayama et al., 1999; Kasai et al., 2002a; Syutsubo et al., 2001).

The predominance of A. borkumensis (that we can observe in figure 1) in early stages of petroleum degradation has been reported in microcosms and mesocosms studies (Röling et al., 2002; Capello et al., 2007) as well as during a field-scale experiment (Röling et al., 2004). Alcanivorax spp. seem to be early colonizers after an oil spill event, because after an initial rapid increase in population size, it declines to much lower numbers within a few weeks. This phenomenon is related to the depletion of saturated hydrocarbons, suggesting a specialized character of Alcanivorax strains to degrade these compounds (Head et al., 2006). Given the relevance that A. borkumensis got in the last years, the genome of this bacterium has been sequenced and annotated (Schneiker et al., 2006), and recent studies have shown the capacity of this strain for adaptation to different environmental stress factors such as UV-radiation, low temperature, or osmotic up-shift (Sabirova et al., 2008). Although genes coding for enzymes which are required for biosynthesis of bacterial storage compounds like polyhydroxyalkanoates (PHAs), triacylglycerols (TAGs), or wax esters (WEs), are present in the genome of A. borkumensis (SK2), only TAGs and WEs seem to be the relevant carbon storage compounds produced by this strain (Kalscheuer et al., 2007).

Figure 1: A. borkumensis

5 BIOAUGMENTATION AND BIOSTIMULATION

Bioremediation, involving bioaugmentation and/or biostimulation, being an economical and eco-friendly approach, has emerged as the most advantageous soil and water clean-up technique for contaminated sites containing heavy metals and/or organic pollutants. Addition of pre-grown microbial cultures to enhance the degradation of unwanted compounds (bioaugmentation) and/or injection of nutrients and other supplementary components to the native microbial population to induce propagation at a hastened rate (biostimulation), are the most common approaches for in situ bioremediation of accidental spills and chronically contaminated sites worldwide. However, many factors as we will see, may lead to their failure (Tyagi et al., 2011).

To improve the bioremediation process, besides a competent microbe able to degrade the contaminant carbon source, other parameters must be taken into account e.g. water, oxygen, and utilizable nitrogen and phosphorous sources (Rosenberg et al., 1992). Lack of any of the mentioned parameters makes the remediation process under natural conditions inefficient (Tyagi et al., 2011).

Bioaugmentation or seeding is the addition of highly concentrated and specialized populations (single strains or consortia) to the site contaminated with recalcitrant toxic compounds (Leahy and Colwell, 1990; Gentry et al., 2004). This technique is best suited for sites that (i) do not have sufficient microbial cells or (ii) the native population does not possess the metabolic routes necessary to metabolize the compounds under concern. On the other hand, biostimulation involves the identification and adjustment of factors such as nutrients that may be looming the biodegradation rate of the contaminants by the indigenous microorganism at the affected site (Swannell et al., 1996). Besides the type and concentration of nutrients, physical and environmental parameters also influence the mineralization rate of hydrocarbons by degrading bacteria. These factors include the chemical composition, physical state and concentration of the crude oil or hydrocarbons; along with the temperature, oxygen availability, salinity, pressure, water activity and pH on the site (Leahy and Colwell, 1990).

The decision to implement either or both of these techniques for bioremediation largely depends on the degrading capability of the indigenous microbes and the extent of contamination of the site to be treated. As several examples show, successful laboratory studies concerning bioremediation do not necessarily lead to reproducible in situ decontamination (Balba et al., 1998; Cassidy et al., 1996; El Fantroussi and Agathos, 2005). This impending gap between laboratorial trials and on-field studies may be due to several factors influencing the remediation process. Among these are strain selection, indigenous microbial ecology, type of contaminants, environmental constraints, and the procedures used for the introduction of the remediation agents (Tyagi et al., 2011).

5.1 Bioaugmentation

A successful bioremediation program usually requires the application of strategies customized for the specific environmental conditions of the contaminated site. The most commonly used options for bioaugmentation are: addition of a pre-adapted pure bacterial strain; addition of a pre-adapted consortium; introduction of genetically engineered bacteria; and addition of biodegradation relevant genes packaged in a vector to be transferred by conjugation into indigenous microorganisms (El Fantroussi and Agathos, 2005).

The best approach for selecting competent microbes should be based on the prior knowledge of the microbial communities inhabiting the target site (Thompson et al., 2005; van der Gast et al., 2004). In the case of co-contaminated sites, e.g. contaminated with both high metal concentrations and organic pollutants, the microbial population ability to degrade the organic compounds may be inhibited by the co-contaminants (Roane et al., 2001). The proposed strategies, in such cases, have involved the use of multi-component systems such as a microbial consortium, which is a better representation of a real environment than models based on single-component systems (Ledin, 2000).

From an applied perspective, using a microbial consortium rather than a pure culture for the bioremediation is more advantageous as it provides the metabolic diversity and robustness needed for field applications (Rahman et al., 2002; Nyer et al., 2002). Alisi et

al. (2009), successfully obtained complete degradation of diesel oil and phenanthrene; a reduction of 60% of isoprenoids; and an overall reduction of about 75% of the total hydrocarbons in 42 days, using a microbial formula made with selected native strains (Alisi et al., 2009). Albeit, adding microbial consortia with five fungi (Phanerochaete chrysosporium, Cuuninghamella sp., Alternaria alternate (Fr.) Keissler, Penicillium chrysogenum, and Aspergillus niger) and three bacteria (Bacillus sp., Zoogloea sp., and Flavobacterium) enhanced the degradation rate significantly (41.3%) (Li et al., 2009).

The amount of biomass to be used as inoculum for bioaugmentation is produced in bioreactors, although the transference of such cultures to the site is often critical. Microbial inoculants are homogeneous cell suspensions produced under optimum conditions which often undergo stress when they contact with the complexity of the natural habitats. In real cases, the introduced population starts decreasing shortly after being added due to several abiotic and biotic stresses. The stresses that difficult microbial growth may include fluctuations or extremes in temperature, water content, pH, depletion of nutrients, and also potentially toxic pollutant levels in the contaminated soil (Gentry et al., 2004).

The suggested possible reasons for bioaugmentation failure were: problems concerning the adaptation of the inoculated microorganisms; insufficiency of substrate; competition between introduced and indigenous biomass; use of other organic substrates in preference to the pollutant; and predation (Goldstein et al., 1985).

The use of carrier materials often provides a physical support for biomass, along with a better access to nutrients, moisture and aeration, which extends the survival rate of the microbes (Mishra et al., 2001). Microbial cell encapsulation or immobilization may offer a better survival rate protecting cells under stressed environmental conditions, usually enabling a faster and more efficient biodegradation as compared to free living cells (Moslemy et al., 2002; Obuekwe and Al-Muttawa, 2001). Encapsulation controls the flow of nutrients, lowers the concentration of toxic compounds in the microenvironment of the cells, minimizes cell membrane damage as it reduces the exposure to the toxic compounds and protects from predation and competition; thereby mimicking a miniature bioreactor in the environment (McLoughlin, 1994). Several materials like agar, agarose, alginate,

gelatin, gum, acrylate copolymers, and polyurethane, have been well studied and tested to encapsulate or immobilize cells (Cassidy et al., 1996). Moslemy et al. (2002) encapsulated an enriched bacterial consortium (isolated from a gasoline-polluted site) in gellan gum microbeads. Encapsulated cells exhibited a shorter lag phase and thus a higher gasoline degradation rate as compared to their free cell counterparts at equivalent microbial concentrations (Moslemy et al., 2002). In a recent study, Liu et al. (2009) compared the biodegradation of phenol by free and immobilized cells of Acinetobacter sp. and Sphingomonas sp. strains collected from activated sludge and phenolcontaminated soil. They found that the mixture of the two strains performed better than the pure cultures; and immobilized cells performed better in what concerned the degradation of phenol at concentrations higher than 500 mg/l and could be used for 20 cycles (Liu et al., 2009).

5.1.1 Bacteria consortia

The success of bioremediation technologies applied to hydrocarbon-polluted environments highly depend on the biodegrading capabilities of native microbial populations or exogenous microorganisms used as inoculants (Venosa and Zhu, 2003). The origin of microorganisms used in bioremediation is a controversial subject. One hand native populations present in contaminated sites are certainly adapted to the climatic, physico- chemical and nutrient conditions, but on the other hand these communities frequently do not include species with the enzymatic abilities needed to allow bioremediation proceed at increased rates, resulting in long time processes (Díaz-Ramírez et al., 2008).

Bioremediation of complex hydrocarbons usually requires the cooperation of more than a single species because the individual microorganism can metabolize only a limited range of hydrocarbon substrates (Mukherjee and Bordoloi, 2011). Possible sources to obtain these microorganisms are: remediated or contaminated sites, commercial suppliers and genetic engineering (Díaz-Ramírez et al., 2008).

Therefore, assemblages of mixed populations with overall broad enzymatic capabilities are required to bring the rate and extent of petroleum hydrocarbon degradation much faster (Wiesel et al., 1993). Some members of the microbial community might be able to secrete important degradative enzymes, growth factors, whereas others may produce

biosurfactants leading to the enhanced solubilization of hydrophobic hydrocarbons for their better utilization by microbes (Ghazali et al., 2004; Mukherjee and Das, 2005).

5.1.1.1 Bioremediation and reclamation of soil contaminated with petroleum oil hydrocarbons by exogenously seeded bacterial consortium: a pilot-scale study

During the last decade, several reports on the microbial degradation of petroleum oil hydrocarbons in soil were published; however, limited studies have been attempted to investigate the ecotoxicity and reclamation of bioremediated soil. Since the major objective of the bioremediation of petroleum oil-contaminated soil should be to reclaim the soil, enabling it to support the growth of crop plants, the next example reinforces the application of bacterial consortium rather than individual bacterium for the effective bioremediation as well as by other bacterial consortia. Furthermore, it has demonstrated that bioremediated soil can support the germination and growth of crop plants, thus vouching for the efficiency of this method (Mukherjee and Bordoloi, 2011).

It has been suggested that synergistic interaction among the members of bacteria in a community is one of the essential factors for achieving successful bioremediation of soil (Ghazali et al., 2004; Kim et al., 2005).

Different bacteria were selected based on their biosurfactant production and hydrocarbon degradation property. For the laboratory-scale bioremediation experiment, the bacterial consortium was formulated by mixing equal proportions of pure cultures; it consists of Bacillus subtilis (DM-04) and Pseudomonas aeruginosa (M and NM strains). It is worthy to mention that these bacteria originally isolated from a petroleum oil-contaminated soil were shown to be efficient degraders of petroleum oil components both in liquid culture as well as in soil (Mukherjee and Das, 2005; Das and Mukherjee, 2007).

The soil samples were biostimulated by adding nitrogen and phosphorus sources, respectively, to give an initial C/N/P ratio of approximately 100:10:1 (Vezquez et al., 2009). The experiment was carried out in 180 days of treatment, as we can see in table 1

along with the level of contamination of the different petroleum compounds. It is to be noted that a supplementation of glucose as a co-carbon source enhanced the rate of biodegradation of policiclic aromatic hydrocarbons (PAH) by the bacterial strains (Das and Mukherjee, 2007). A control was run in parallel where the soil was treated with un- inoculated medium, in order to be able to compare afterwards.

The amount of residual residual total petroleum hydrocarbons (TPH) was determined gravimetrically. TPH was fractionated into alkane, aromatic, asphaltene, and NSO- containing fractions on a silica gel column (Das and Mukherjee, 2007). The ecotoxicity of the elutriates obtained from petroleum oil contaminated soil before and post-treatment with microbial consortium at different time intervals was tested on germination of Bengal gram (Cicer aretinum) and green gram (Phaseolus mungo) seeds (Petukhov et al., 2000).

The water holding capacity of the soil increased from 45% to 70% at the end of the experiment. The alkane fraction, which is the largest constituent of TPH, was detected as 56% of total TPH followed by aromatic fraction (20%), asphaltene fraction (16%), and then NSO (5%) fraction. The biodegradation study showed that the TPH level of contaminated soil was reduced from 179.5 to 43 gkgˉ ˡ (76% degradation). The in situ degradation of TPH (alkane, aromatic, NSO, and asphaltene fractions) by indigenous microbes present in petroleum oil-contaminated soil (control soil) was insignificant compared to the biodegradation of the same fractions by applied bacterial consortium (Mukherjee and Bordoloi, 2011).

Table 1: Biodegradation of crude petroleum oil in soil by bacterial consortium.

The GC analysis confirmed that bacterial consortium was more effective in degrading the alkane fraction compared to the aromatic fraction of crude petroleum oil hydrocarbons in soil. It may be inferred that n-alkanes, particularly with a hydrocarbon chain length in between C12 and C21, are the most suitable substrates for microbial attack owing to their low hydrophobicity (Ghazali et al., 2004). Furthermore, bioavailability is the key deciding factor for the biodegradation of any component of crude petroleum oil. Therefore, it may be presumed that the lower degradation of short-chain alkanes (C8–C11) is related to their less availability due to the easy penetrability in soil (Hornick et al., 1983) as well as their toxicity to bacterial cells (Wang et al., 2002).

Benzene, toluene, and xylene (BTX) degradation in soil showed that added bacterial consortium preferentially degraded the benzene (98%) followed by toluene (96%), whereas xylene (76%) was least degraded. The BTX degradation potential of the formulated bacterial consortium was significantly higher as compared to the rate of biodegradation of these compounds by individual isolate of the same consortium and also better than some previously reported values on the degradation of BTX with individual isolates (Jean et al., 2008; Wang et al., 2008). In the bioremediated soil, the level of N and K was found to be higher than the normal soil. Phosphorous level was found to be more or less similar to that of normal soil. There was no difference in pH of the bioremediated and normal soils (Mukherjee and Bordoloi, 2011).

Table 2: Nutrient (NPK) and pH status of the bioremediated and normal soils

The crop plants showed better growth performance in bioremediated soil as compared to normal soil, which is showed in table 2. This might be due to the increased levels of nutrients in bioremediated soil as depicted by higher levels of nitrogen, potash, and phosphate. It suggested that the bioremediation of petroleum oil-contaminated soil by seeded bacterial consortium effectively restored the conditions suitable for the germination, growth, and development of crop plants (Mukherjee and Bordoloi, 2011).

Since the bioremediation process is an eco-friendly and cost-effective alternative treatment, therefore field-scale experiment using bacterial consortium used in the present study may be recommended for the effective bioremediation of petroleum oil-contaminated soil (Mukherjee and Bordoloi, 2011).

5.1.1.2 Role of a moderately halophilic bacterial consortium in the biodegradation of polyaromatic hydrocarbons Polycyclic aromatic hydrocarbons are ubiquitous pollutants in the environment, and most high molecular weight PAHs cause mutagenic, teratogenic and potentially carcinogenic effects. Since aromatic hydrocarbons are toxic it is necessary to completely degrade them to harmless end products such as CO2 and water (Arulazhagan and Vasudevan, 2009).

The hydrophobic nature of polyaromatic hydrocarbons makes their clean-up extremely difficult as they persist for a long period of time. In addition to increasing environmental persistence with increasing molecular size, evidence suggests that in some cases PAH genotoxicity also increases with size, up to at least four or five fused benzene rings (Cerniglia, 1992).

Low molecular weight (LMW) PAHs, such as anthracene, phenanthrene, and naphthalene, are easily biodegradable under laboratory conditions, since they are readily available in water for biodegradation, but high molecular weight (HMW) PAHs, such as benzo(a)pyrene and benzo(a)anthracene, show extensive resistance to degradation (Kanaly and Harayama, 2000).

The present study was focused on the degradation of PAHs in a marine environment by a moderately halophilic bacterial consortium. The bacterial consortium was isolated from a mixture of marine water samples collected from seven different sites in Chennai, India with different salt concentrations. The PAHs were selected for the study based on their number of rings, they are: the low molecular weight (LMW) PAHs phenanthrene and fluorine, and the high molecular weight (HMW) PAHs pyrene and benzo(e)pyrene. The consortium was able to metabolize both LMW and HMW PAHs and also the ones present in crude oil-contaminated saline wastewater. The bacterial consortium was able to degrade 80% of HMW PAHs and 100% of LMW PAHs in the saline wastewater. The strains presented in the consortium were identified as Ochrobactrum sp., Enterobacter cloacae and Stenotrophomonas maltophilia. The bacterial consortium grown on LMW-PAHs (phenanthrene/ fluorene), where phenanthrene was the sole carbon source, showed better growth. The consortium degraded phenanthrene (3 mg/l) almost completely (99%) by releasing CO2 (approximately 97%) (Arulazhagan and Vasudevan, 2009).

An increase in CO2 evolution (example with phenanthrene in figure 2) with a corresponding increase in the percent degradation showed that the consortium has the potential to degrade phenanthrene to harmless end products. The ratio between the amount of carbon dioxide produced and residual hydrocarbons provides all the information on hydrocarbon degradation (Penet et al., 2004). Metabolites were formed during phenanthrene and fluorene degradation, the phthalic acid formed in both processes may be further degraded to CO2 and water. When the bacterial consortium was grown on a PAH- coated mineral salts agar medium, clearing zones were visualized around the conso rtium, indicating PAH utilization (Arulazhagan and Vasudevan, 2009).

Figure 2: CO2 Evolution during degradation of phenanthrene (NaCl: 30 g/l)

Growth of a bacterial consortium on the mineral salts medium (MSM) containing LMW PAHs as the sole source of carbon and energy was demonstrated by an increase in bacterial cell number and concomitant PAH degradation (Arulazhagan and Vasudevan, 2009).

Degradation of LMW PAHs (phenanthrene and fluorene): Phenanthrene (a LMW PAH) was almost completely degraded (>95%) by the bacterial consortium in 4 days at concentrations of 5, 10, and 20 ppm as we can observe in figure 3. When the concentration of phenanthrene increased to 50 ppm, the rate of degradation reduced to 89%, and at 100 ppm degradation was 74%. In the case of fluorene (figure 4), similar results were obtained. The reason for a decrease in the degradation from the bacterial consortium may be due to the increase in the concentration of PAHs, which leads to a toxic effect on the bacterial consortium. Accumulation of toxic metabolites in the degradation process also leads to a decrease in the degradation rate (Vidali, 2001).

The consortium in the present study degraded 89% of phenanthrene (50 mg lˉ ˡ ) in 4 days at 30 g/l salinity. Thus, when compared to the previous studies the consortium degraded phenanthrene at a faster rate, without any additional carbon source, probably due to an active potential enzymatic profile. This is evident from the percent degradation, CO2 evolution study and the metabolites formed during the degradation process (Arulazhagan and Vasudevan, 2009).

Figure 3: Degradation of different concentrations phenanthrene

Figure 4: Degradation of different concentrations fluorene

Degradation of HMW PAHs (pyrene and benzo(e)pyrene): When HMW PAHs such as pyrene (figure 5) and benzo(e)pyrene (figure 6) were the sole carbon source, at a salinity of 30 g/L the bacterial consortium degraded 89% and 88% of pyrene at 5 and 10 ppm, respectively, in 4 days. At 20 ppm, 81% of pyrene was degraded, but at 50 and 100 ppm the rate of degradation was 74% and 52%, respectively. This low rate may be due to its hydrophobity. However, the consortium was able to degrade 70% of benzo(e)pyrene at 1 ppm. The rate of degradation decreased to 51% at 2 ppm and 21%

degradation was achieved at 5 ppm. As the benzene ring number increased the rate of degradation decreased (Arulazhagan and Vasudevan, 2009). Most of the studies showed that bacterial cells require an additional substrate for the degradation of HMW PAHs (Sohn et al., 2004).

The consortium was able to degrade only 21% of 5 mg lˉ ˡ benzo(e)pyrene in MSM but significantly degraded 80% of it in the crude oil-contaminated saline wastewater (Arulazhagan and Vasudevan, 2009).

Figure 5: Degradation of different concentrations pyrene

Figure 6: Degradation of different concentrations benzo(e)pyrene

The capacity of the bacterial consortium to metabolize low and high molecular weight PAHs at different concentrations in MSM and the degradation of PAHs present in crude oil-contaminated saline wastewater was demonstrated (Arulazhagan and Vasudevan, 2009).

5.1.1.3 Design of bacterial defined mixed cultures for biodegradation of specific crude oil fractions using population dynamics analysis by Denaturing Gradient Gel Electrophoresis (DGGE) Hydrocarbon-degrading mixed cultures could be prepared by combining a number of microorganisms to establish a defined mixed culture with enhanced biodegrading capabilities. Selection of strains forming a defined mixed culture (DMC) should be based on ability to efficiently metabolize petroleum fractions (Díaz-Ramírez et al., 2008).

Molecular ecology tools such as TGGE/DGGE have been used for screening of environmental dominant isolates (Ramírez-Saad et al., 2000; Viñas et al., 2005). These methods could be also useful to evaluate the predominance of microbial populations during hydrocarbon biodegradation process (Torsvik et al., 1998). It is important to recognize the relative limited scope of the DGGE approach to correlate presence of the different strains and their metabolic activity. However, this community profiling approach has proven effective for the selection and preparation of defined mixed cultures. Even though, in this work we started from oil-degrading strains susceptible of plating, DGGE resulted faster and more straightforward approach, avoiding the biases of plating, discriminating (based in colony morphology) and counting different strains in a mixed culture plate (Díaz-Ramírez et al., 2008).

The aim of this experiment was to design three defined mixed cultures (DMC) for the biodegradation of Maya crude oil specific fractions. The design was based on both; the biodegrading efficiency and the dynamics of the involved bacterial strains during growth in the presence of petroleum hydrocarbons (Díaz-Ramírez et al., 2008).

Hydrocarbon-degrading bacteria originally isolated from the rhizosphere of Cyperus laxus (Lam) a plant naturally growing on oil-polluted soils, were used to conform the defined mixed culture A (DMC A), made up with ten strains. The strains have been previously assayed as degraders of at least one petroleum fraction and TH. Strains identification was carried out by Microbial ID (Newark, Del) on the basis of cellular fatty acid analysis (Díaz-Ramírez et al., 2003).

From this ten-strain defined mixed culture (DMC A), we have designed DMC B (six strains) and DMC C (three strains) by selecting the predominant strains after degradation assays of specific petroleum fractions. Selection was based on PCR-DGGE profiles that were obtained for each bacterial strain and the DMCs (Díaz-Ramírez et al., 2008).

The DGGE profiles of each assay were obtained at different times, and analyzed to evaluate the population dynamics on each Maya crude oil fraction. The use of DGGE- fingerprinting to track microbial populations, allowed selecting strains to design efficient oil-degrading defined mixed cultures. These ones were made by blending standardized suspensions of each strain (Díaz-Ramírez et al., 2008).

Selection of bacterial strains was based on their hydrocarbon biodegrading abilities and relative abundance after biodegradation assays, these, were performed in a mineral base medium (MBM). The MBM was added with either; aliphatics, a mixture of aromatic–polar fractions or total hydrocarbons (TH), as carbon source. In order to maintain the C/N ratio value around 20 the nitrogen source concentration was adjusted according to the initial concentration of the corresponding carbon source (Díaz-Ramírez et al., 2008).

The hydrocarbon fractions were obtained from Maya crude oil. Briefly, volatile oil compounds were eliminated by heating (5 days, 70 ºC) and asphaltenes were extra cted by n-pentane precipitation. The remaining oil is hereafter referred as TH. TH was further fractionated by column chromatography to obtain the aliphatic, aromatic and polar fractions. The aromatic and polar fractions were obtained separately and then mixed, to yield a hydrocarbons mixture of the more recalcitrant compounds present in TH (Díaz- Ramírez et al., 2008).

Biodegradation of the aliphatic fraction (10 000 mglˉ ˡ ) and an aromatic–polar mixture (5000 mg lˉ ˡ ) was evaluated for the DMC B, indicating high biotransformation of the aliphatics. Biodegradation of total hydrocarbons (10 000 mglˉ ˡ ) and its fractions was evaluated for DMC B and DMC C (Díaz-Ramírez et al., 2008).

There are O2 consumption and CO2 production that are assessed by respirometry. Biodegradation end-point was established when oxygen consumption rate values were near to zero. The higher CO2 production and O2 consumption period was reached during the first 10 days of the assay. This high oxidation period correlates with the behavior of the residual hydrocarbons concentration throughout the biodegradation assay, where the larger aromatic–polar mixture degradation occurs during the first 13 days of culture by the DMC B (Díaz-Ramírez et al., 2008).

Interestingly, aliphatic fraction was completely degraded by the DMC B after 18 days of culture as is is showed in figure 8. The composition of residual aliphatics throughout biodegradation was determined by gas chromatography analysis and it is observed in figure 7 (Díaz-Ramírez et al., 2008).

Figure 7: Gas chromatography analysis for residual aliphatics through biodegradation (DMCB)

The DMC B was able to degrade the high concentration of TH obtained from Maya crude oil. As expected, the higher biodegradation values were registered for the aliphatic fraction, followed by the aromatics and polars. The biodegradation rates of aliphatic and aromatic fractions were 1.6 fold higher than those previously obtained with the DMC A. In the case of the polar fraction, DMC B was able to degrade about 19%, while, DMC A was previously found to be unable to degrade this fraction (Díaz-Ramríez et al., 2003).

Figure 8: Residual concentrations in independent assays (DMC B)

The achieved total biodegradation values with the DMC C were lower than those obtained with the DMC B. Regarding the biodegradation of TH fractions by the DMC C, again the aliphatic fraction was consumed in higher extent than the aromatic and polar fractions. Comparing these results with the respective values from the other two DMCs tested, the TH-aliphatics biodegradation extent (BE) was roughly similar to those obtained with DMC A and B. Statistical comparison of TH-aliphatics biodegradation values from DMC B and DMC C showed significant differences at BE level, while HBR were not considered different. In contrast, the BE values of the aromatic and polar fractions obtained with the DMC C were significantly lower than the values reached with DMC B. The same behavior was registered for the HBR, showing that the biodegrading capacity of DMC C was not improved by the reduction to three strains from six (DMC B). When DMC C was cultured for 18 days, the final biodegradation of the aromatics and polars was around 39.0% and 14.2%, respectively. These values were comparable to those obtained with the DMC B, however, biodegradation in that case was achieved in 11 days of culture (Díaz-Ramírez et al., 2008).

The most active strains in the biodegradation of the different fractions tested with DMC B belong to the following genera: Pseudomonas, Gordonia (Gr 3A) and Bacillus. Strains

belonging to the genus Pseudomonas are the most commonly isolated microorganisms in oil-contaminated sites; many of them demonstrated hydrocarbon-degrading activity in vitro (Chaineau et al., 1999; Barathi and Vasudevan, 2001).

In contrast with these results, the reduction in DMC C to three bacterial strains did not improve the biodegrading ability of the mixed culture. Although the biodegradation ability of DMC C was not significantly improved, the bacterial strains conforming this culture were able to partially degrade the aromatic and polar fraction present in the TH (Díaz- Ramírez et al., 2008).

In our case, better results were obtained with DMC B, where three Bacillus strains (Bc 1A, Bs 7A and Bc 9A) were present, when compared to DMC C, which has no members of these genera. According to the DGGE profiles these Bacillus strains only had presence during the first 10 days of culture with the different hydrocarbons (Díaz-Ramírez et al., 2008).

We developed a strategy to design highly efficient defined mixed cultures whose application as inoculants, in bioremediation of hydrocarbon-contaminated tropical soils, could represent an alternative with high potential benefits (Díaz-Ramírez et al., 2008).

5.2 Biostimulation

Spilled petroleum hydrocarbons represent a substantial C-source for the indigenous microorganisms, whereas, in most environments, the presence of nitrogen and phosphorous is limited. Thus, biostimulation accelerates the decontamination rate, as the addition of one or more rate-limiting nutrients to the system improves the degradation potential of the inhabiting microbial population (Nikolopoulou and Kalogerakis, 2009; Prince, 1997).

Sarkar et al. (2005) enhanced the biodegradation of petroleum hydrocarbons by up to 96% after the addition of biosolids (nutrient-rich organic matter resulting from the treatment of domestic sewage) and inorganic fertilizers (rich in N and P) to diesel contaminated soils (Sarkar et al., 2005). Similar results were obtained for diesel oil remediation in the Antarctic coastal sea by biostimulation using a commercial fertilizer (Delille et al., 2009).

In the case of marine environments, the addition of biostimulants is more critical, as firstly they should target the microbes near the oil droplets in the seawater and secondly they should not readily get diluted and washed out by the wave action. Moreover, higher concentrations of N and P sources can cause eutrification, thereby enhancing algal growth and ultimately reduce the dissolved oxygen concentration in the water (Nikolopoulou and Kalogerakis, 2009).

Besides nutrients there are several other factors that largely determine the PHCs degradation rate under natural conditions. For instance, it was observed that factors including the intensity of physical mixing, the pre-treatments (for example shore washing, manual removal, manual raking, bioremediation using biostimulation agents, mechanical tilling, mechanical relocation), and the availability of alternative carbon sources influenced the mineralization potential of the microbes after the Exxon Valdez oil spill (Sugai et al., 1997). Another parameter that has a considerable effect on biodegradation is temperature, due to its effect on the viscosity, water solubility and chemical composition of the oil. It also influences the rate of hydrocarbon metabolism and the composition of the microbial community (Atlas, 1981).

Mulkins-Phillips and Stewart (1974) studied the effect of temperature (ranging from 5 to 28ºC) on the degradation of bunker C fuel oil following the spillage at Chedabucto Bay, using enriched mixed microbial cultures. After incubation for 7 days at 15ºC, 41–85% of benzenesoluble components disappeared, whereas, 21–52% degradation was obtained after 14 days of incubation at 5ºC (Mulkins-Phillips and Stewart, 1974). Horel and Schiewer (2009) studied the impact of temperature and moisture on biostimulation of syntroleum (synthetic diesel oil) using fertilizer as stimulating agent. The bioremediation process started much earlier at a higher than at a lower temperature (20 and 6ºC, respectively), for short incubation periods (4–6 weeks). Nevertheless, microbes adjusted to the lower temperature during long incubation periods (12–17 weeks) and degraded up to 50% syntroleum. Nutrient supply was found to be essential for microbial degradation under all conditions in the soil. The degradation of the contaminant after 17 weeks was almost 3 times higher at 20ºC and 8 times higher at 6ºC when compared to nutrient-deficient sands.

However, moisture content (2–12% by weight) and regular mixing to enhance soil porosity did not influence degradation significantly (Horel and Schiewer, 2009). On the contrary, it was found that temperature showed limited influence on petroleum degradation for the Antarctic seawater samples in a laboratory microcosm study, where commercial fertilizer enhanced bioremediation (Delille et al., 2009).

Inadequate bioavailability of the hydrocarbons to microorganisms due to low water solubility has been addressed as a limiting step in biodegradation. The use of biosurfactants to enhance the biodegradation rate has been well studied (Rahman et al., 2002; Bordoloi and Konwar, 2009; Ron and Rosenberg, 2002). Nikolopoulou and Kalogerakis showed that biostimulation using a N and P fertilizer together with biosurfactants enabled naturally occurring microbes to adapt better and faster to the oil spill contamination, ensuring a relatively shorter lag phase and faster degradation rates (Nikolopoulou and Kalogerakis, 2008). Apparently, the combination of bioaugmentation, biostimulation, and biosurfactant addition, depending on the characteristics of the contaminated site, might be a promising strategy to speed up bioremediation (Baek et al., 2007). However, any planned intervention must be followed by ecotoxicity and quality studies of the contaminated site to ascertain that it has regained its natural biological activity and integrity (Hamdi et al., 2007; Liu et al., 2010). These toxicity tests and biological activity measurements should be used as monitoring tools or bioindicators during and after bioremediation of contaminated sites.

5.2.1 Using guano as fertilizer

The biodegradation of hydrocarbon pollutants in open systems, such as oceans, is generally limited by the availability of utilizable nitrogen and phosphorus sources. In considering the bioremediation of petroleum, it is useful to distinguish between open systems, e.g., oceans, and closed systems, e.g., reactors. In closed systems, it is simple to satisfy the necessary nitrogen and phosphorus requirements for efficient microbial growth and hydrocarbon degradation by the addition of water-soluble inorganic compounds (Knezevich et al., 2006).

An early example of a controlled experiment demonstrating enhanced petroleum bioremediation in a closed seawater system was the cleaning of a cargo compartment of an

oil tanker by addition of urea and phosphate (Rosenberg et al., 1974; Gutnick and Rosenberg, 1977). However, there is at present no practical microbial solution to the reoccurring problem of petroleum pollution in the sea, because of the difficulty of providing the necessary nitrogen (N) and phosphorus (P) supplements. Addition of common fertilizers containing nitrogen and phosphorus is not effective in open systems because of rapid dilution (Knezevich et al., 2006).

Uric acid is the major nitrogen waste product of birds, terrestrial reptiles, and many insects. It has a low solubility in water and adheres to hydrocarbons (Koren et al., 2003). These properties, in addition to the fact that uric acid is the major component of inexpensive and commercially available guano fertilizer, suggested to us that it might be a useful nitrogen source for bioremediation of petroleum pollutants in open systems. Many different species of bacteria are known to degrade uric acid (Vogels and van der Drift, 1976) and one species, Bacillus fastidiosus, was reported to grow only on uric acid and allantoin (Bongaerts and Voegls, 1976).

Using a simulated open system in one experiment, the first set utilized uric acid (the major nitrogen component of Guano) and a pure culture of the isolated hydrocarbon-degrading bacterium Alcanivorax sp. (OK2) to establish the principle that uric acid binds to crude oil and is available for bacterial growth and petroleum degradation (Knezevich et al., 2006).

Since uric acid has a density of 1.89, it rapidly sediments in seawater medium. However, when the seawater medium contained crude oil, the added uric acid did not settle, but remained on the surface bound to the oil. Clearly, in a true open system, such as the sea, the complex would be exposed to a much larger body of water and the uric acid would have to remain bound to the oil for at least a few days (Knezevich et al., 2006).

Some bacteria utilize uric acid as both a carbon and nitrogen source; others use it only as a nitrogen source. Strain OK2 probably fits into the latter group since it failed to grow on uric acid without a carbon source (Knezevich et al., 2006).

The second set of experiments utilized guano as a potential commercial source of nitrogen and phosphorus for the treatment of oil pollution in seawater. The guano used in the experiments contained 14% nitrogen (mainly as uric acid). In a simulated open system, using fresh seawater as the inoculum, cultures containing guano and crude oil reached over 10^8 cells per ml after 7 days, indicating that sufficient guano remained bound to the oil to serve as an efficient fertilizer (Knezevich et al., 2006).

Under the same conditions, seawater washed cultures containing ammonium sulfate and phosphate in place of guano, reached only 10^6 cells per ml . Data on oil degradation supported the conclusion that guano was a superior source of nitrogen and phosphorus than ammonium sulfate and phosphate in the simulated open system. No significant hydrocarbon degradation was detected when the salts were used in the wash-out experiments, whereas 70% degradation was measured when guano served as the fertilizer (Knezevich et al., 2006).

Most of the bacteria isolated after growth on crude oil and guano in seawater belonged to the γ –proteobacteria group, which is typically characteristic of bacterial communities associated with recently spilled oil (Kasai et al., 2001; Röling et al., 2002). Within the most abundant bacteria isolated were two strains from the Alcanivorax genus. This genus has been characterized as hydrocarbonoclastic Bacteria. One of the other strains isolated was an Altermonas species. This group of bacteria is widespread in marine environments (Eilers et al., 2000; Uphoff et al., 2001).

We demonstrate here that (1) uric acid is an effective nitrogen source for the degradation of petroleum in a similar seawater open system by an isolated Alcanivorax species; and (2) a commercial source of uric acid, guano, can serve as a nitrogen and phosphorus source for petroleum degradation by bacteria in a simulated open seawater system (Knezevich et al., 2006).

Results of oil extraction from the different experimental cultures, based on gravimetric analysis, showed high degrees of oil removal in the ―guano‖ cultures (Knezevich et al., 2006). In figure 9 we can compare the results of guano and no guano experiments.

Petroleum Analysis: Oil samples extracted from the ―guano‖ flasks after 3 days of growth were subjected to fractionation and gas chromatography in order to identify the components that were degraded. All of the compounds analyzed were extensively degraded during the incubation. The n-alkanes (C10, C12, and C14) were more than 95% degraded, whereas the longer chain n-alkanes (C16 toC36) were 63% to 87%degraded. The five aromatic hydrocarbons examined showed 56% to 81% degradation. It should be pointed out that the compounds analyzed represent only a small portion of the hydrocarbons present in crude oil (Knezevich et al., 2006).

Figure 9: Oil degradation using guano or ammonium sulfate as the nitrogen source

5.3 Bioaugmentation versus biostimulation

There is a debate on which of the two techniques, bioaugmentation or biostimulation, is a better strategy for bioremediation. Apparently, the circumstances prevailing at the site influence significantly the choice of the technique to be applied. Hamdi et al. (2007) found that selectivity and specialization of added microorganisms mainly defines the bioremediation efficacy, rather than the nutrient load (Hamdi et al., 2007). Bento et al.

(2005) compared bioremediation of diesel oil by natural attenuation, biostimulation and bioaugmentation. They concluded that the best approach for bioremediation of diesel oil was the bioaugmentation performed by inoculating microorganisms pre-selected from a contaminated site. Apparently, indigenous microbes (pre-selected for bioaugmentation) are more likely to survive and propagate when reintroduced into the site, as compared to transient or alien strains to such a habitat (Bento et al,. 2005; Thompson et al., 2005). On the contrary, there are reports of microcosm and field studies where inoculation with enriched cultures originating from the site itself did not affect hydrocarbon removal rates whereas stimulation was effective (Thomassin-Lacroix et al., 2002). Biostimulation can provide suitable nutrients and conditions for both indigenous and exogenous microbes. Thus, biostimulation becomes a viable approach in those cases where microbial population gets acclimatized due to exposure to hydrocarbons at contaminated sites. Natural acclimatization by the indigenous microbial population often requires a longer time due to an extended lag phase leading to prolonged bioremediation processes (Lendvay et al., 2003).

Due to limitations associated with bioaugmentation and biostimulation when applied individually, these techniques are emerging as complementary. Hamdi et al. amended aged PAH contaminated soil using both bioaugmentation and biostimulation and obtained higher PAH dissipation rates, especially for anthracene and pyrene, than those observed in unamended PAH-spiked soils (Hamdi et al., 2007).

5.4 Commercial products: ongoing scenario

Bioaugmentation and biostimulation may contribute to overcome a critical bottle neck of the bioremediation technology. The booming up of bioaugmentation and biostimulation technologies as preferred in situ remediation techniques has attracted commercial interest. Emerging formulations and products are gaining attention and application, claiming fast decontamination rates. However, these products are not panacea and need to be evaluated according to the requirements of the site before implementation. Dott et al. (1989) compared the biodegradability of fuel oil using the commercial product DBCTM (containing nine different dried bacterial mixed cultures) and a mixed culture from a domestic treatment plant during a 31-day incubation period. They asserted that the native

microorganisms possessed sufficient biodegradation capability due to eventual adaptations and therefore the rate and extent of biodegradation was higher than that of highly adapted commercial microbial cultures (Dott et al., 1989). Later in 1991, Venosa et al. conducted a detailed study for the screening of ten commercial products (eight bacterial cultures and 2 non-bacterial products) from ten different companies for the bioremediation of samples from Prince William Sound, Alaska. Their results suggested that the degradation of the hydrocarbons could be primarily ascribed to the activity of Alaskan microorganisms (Venosa et al., 1992).

The metabolic activity, adaptability and ecological competence of commercial inoculants appear to be a major limitation for a successful tailor made remediation. However, enriched cultures of selected indigenous microbes could produce a more appropriate and cost-effective product for local conditions than commercial bioremediation products (Mohammed et al., 2007)

5.5 Confined systems and real-case studies: bridging the gap

A successful process in the laboratory under controlled conditions does not imply similar success in an uncontrolled environment (Mueller et al., 1992). Indeed it is a pioneering step forward to address potential problems in the environment (Goldstein et al., 1985). Bioaugmentation and biostimulation studies comprising laboratory, simulated field and in situ tests, are very few so far.

The existence of competent microorganisms for a given bioremediation, their nutrient requirements (e.g. carbon, nitrogen, phosphorous, oxygen or an alternative electron acceptor) along with suitable growth conditions (temperature, redox potential, salinity, pH, etc.) should be firstly determined by laboratory and field trials (Rosenberg et al., 1992), more factors are explained in table 3. After optimizing conditions for bioremediation of crude oil using a combination of bioaugmentation and biostimulation technique in the laboratory, successfully implemented the same for field and in situ beach remediation (Rosenberg et al., 1992).

Based on these considerations, a deeper understanding of microbial degradation abilities, together with their metabolic networks as well as their cellular resistance and adaptation mechanisms, will bring out a variety of appropriate ‗‗microbial formula‘‘ tailored for a specific contamination site (Tyagi, 2011).

Table 3: Factors influencing bioaugmentation and biostimulation processes

Table 4: Bioaugmentation and biostimulation: field and real case studies

6 PRODUCTION OF LIPIDS AND STORAGE COMPOUNDS BY HCB BACTERIA

Crude oil pollution constitutes a temporary condition of carbon excess coupled to a limited availability of nitrogen that prompts marine oil-degrading bacteria to accumulate storage compounds. Storage lipid compounds such as polyhydroxyalkanoates (PHAs), triacylglycerols (TAGs), or wax esters (WEs) constitute the main accumulated lipophilic substances by bacteria under such unbalanced growth conditions. Biosynthesis and accumulation of storage lipid compounds seems to be common among hydrocarbonoclastic bacteria (Tyagi, 2011).

6.1 Production of lipids by hydrocarbonoclastic marine bacteria

The synthesis of at least one type of storage lipids (SL) is a common process among microorganisms. Many microorganisms synthesize lipophilic substances as an integral part of their metabolism and as storage compounds in response to a nutrient limitation in the environment. This process occurs if an essential element becomes unavailable and if there is an excess of the carbon source at the same time (Manilla-Pérez et al., 2010).

Members of most genera synthesize polymeric lipids such as PHAs (polyesters of hydroxy fatty acids) (Steinbüchel, 2001), whereas the accumulation of TAGs (trioxoesters of glycerol and long-chain fatty acids) or WEs (esters of primary long-chain fatty acids and primary long-chain fatty alcohols) is a property of a few prokaryotes (Wältermann and Steinbüchel, 2006), it has recently been reported in A. borkumensis (SK2) (Kalscheuer et al., 2007). They serve as depot for carbon and energy needed for maintenance and metabolic activity during starvation and particularly if growth resumes (Wältermann and Steinbüchel, 2005). These lipophilic compounds represent an ideal way of carbon storage, since they are compact, anhydrous, show a minor grade of oxidation, and posses a higher calorific value than proteins or carbohydrates (Lenz and Marchessault, 2005; Alvarez et al., 1996). Moreover, they represent low biological toxicity compared to free fatty or hydroxy fatty acids and are water-insoluble and osmotically inert (Wältermann and Steinbüchel, 2006).

The occurrence of storage lipids seems to be associated to the habitat of bacteria, because in species existing in nutrient-rich habitats, biosynthesis of storage lipids seems insignificant (Wältermann and Steinbüchel, 2006). On the contrary, biosynthesis and accumulation of storage lipids represent only one of the different strategies developed by microorganisms to survive in variable environments (Alvarez et al., 1997). Another strategy developed by HDB is the increase of the contact area between bacteria and water- insoluble hydrocarbons through emulsification of hydrocarbon. Therefore, production of bioemulsifiers is common among HDB (Manilla-Pérez et al., 2010).

6.2 Storage compound accumulation in hydrocarbonoclastic bacteria

For marine oil-degrading bacteria, oil pollution constitutes a temporary condition of carbon excess coupled to a limited availability of nitrogen that prompts bacteria to accumulate storage compounds (Sabirova et al., 2006). Due to their oligotrophic way of life and the sporadic availability of hydrocarbons as substrates for their growth, marine hydrocarbonoclastic bacteria are frequently facing extended phases of starvation conditions. Therefore, these bacteria must have developed adaptation and survival strategies under such unfavorable environmental conditions (Kalscheuer et al., 2007) such as the accumulation of intracellular carbon storage for starvation periods. Production of PHAs in Pseudomonas oleovorans, of TAGs in Rhodococcus opacus (strain PD630) and of WEs in Acinetobacter baylyi (strain ADP1) constitute good examples of hydrocarbon utilization to produce high-value products (Manilla-Pérez et al., 2010).

6.3 Production of biosurfactants by oil-degrading bacteria

Although the production of extracellular lipids has already been reported in bacteria such as Acinetobacter sp. (Singer et al., 1985), Alcaligenes sp. (PHY9), Pseudomonas náutica (strain 617) (Goutx et al., 1990), the physiological role of lipid export is unclear. As part of surface-active compounds also known as biosurfactants or bioemulsifiers, specialized lipids of diverse chemical structures could play a role in the modification of medium properties and could therefore enhance the contact between bacteria and water-insoluble hydrocarbons. For that reason, most HDB produce biosurfactants of diverse chemical properties and molecular size (Ron and Rosenberg, 2002). These surfactants disperse

hydrophobic compounds, thereby increasing their surface to enhance growth of cells (Desai and Banat, 1997; Rosenberg and Ron, 1999). Lipid moieties are present not only in low but also in high molecular weight biosurfactants like glycolipids, lipoproteins, or lipopolysaccharides (Ron and Rosenberg, 2002; Maneerat, 2005). Biosurfactants should not only be considered as prerequisite of microbial alkane degradation, but also as products of secondary metabolism (Hommel, 1990).

6.4 Production and export of lipids by Alcanivorax strains

The export mechanism of lipids by Alcanivorax strains remains unclear to date. It has been proposed that the glycolipids of A. borkumensis (SK2) might be precursors of biosurfactants or could play a role in the function of the outer cell membrane (Abraham et al., 1998). Species such as P. aeruginosa or Acinetobacter (RAG-1) can use their biosurfactants to regulate their cell-surface properties (Ron and Rosenberg 2002). Therefore, production and export of lipids by Alcanivorax may constitute a step in the synthesis of biosurfactants to increase the availability of hydrophobic substances and hence improving the probability of survival in habitats with presence of hydrophobic compounds. Alcanivorax strains accumulate TAGs and WEs as main storage compounds (Manilla- Pérez et al., 2010).

Two recent directions of using bacteria with long-chain alkane utilization capabilities for biotechnological applications have been discussed: (1) the bacterial production of high- value wax esters, and (2) the biodegradation of paraffin waxes from complex mixtures like crude oils (Wentzel et al., 2007).

6.5 Concluding remarks and future perspectives

―Lipid biotechnology‖ covers the microbial production and biotechnological transformation of lipids and lipid-soluble compounds (Schörken and Kempers, 2009). Storage lipids like TAGs and different fatty acid types are the main targets for biotechnological product development, while phospholipids, sphingolipids, glycolipids, sterols, carotenoids as well as other lipid-soluble compounds are utilized for the production

of bioactive molecules in cosmetics and pharmaceuticals or for nutrition (Schörken and Kempers, 2009).

Since TAGs and fatty acids are precursors of biodiesel (Li et al., 2008), their production has recently attracted considerable attention. Screening for bacteria able to produce high amounts of TAGs is the first and essential step in the search for oleaginous microorganisms (bacteria that accumulate more of 20% of their weight as lipid). TAG accumulation by A. borkumensis (SK2) of up to 23% of its cell dry weight has already been reported (Kalscheuer et al., 2007). However, additional studies are necessary to further increase the content of TAGs in A. borkumensis (SK2), before this strain can be considered as candidate for biotechnological production of neutral lipids (Manilla-Pérez et al., 2010).

The ability of OHCB to metabolize almost exclusively hydrocarbons can be used to produce storage compounds (like PHAs, TAGs, and/or WEs as bulk chemicals), while the extracellular deposition of these compounds could represent a major advantage concerning the downstream processes, if they are produced biotechnologically (Manilla-Pérez et al., 2010).

7 DISPERSANTS AND OIL MINERAL AGREGATES

7.1 Use of dispersants

Using chemical dispersants as an oil spill countermeasure is the most frequently employed clean-up method because such liquids can be readily applied to large oil spills and also this method is generally more cost-effective than physical remediation methods. On the other hand the exclusive property of these dispersants to make oil spills dispersed into water will enhance the biodegradability of crude oil due to the increased exposed surface of the spills to such agents. Nevertheless, the surfactant compositions are efficient in dispersing spilled oil products and many of the effective ones have the drawbacks of being toxic and/or not biodegradable (Baghbaderani et al., 2009). It is not a sustainable technique since it does not remove oil from the sea and just transfers oil on water layer and/or sediments (Zahed, 2011).

Several factors may affect hydrocarbon degradation and, in particular, the oil concentration is an important consideration in determining whether bioremediation is a viable option (Zahed, 2011)

The dispersant Corexit 9500® was formulated in 1992. This dispersant effectively extends the ―window of opportunity‖ because it is effective on crude with viscosities up to 20,000 cp and it is also lower in toxicity than other dispersants for most species (Zahed, 2011).

A comparison of bioremediation of crude oil and dispersed crude oil at different initial oil concentrations were carried out using indigenous microorganisms. Some experiments confirmed that the use of dispersant improved the rate of biodegradation in bioreactors simulating marine environments contaminated with crude oil (Zahed, 2011).

7.2 Oil–mineral interaction in particle image velocimetry (PIV)

The mechanism of the formation of oil–mineral-aggregates (OMA) is studied by particle image velocimetry (PIV) to evaluate factors that may influence its application as an oil spill countermeasure. Both stationary and moving oil droplet strategies are emplo yed. The movements of mineral particles near an oil droplet are captured and the interactions between mineral particles and oil droplets are visualized through the vector fields of mineral particles. The interaction between two different crude oils and three types of minerals were evaluated in this work. The results show that hydrophobicity plays an important role in the interaction between mineral particles and oil droplets. Hydrophobic minerals (modified Kaolin) tend to have more oil–mineral attraction than hydrophilic minerals (original Kaolin and diatomite) (Wang et al., 2011).

In this work, a new approach was developed using particle image velocimetry (PIV) to study the formation of OMAs in situ. Particle image velocimetry is a laser photograph system that is designed to track the movement of particles by detecting the velocity of the moving particles in a large view of field (Adrian, 1986; Adrian, 1991; Buchhave, 1992). Two successive photos captured by a charge coupled device (CCD) with a certain inte rval were compared to calculate the displacements of the particles. The feature of non- intrusively tracking particles, one of the prominent features of PIV, makes PIV a good tool in studying the interactions between mineral particles and oil droplets. The moving tracks of mineral particles can be captured by a two-dimensional PIV (Wang et al., 2011).

During the recent Gulf of Mexico oil spill response operations, enhanced dispersion by the addition of chemical dispersants has been given serious consideration. Chemical dispersants comprised of surfactants and solvents can reduce the interfacial tension between oil and water so as to significantly decrease oil droplet sizes to micron-scale under mixing energy conditions encountered in the field (Lessard and Demarco, 2000; Venosa et al., 2002; NRC, 2005). The dispersed oils were then transported into the water column where the oil is diluted into concentrations lower than the toxicity threshold to pelagic and benthic species. However, the toxicities of dispersant and chemically dispersed oil were still a concern, especially in nearshore areas where dilution capacity is less than open waters (Bayat et al., 2005). Therefore, other processes with minimal or no dispersant

addition are being considered. Enhanced oil–mineral fine aggregate (OMA) formation has been identified as an effective alternative technology for oil spill cleanup (Bragg and Owens, 1994; Bragg and Owens, 1995; Lee, 2003). The oil droplets trapped in OMAs were stabilized and prevented from re-coagulating, which led to the significant reduction of droplet size (Ajijolaiya, 2006). Smaller oil droplets with a larger surface area exhibit higher rates of oil biodegradation due to enhanced availability to nutrients and oxygen (Bragg and Owens, 1994; Ajijolaiya, 2006; Weise et al., 1999). Oil associated with OMA, that is stabilized by mineral fines, will also have less chance to adhere onto bedrock and facilities (Kepkay et al., 2002; Li et al., 2007; Owens, 1999; Owens and K. Lee, 2003). Furthermore, the efficacy of this enhanced dispersion process may be improved by the presence of oil dispersants (Li et al., 2007; Guyomarch et al., 2002; Ma et al., 2008).

Poirier and Thiel reported that Kaolin was the most efficient mineral among the hydrophilic minerals tested (Poirier and Thiel, 1941). Omotoso et al. studied the mechanisms of crude oil–mineral interactions and reported that hydrophilic minerals (Kaolin) have less oil–mineral interaction than the hydrophobic minerals (calcite) (Omotoso et al., 2002). Other studies showed that hydrophobic compounds in crude oil do not adsorb directly fine particles unless they have an organic coating (Muschenheim and Lee, 2002; Stoffyn-Egli and Lee, 2002). It was also reported that the polarity of the oils played an important role in forming OMAs (Guyomarch et al., 2002; Omotoso et al., 2002). Negatively charged particles have an affinity to polar compounds of the oil. Asphaltenes and resins are amphiphilic components in crude oils (Quintero et al., 2009). The polarities of crudes increase with increasing concentrations of resins and asphaltenes [Guyomarch et al., 2002; Goual and Firoozabadi, 2002). The interactions between oil and hydrophilic minerals decrease as the polar content of crudes increases (Omotoso et al., 2002).

The two crudes studied were: Alaska North Slope (ANS hereafter) and Medium South American (MESA hereafter). ANS has a lower % of asphaltenes compared to MESA (composition showed in table 5). The dispersant used in the paper was Corexit 9500. The interaction between oils and minerals can be reflected by the shape and movement of oil droplets. The three minerals were: Kaolin, modified Kaolin and diatomite (physical

properties showed in table 6). Oval shaped oil droplet can be observed when hydrophilic minerals (Kaolin and diatomite) are present in brine solution. As the mineral is replaced by hydrophobic one (modified Kaolin), the oil droplet becomes a sphere (Wang et al., 2011).

Table 5: Composition of oils

Table 6: Physical properties of minerals

Original Kaolin and modified Kaolin, have the same properties such as density, porosity, except hydrophobicity. This evidence supports our hypothesis that hydrophobicity can promote the affinity between oil droplets and mineral particles and enhance the interaction between oil droplets and mineral particles (Wang et al., 2011).

Dispersant plays an important role in modifying the interfacial forces between oil droplets and water. Without dispersant, oil droplets frequently show regular shape, as recorded in figure 10. Introduction of dispersant can result in a significant change to the shape of oil droplets. The dispersant (Corexit 9500) was mixed with ANS crude at a ratio of dispersant to oil 1:15 (Wang et al., 2011).

Figure 10: ANS oil droplet with different types of minerals

Figure 11: Mineral–oil interactions (green area= crude oil and red spots = minerals)

The irregular shaped oil droplets (figure 12 c–h) can always be observed but the sequence of their appearance is not fixed. Figure 12 i–j shows that one large oil droplet breaks up into smaller droplets (Wang et al., 2011).

Figure 12: Effect of dis persants

The ANS with dispersant maintained its hydrophobic property and oil droplets were formed as it is introduced to brine solution. As the oil droplets rise, hydrophilic heads of dispersant make contact with the salt water, thus the oil droplets become irregularly shaped and disperse into smaller oil droplets. A similar phenomenon was observed in previous work (Li et al., 2007; Gopalan and Katz, 2010). However, a distinct phenomenon was observed when dispersant was introduced to the brine solution instead of ANS crude. Surface agents create a surface film on the brine solution, which repels the oil. It is surprising to observe a thin line of oil when it was injected into the brine solution with no formation of oil droplets (figure 13). This indicates that the method of dispersant addition is of great importance in the dispersion of oil droplets (Wang et al., 2011).

Figure 13: Effect of dis persants

The vectors represent the movements of the Kaolin particles, close to the surface of the oil drop, the vectors shown in region b) of figure 14 essentially point towards the oil drop while the vectors in region a) of figure 14 point to the opposite direction. This suggests an opposite interaction between oil and Kaolin particles in the area close to the surface of the oil droplet. In the area relatively further from the oil droplet (region c of figure 14), it is seen that Kaolin particles essentially move downwards. This indicates that the interaction between oil and mineral particles becomes weak further from the surface of oil droplet. Since the oil spills in the sea would never stop moving, the moving oil droplet was used to investigate the oil–mineral interaction (Wang et al., 2011).

Figure 14: Vector field at the surface of an oil droplet of ANS

Mineral particles move fast in the center of the oil droplet (figure 15) and in the middle of the tail that is close to the head. This indicates that intensive interaction between the modified mineral particles and oil droplet occurred at these areas (Wang et al., 2011).

Comparing figure 15a–e (modified Kaolin and Alaska North slope oil) to figure 15 f–j

(Original Kaolin and Alaska North slope oil), one may note that the tail formed with original Kaolin is shorter than the one formed with modified Kaolin and original Kaolin

particles have lower movement velocity. This indicates that the modified Kaolin particles have a stronger interaction with oil droplets compared to the original Kaolin particles, which can enhance the formation of oil–mineral aggregation (Wang et al., 2011).

Figure 15: Vector field around a moving oil droplet at different positions

These results suggested that modified Kaolin has stronger interaction with Alaska North slope oil in terms of both intensity and duration. Again, this was validated by the fact that hydrophobic minerals play an important role in the formation of oil–mineral aggregates (Wang et al., 2011).

Each vector can be divided into a horizontal component and a vertical component. To eliminate the effects of gravity, only the horizontal component is considered. When it

points towards the oil droplet, this horizontal component is denoted as H(+) and the corresponding mineral particle is considered to move towards the oil droplet (Wang et al., 2011).

Figure 16: Vector field of modified Kaolin and a static oil droplet (MESA)

The physical–chemical properties of oil and minerals may play a role in this interaction. Although both Kaolin and diatomite are hydrophilic minerals, diatomite is more hydrophilic than Kaolin. This can be evidenced by their contact angles shown in Table 2 (Wang et al., 2011).

For both MESA and ANS, modified Kaolin displayed enhanced OMA formation with the crudes compared to the original Kaolin, where percentage of mineral particles moving towards the oil droplet reaches 71% and 90% for MESA and ANS, respectively, which being greater than 50% suggests that attractive forces dominate between mineral particles and oil droplet. This shows that hydrophobic property of minerals can improve the formation of oil–mineral aggregates (Wang et al., 2011).

Several key factors that influence the interaction have been investigated. The major conclusions are: (1) Dispersant plays an important role in the formation of oil–mineral aggregates. The way to introduce dispersant to crude oil can have significant impact on the oil–mineral

interaction. The results showed that oil drops exhibit irregular shapes when the crude was mixed with dispersant. It would be more difficult to form oil droplets if brine solution is mixed with dispersant (Wang et al., 2011). (2) Confocal microscopy images showed that hydrophobicity can promote the affinity between oil droplets and mineral particles and enhance the interaction between oil droplets and mineral particles (Wang et al., 2011).

Figure 17: Comparison of interaction of hydrophilic minerals with different oil types

Figure 18: Effect of hydrophobic property on oil–mineral interaction

(3) For the crudes studied in this work, MESA was more repulsive against hydrophilic minerals (both Kaolin and diatomite) than ANS. The hydrophilic property of minerals plays a role in the formation of OMA (Wang et al., 2011).

(4) The vector field of moving oil droplets shows that ANS interacts with modified Kaolin more intensively than with original Kaolin. Movement of minerals can enhance the formation of OMA (Wang et al., 2011).

8 COLD ENVIRONMENTS

Oil spills have become a serious problem in cold environments with the ever-increasing resource exploitation, transportation, storage, and accidental leakage of oil. During the past decade, hydrocarbon pollution increased in cold regions, such as northern Russia, Canada, Alaska, and the Qinghai-Tibet Plateau (Collins et al., 1993; Chuvilin et al., 2000; Jin et al., 2006). An average of 407 spills of oil products occurred annually in Alaska between 1996 and 1999 (Poland et al., 2003). In former Soviet Union, 19% of the total reported incidents occurred in permafrost areas (ESMAP 2003).

Oil spilled onto permafrost can influence the microbial populations (Atlas, 1981), freeze- thaw processes and soil stress (Grechishchev et al., 2001), thermal and moisture regimes (Balks et al., 2002), as well as the soil pH and nutrient availability (Everett, 1978). Most of all, the same levels of contamination may have a greater impact on the environments of cold regions than on the other environments, as the cold ecosystems have adapted to harsh conditions in ways that make them more sensitive (Snape et al., 2003).

Physical cleanup methods, such as physical barrier, excavation, and incineration, are detrimental to frozen ground and cold region ecosystems, and can quickly attenuate contaminants to a considerable extent and scope. Bioremediation normally follows physical treatment but since it is effective and economic in removing oil with less undue environmental damages is a promising option for remediation. In practice, physical and biological remediation processes need to collaborate with consideration to geography, water bodies, habitation, logistics, environmental guidelines, regulations, and remediation protocols (Yang et al., 2009).

In spite of foregoing, bioremediation is a relatively slow process in cold regions and the degree of success depends on a number of factors, including the properties and fate of oil spilled in cold environments, and the major microbial and environmental limitations of bioremediation. The microbial factors include bioavailability of hydrocarbons, mass transfer through the cell membrane, and metabolic limitations. As for the environmental limitations in the cold regions, the emphasis is on soil temperatures, freeze-thaw processes, oxygen and nutrients availability, toxicity, and electron acceptors. There have been several

cases of success in the polar regions, particularly in the Arctic and sub-Arctic regions (Yang et al., 2009).

8.1 Properties of spilled oil

Both crude and refined oil are complexes and their physical and chemical properties are determined by constituents and each percentage (Yang et al., 2009).

Low ambient temperatures usually result in increased viscosity of oil, reduced evaporation of volatiles, and increased water solubility, and thus delayed onset of biodegradation (Atlas, 1991; Margesin and Schinner, 2001). Oil spills in winter on taiga forests underlain with permafrost were long-term and chronic, while the summer spills were relatively shorter-term and more acute (Collins et al., 1993). In addition, the volume of spilled oil, the impacted area, the duration of the spill should be the major concerns (Barnes and Filler, 2003).

The susceptibility to microbial attack differs in hydrocarbons. Aliphatic compounds are more prone to degradation than polycyclic aromatic hydrocarbons (PAHs) (Margesin and Schinner, 1999) because of the low water solubility and high sorption capacity of PAHs in cold environments (van Hamme et al., 2003). Isoprenoids, phytanes, and pristanes are more resistant to bacterial metabolism at 4 ºC (Atlas, 1991).

8.2 Fate of oil spilled in the cold environments

Containment and removal are the most important factors in minimizing the environmental and economic impacts resulting from oil spilled in the cold environments are as well as the rapidity and degree of success. The methods employed for containment differ in summer and winter. The degree of success may be extensively modified during the transitio n periods of spring and fall, especially if there are nearby bodies of water, involvement of surface ice, or heavy runoffs. The degree of success also depends on the type and the viscosity of the spilled oil. Bioremediation enters the cleanup program after the containment and removal. The containment of a summertime oil spill requires rapid (within hours) deployment of a surrounding boom to be followed by vacuum pumping and

physical removal. This is often followed by the mopping up of as much oil as possible using sheets of absorbent materials (Yang et al., 2009).

Rapid containment and removal of the oil from the surface are also equally important first- priority actions for wintertime spilled oil, which, especially the crude oil usually will not attenuate as rapidly and thus will provide smaller areas of contamination. If snow is readily available, it can be used, when tamped down, to provide an effective containment boom. Snow can also act as an inexpensive, readily removable absorbent material. Except when the surface materials are sands and gravels, the frozen surface of the ground often slows the penetration of the oil spilled. Lakes and ponds with ice can be the most effectively cleaned, but containment booms and absorbent treatment are required when the ice melts in spring. Attenuation and biodegradation, after the physical removal of as much of the spilled oil as possible, are the natural methods of reducing the environmental impacts of spilled oil. The processes are long-term owing to environmental limitations (Filler et al., 2006), especially in the cold regions (Yang et al., 2009).

8.3 Cold-adapted, oil-degrading microorganisms

Cold-adapted microorganisms are able to grow at temperatures around 0 ºC and are widely distributed in the biosphere at temperatures below 5 ºC (Margesin and Schinner, 1999), which plays a crucial role in the in-situ biodegradation. The degraders are a diverse group of bacteria and fungi with more than 100 species representing 30 microbial genera involved (Atlas, 1981).

Hydrocarbon degraders may comprise less than 0.1% of the microbial community in unpolluted environments but can constitute up to 100% of the viable microorganisms in oil-polluted ecosystems (Atlas, 1981). Effective biodegradation requires an appropriate population of degraders that have adequate tolerance to environmental changes (Mosbech, 2002), and the environment should be conducive to potential active microorganisms (Thomassin-Lacroix, 2000). The length of time that degraders become acclimated and enriched before the degradation provides an approximate clue of how rapidly the microbial populations in different cold environments respond, and thus helps in making cleanup decisions (Yang et al., 2009).

8.4 Factors affecting biodegradation

Bioremediation normally takes place in the active layer, which is on the top of permafrost; therefore, the bioremediation effectiveness mainly depends on some limitations which are often quoted in cold environments (Yang et al., 2009).

Temperature: Temperatures on the surface area have particular effects on the microbial colonization and the residual hydrocarbons after evaporation for microbial attack. Ambient temperatures influence the physical nature and chemical composition of oil, rate of hydrocarbon degradation, and composition of microbial communities (Atlas, 1981), as well as the mass transfer of substrate and/or electron acceptors in frozen ground, which are crucial to the cold-adapted microbes and consequent bioremediation (Ostroumov and Siegert, 1996; Aislabie et al., 2006). Fluctuation, duration, and variable frequency of temperatures differ from site to site and the resultant biodegradation may be diverse.

Although the microbial biodegradation activity does not cease at subzero temperatures, the optimum temperature for biodegradation is usually 15–30 ºC for aerobic processes and 25– 35 ºC for anaerobic processes (Rawe et al., 1993).

Biodegradation of heavy fuel (Bunker C) by indigenous organisms in the North Sea was four times greater in summer (18 ºC) than in winter (4 ºC) (Balks et al., 2002). In the Arctic/sub-Arctic environments, the biodegradation decreases during winter period and the temperature threshold for remarkable oil biodegradation is around 0 ºC (Siron et al., 1995). Therefore, bioremediation should take advantage of the warm season in the cold regions since warmer months correlate with better degradation rates (Yang et al., 2009).

Bioavailability: Bioavailability is the tendency of individual oil components to be taken up by microorganisms (Yang et al., 2009).

As for the microbial aspects, difficulties in bioavailability result from the obstacles for hydrocarbons transferring into cellulous enzymes and from limitations in energy for maintaining degradation (Yang et al., 2009).

Access to microbial cells: The first obstacle is the access of hydrocarbons to the cellular enzymatic system because the biodegradation of chemicals relies on enzymes within the bacterial cell. The microorganisms can be metabolically active only when mass transfer across the cell membrane occurs. When the ambient temperature is lowered toward the freezing point, the channels in the cell membrane tend to be closed and cytoplasm is subject to cryogenic stress. When the cytoplasmic matrix becomes frozen, the cell will stop functioning (Finegold, 1996).

The aqueous solubility of a pollutant is important in bioremediating contaminants because the soil adsorption of contaminants correlates directly with the octanol-water partition coefficient (Kow) and inversely with the aqueous solubility (Bressler and Gray, 2003). When a compound is absorbed strongly in soil organics, the bioavailability tends to decline, therefore impeding biodegradation. The value of logKow is a good overall indicator for the solubility of a contaminant in biological membranes and lipids. Contaminants with logKow of 1 to 3.5 tend to have the highest rates of aerobic and anaerobic biodegradation (Bressler and Gray, 2003). With very low water solubility, the maximum rate of bioremediation is dictated solely by mass transfer limitations. However, mass transfer in frozen soils depends on the liquid water or water films, which is a limitation especially in permafrost environments (Ostroumov and Siegert, 1996).

Metabolic limitations: Metabolic limitations may result from the enzyme-substrate interaction and the energy needed to activate metabolism. If the proper enzyme already exists, the degradation rate may then be determined by specific interactions of the compound with the enzyme (Schwarzenbach et al., 1993).

The reactions require the activation energy produced by microorganisms to support enzyme activities. The level of energy consumption may serve as an indicator of bioremediation levels. Considerably lower adenosine triphosphate (ATP) levels at the oil spill site may indicate considerably lower levels of microbial biomass or lower microbial activity there (Sparrow and Sparrow, 1988). Decreased ATP levels were also found after a crude oil spill at an Arctic tundra site and after an oil spill in Antarctic tundra (Karl, 1992).

Oxygen: Oxygen is usually severed as the terminal electron acceptor in metabolism and oxygen limitation is one of the crucial reasons for bioremediation failures in cold regions. The importance of oxygen comes from the participation of oxygenases and molecular oxygen involved in the major degradation pathways for the hydrocarbons. Aerobic processes mostly yield a considerably greater potential energy yield per unit of substrate and tend to occur considerably more rapidly. Theory suggests that the mass of oxygen necessary to remediate the hydrocarbon load is about 0.3 g oxygen for each gram of oil oxidized (Atlas, 1981). Oxygen supply, however, is a common constraint to the bioremediation in frozen ground because oxygen is scarce and the oxygen diffusion is partly or completely blocked. Within these environments, oxygen transport is considered to be the rate-limiting step in aerobic bioremediation. Oxygen may be consumed faster than it can be replaced by diffusion from the atmosphere, and the soil may become anaerobic. In this circumstance, aerobic degradation will be limited, the transformation rates will decline, and obligate anaerobic organisms gradually become the dominant populations (Atlas, 1981; Bressler and Gray, 2003).

Alternate electron acceptors: Aerobic organisms utilize elemental oxygen as their ultimate electron acceptor. Anaerobes utilize nitrates, sulfates, CO2, and ferrous metals as electron acceptors. The redox potential of frozen ground, varying owing to freeze-thaw processes, can influence the metabolic pathway and biodegradation (McFarland and Sims, 1991). It should be greater than 50 Mv for aerobic metabolism, and less than 50 mV for anaerobic metabolism. At low redox potentials, alternative electron acceptors act as electron acceptors (McFarland and Sims, 1991). Therefore, if the redox potential changes with soil

freezing and thawing, the metabolic pathways respond to the changes, and so do the activities and composition of enzymes involved in bioremediation (Yang et al., 2009).

Nutrients: A group of nutrient elements or organic compounds is required as a source of carbon or electron donor/acceptor (Yang et al., 2009).

Inorganic nutrients including exchangeable cations, nitrates, and phosphates are important for bioremediation. However, nitrogen, and to a less extent, phosphorus are in low concentration in cold regions such as the Arctic environments, and low concentrations of some amino acids, vitamins, or other organic molecules are also needed for bioremediation (Thomassin-Lacroix, 2000).

Microbial activities can be constrained by the limitations of both nutrient supply and transport affected by freeze-thaw processes of soils. In some cases, slow-releasing fertilizers should be used if rapid dissolution and dilution of fertilizers in water systems fail to effectively stimulate biodegradation. Excessively high nitrogen levels, e.g., C/N ratios less than 20, may result in inhibited soil microbial activity possibly owing to nitrite toxicity (Thomassin-Lacroix, 2000).

Toxicity: Experiments show that lichens and mosses suffer particularly heavy mortality from toxicity. A hydrophobic coat of oil, which covers the root, may disrupt the root nutrient uptake (Jenkins et al., 1978).

The toxicity depends on the petroleum composition and concentration. Refined oil products, for example, are found to be more toxic to plant cover than crude oil (Jenkins et al., 1978). Acute toxicity usually results from low-molecular-weight alkanes and aromatics, while chronic toxicity is from PAHs. Toxicity is also related to ambient temperatures and the consequent weathering of volatile compounds. With higher air temperatures, more toxic components will be lost through weathering. Microbes are able to

degrade a contaminant when its concentration is below the toxic threshold, but their growth and viability are restricted when the contaminant is above the threshold concentration (Bressler and Gray, 2003).

Freeze-thaw processes in the active layer: The active layer, subject to freeze-thaw cycles, shows distinct seasonal variations in soil temperature and moisture. When soil freezes, moisture migration to the freezing surfaces leads to desiccation in some layers, and the unfrozen water content decreases with the increased ice content. The unfrozen water can function as a cryoprotector against cell damage from growing ice crystals, as well as serve as the solvent and medium for nutrients (Steven et al., 2006). This water phase change can result in both physical constraints to oil migration and biological effects on microbial habitats (Gilichinsky et al., 1993). For example, the number of viable cells isolated from permafrost declines to tens of thousands lower than that in an environment with positive temperatures, (Walker et al., 2006). The changes of moisture contents can also result in changes in soil water osmotic pressure. With the repulsion of the dissolved salt to unfrozen water when ice grows, the osmotic pressure balance of the cell will be broken (Nadelhoffer, 1992). Thus, membrane lipid, protein synthesis, and enzyme activity sensitive to the freeze-thaw processes will be damaged (Margesin and Schinner, 1999). The microbial physiology should be in a resting state, and the metabolism may cease under these circumstances (Vorobyova et al., 1997). On the contrary, when ice melts, the accumulation of water in the deeper active layer tends to drive the oxygen out of the free voids, resulting in relatively reductive environments unsuitable for aerobic microbes (Finegold, 1996). As a result, the metabolic activity of microbes will be limited, at least during the early and late periods of the freeze-thaw processes in the active layer of permafrost (Yang et al., 2009).

8.5 Techniques, successful cases, challenges and constraints

Biostimulation: The most widely used technique is to add inorganic nutrients or oleophilic fertilizers, which reportedly has a positive effect for oil decontamination in cold ecosystems (Allard and Neilson, 1997).

Studies have been conducted on the biodegradation of oil in the cold Arctic (Pritchard and Costa, 1991) and Antarctica (Kerry, 1993), all reporting a favorable effect of nutrient addition when aeration was provided (Yang et al., 2009).

During bioremediation treatments, it is important to monitor the availability of nitrogen and phosphorus and the concentration of potassium. However, the degree to which the oxygen and nitrogen availability can limit the biodegradation and the appropriate interval for the re-supply of oxygen and nitrogen are both uncertain (Yang et al., 2009).

Bioaugmentation: Some studies on experimentally and chronically oilpolluted cold alpine soils elucidated that augmentation with cold-adapted biodegraders was unsuccessful (Margesin and Schinner, 1997). A bioventing (bioventilacion) simulating the cold alpine soil system showed decreased hydrocarbon utilization by the inoculum in the soil while increasing degradation by the indigenous soil microorganisms with increasing incubation time and temperature (Margesin and Schinner, 1999). However, it often failed to enhance hydrocarbon degradation rates or the total hydrocarbon removal (Møller et al., 1995; Vogel, 1996; Allard and Neilson, 1997).

The cold environments will exert selective pressures on both exotic and indigenous species once these microbes are released. The inoculated microbes will probably be outcompeted by the indigenous organisms in some cases (Yang et al., 2009).

Advances in genetic engineering may lead to the development of a new field of metabolic engineering that involves the improvement of cellular activities by manipulation of enzymatic, transport, and regulatory functions of the cell using recombinant DNA technology (Yang et al., 2009).

Successful cases: A well-known example of bioremediation, which highlighted the usefulness of this treatment strategy and accelerated its development, was in the biological cleanup in the large accidental oil spill by the tanker Exxon Valdez in Alaska in March 1989. The accident

spilled approximately 41 000 m³ of crude oil and contaminated about 2 000 km of coastline. Bioremediation was extensively used. Nutrient addition was used in coastal environments including beaches and marshes (Bragg et al., 1994; Wright et al., 1997). Fertilizers were typically applied on the surfaces of sand and sediments contaminated with oil, but the application was not feasible for large areas of contamination because it required huge quantities of nutrients. The study of using fertilizers in one shoreline following the Valdez spill resulted in a fivefold increase in oil degradation (Bragg et al., 1994). Recent Alaskan bioaugmentation projects suggest that commercially available fertilizers are as or more effective than commercial bioproducts (Braddock et al., 1997).

Westlake et al. (1978) monitored over three years the effect of fertilization on the in-situ degradation of oil in a soil in northern Canada. Application of urea phosphate caused a rapid increase in bacterial number, resulting in a rapid disappearance of n-alkanes and isoprenoids, and a continuous loss in weight of saturated compounds in the recovered oil. A test in Antarctic mineral soils showed promise after one year only in the fertilizer-treated plots (C:N 61:1, C:P 607:1), and the fertilized soils had the highest level of microbial activity relative to untreated plots (Margesin and Schinner, 1999).

Challenges and constraints: It has been demonstrated that snow, ice, and frozen ground can act as physical containments or barriers to limit the scope, rate, and extent of oil spills (Biggar et al., 1998; Mosbech, 2002). In winter months, oil can be essentially sequestered with little weathering owing to the blockage by existing snow and ice, and therefore it favors physical clea nup approaches. However, a comparative test showed that the presence of ice clearly kept the number of hydrocarbon degraders about a factor of 10 lower than systems without ice (Mosbech, 2002). However, during summer months, melting snow and rain can potentially disperse oil laterally and horizontally; the trapped oil will be released and consequently become a potentially long-term source of oil slicks in the runoff waters. Freeze-thaw processes make bioremediation more complicated because of the nonlinear coupling of temperature and water regimes, mass transfer, and metabolic activities (Yang et al., 2009).

It should be emphasized that all the procedures are site-specific since they must take into account the local physical, environmental, and microbiological issues. They also are contaminant-specific, and should take into account both the pathways and the regulative mechanisms for degradation of the complex of contaminants. It depends largely on understanding of the migration mechanisms of spilled oil, as well as its interactions with local geology, ecology, and climate (Yang et al., 2009).

Generally, biological communities in the cold regions are more sensitive than those in temperate areas because of the complexity and variability of climate, population, permafrost, and environmental sensitivity in the cold regions. However, hydrocarbons can be degraded by microorganisms when major factors, such as nutrient availability (including oxygen), organic compound bioavailability, and temperatures are optimized (Yang et al., 2009).

It is critical to assess the potential of and the biodegradation response measures to accidental oil spills in permafrost areas in advance. It is also urgent for us to conduct research relevant to the cleanup technologies, especially using bioremediation approaches (Yang et al., 2009).

9 CONCLUSIONS

(1) The combination of bioaugmentation, biostimulation, and biosurfactant addition, depending on the characteristics of the contaminated site, is a promising strategy to speed up bioremediation.

(2) Production and export of lipids by Alcanivorax may constitute a step in the synthesis of biosurfactants to increase the availability of hydrophobic substances and hence improving the probability of survival in habitats with presence of hydrophobic compounds.

(3) Oil–mineral aggregate (OMA) formation is an effective alternative technology for oil spill cleanup. The oil droplets trapped in OMAs were stabilized and prevented from re-coagulating reducing their droplet size. Smaller oil droplets with a larger surface area exhibit higher rates of oil biodegradation.

(4) Hydrophobicity plays an important role in the interaction between mineral particles and oil droplets. Hydrophobic minerals tend to have more oil–mineral attraction than hydrophilic minerals.

(5) Snow, ice, and frozen ground act as physical containments or barriers to limit the scope, rate, and extent of oil spills. However, the presence of ice keeps the number of hydrocarbon degraders about a factor of 10 lower than systems without ice.

(6) Adding inorganic nutrients or oleophilic fertilizers, which reportedly has a positive effect for oil decontamination in cold ecosystems.

(7) Bioremediation process is an eco-friendly and cost-effective alternative treatment.

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