Bioremediation of PAHs - Limitations and soultions

Introduction 1.1 Occurrence of polycyclic aromatic hydrocarbons (PAH) in the environment Worldwide industrial and agricultural developments have released a large number of natural and synthetic hazardous compounds into the environment due to careless waste disposal, illegal waste dumping and accidental spills. As a result, there are numerous sites in the world that require cleanup of soils and groundwater. Polycyclic aromatic hydrocarbons (PAHs) are one of the major groups of these contaminants (Da Silva et al., 2003). PAHs constitute a diverse class of organic compounds consisting of two or more aromatic rings with various structural configurations (Prabhu and Phale, 2003). Being a derivative of benzene, PAHs are thermodynamically stable. In addition, these chemicals tend to adhere to particle surfaces, such as soils, because of their low water solubility and strong hydrophobicity, and this results in greater persistence under natural conditions. This persistence coupled with their potential carcinogenicity makes PAHs problematic environmental contaminants (Cerniglia, 1992; Sutherland, 1992). PAHs are widely found in high concentrations at many industrial sites, particularly those associated with petroleum, gas production and wood preserving industries (Wilson and Jones, 1993). 1.2 Remediation technologies Conventional techniques used for the remediation of soil polluted with organic contaminants include excavation of the contaminated soil and disposal to a landfill or capping - containment - of the contaminated areas of a site. These methods have some drawbacks. The first method simply moves the contamination elsewhere and may create significant risks in the excavation, handling and transport of hazardous material. Additionally, it is very difficult and increasingly expensive to find new landfill sites for the final disposal of the material. The cap and containment method is only an interim solution since the contamination remains on site, requiring monitoring and maintenance of the isolation barriers long into the future, with all the associated costs and potential liability. A better approach than these traditional methods is to completely destroy the pollutants, if possible, or transform them into harmless substances. Some technologies that have been used are high-temperature incineration and various types of chemical decomposition (for example, base-catalyzed dechlorination, UV oxidation). However, these methods have significant disadvantages, principally their technological complexity, high cost , and the lack of public acceptance. Bioremediation, on the contrast, is a promising option for the complete removal and destruction of contaminants. 1.3 Bioremediation of PAH contaminated soil & groundwater Bioremediation is the use of living organisms, primarily microorganisms, to degrade or detoxify hazardous wastes into harmless substances such as carbon dioxide, water and cell biomass Most PAHs are biodegradable unter natural conditions (Da Silva et al., 2003; Meysami and Baheri, 2003) and bioremediation for cleanup of PAH wastes has been extensively studied at both laboratory and commercial levels- It has been implemented at a number of contaminated sites, including the cleanup of the Exxon Valdez oil spill in Prince William Sound, Alaska in 1989, the Mega Borg spill off the Texas coast in 1990 and the Burgan Oil Field, Kuwait in 1994 (Purwaningsih, 2002). Different strategies for PAH bioremediation, such as in situ , ex situ or on site bioremediation were developed in recent years. In situ bioremediation is a technique that is applied to soil and groundwater at the site without removing the contaminated soil or groundwater, based on the provision of optimum conditions for microbiological contaminant breakdown.. Ex situ bioremediation of PAHs, on the other hand, is a technique applied to soil and groundwater which has been removed from the site via excavation (soil) or pumping (water). Hazardous contaminants are converted in controlled bioreactors into harmless compounds in an efficient manner. 1.4 Bioavailability of PAH in the subsurface Frequently, PAH contamination in the environment is occurs as contaminants that are sorbed onto soilparticles rather than in phase (NAPL, non aqueous phase liquids). It is known that the biodegradation rate of most PAHs sorbed onto soil is far lower than rates measured in solution cultures of microorganisms with pure solid pollutants (Alexander and Scow, 1989; Hamaker, 1972). It is generally believed that only that fraction of PAHs dissolved in the solution can be metabolized by microorganisms in soil. The amount of contaminant that can be readily taken up and degraded by microorganisms is defined as bioavailability (Bosma et al., 1997; Maier, 2000). Two phenomena have been suggested to cause the low bioavailability of PAHs in soil (Danielsson, 2000). The first one is strong adsorption of the contaminants to the soil constituents which then leads to very slow release rates of contaminants to the aqueous phase. Sorption is often well correlated with soil organic matter content (Means, 1980) and significantly reduces biodegradation (Manilal and Alexander, 1991). The second phenomenon is slow mass transfer of pollutants, such as pore diffusion in the soil aggregates or diffusion in the organic matter in the soil. The complex set of these physical, chemical and biological processes is schematically illustrated in Figure 1. As shown in Figure 1, biodegradation processes are taking place in the soil solution while diffusion processes occur in the narrow pores in and between soil aggregates (Danielsson, 2000). Seemingly contradictory studies can be found in the literature that indicate the rate and final extent of metabolism may be either lower or higher for sorbed PAHs by soil than those for pure PAHs (Van Loosdrecht et al., 1990). These contrasting results demonstrate that the bioavailability of organic contaminants sorbed onto soil is far from being well understood. Besides bioavailability, there are several other factors influencing the rate and extent of biodegradation of PAHs in soil including microbial population characteristics, physical and chemical properties of PAHs and environmental factors (temperature, moisture, pH, degree of contamination). Figure 1: Schematic diagram showing possible rate-limiting processes during bioremediation of hydrophobic organic contaminants in a contaminated soil-water system (not to scale) (Danielsson, 2000). 1.5 Increasing the bioavailability of PAH in soil Attempts to improve the biodegradation of PAHs in soil by increasing their bioavailability include the use of surfactants , solvents or solubility enhancers.. However, introduction of synthetic surfactant may result in the addition of one more pollutant. (Wang and Brusseau, 1993).A study conducted by Mulder et al. showed that the introduction of hydropropyl-ß-cyclodextrin (HPCD), a well-known PAH solubility enhancer, significantly increased the solubilization of PAHs although it did not improve the biodegradation rate of PAHs (Mulder et al., 1998), indicating that further research is required in order to develop a feasible and efficient remediation method. Enhancing the extent of PAHs mass transfer from the soil phase to the liquid might prove an efficient and environmentally low-risk alternative way of addressing the problem of slow PAH biodegradation in soil.

DNA barcoding taxonomy and diversity of the Mediterranean sharks and rays

Elasmobranchs are an important by-catch of commercial fisheries targeting bony fishes. Fisheries targeting sharks are rare, but usually almost all specimen bycatched are marketed. They risk extinction if current fishing pressure continues (Ferretti et al., 2008). Accurate species identification is critical for the design of sustainable fisheries and appropriate management plans, especially since not all species are equally sensitive to fishing pressure (Walker & Hislop 1998). The identification of species constitutes the first basic step for biodiversity monitoring and conservation (Dayrat B et al., 2005). More recently, mtDNA sequencing has also been used for species identification and its use has become widespread under the DNA Barcode initiative (e.g. Hebert et al. 2003a, 2003b; Ward et al. 2005, 2008a; Moura et al 2008; Steinke et al. 2009). The aims of this work were: 1) identify sharks and skates species using DNA barcode; 2) compare species of different provenance; 3) use DNA barcode for misidentified species. Using DNA barcode 15 species of sharks (Alopias vulpinus, Centrophorus granulosus, Cetorhinus maximus, Dalatias licha, Etmopterus spinax, Galeorhinus galeus, Galeus melastomus, Heptranchias perlo, Hexanchus griseus, Mustelus mustelus, Mustelus punctulatus, Oxynotus centrina, Scyliorhinus canicula Squalus acanthias, Squalus blainville), 1 species of chimaera (Chimaera monstrosa) and 21 species of rays/skayes (Dasyatis centroura, Dasyatis pastinaca, Dasyatis sp., Dipturus nidarosiensis, Dipturus oxyrinchus, Leucoraja circularis, Leucoraja melitensis, Myliobatis aquila, Pteromylaeus bovinus, Pteroplatytrygon violacea, Raja asterias, Raja brachyura, Raja clavata, Raja miraletus, Raja montagui, Raja radula, Raja polystigma, Raja undulata, Rostroraja alba, Torpedo marmorata, Torpedo nobiliana, Torpedo torpedo) was identified.

Molecular phylogeny and historical zoogeography of four genera of Eastern Atlantic skates (Rajiformes, Rajidae)

The aim of this study was to reconstruct a solid phylogeny of four genera of the Rajidae family (Chondrichthyans: Batoidea) using a concatenated alignment of mtDNA genes. Then use the resultant tree to estimate divergence time between taxa based on molecular clock and fossil calibration and conduct biogeographic analysis. The intent was to prove that the actual distribution of species of Eastern Atlantic and Mediterranean skates is due to a series of vicariant events. The species considered belongs to two different tribe: Rajini (Raja and Dipturus) and Amblyrajini (Leucoraja and Rajella). The choice of this genera is due to their high presence in the area of interest and to the richness of endemic species. The results show that despite the ancient origin of Rajidae (97 MYA), the Eastern Atlantic and Mediterranean faunas originated more recently, during Middle Miocene-Late Pliocene, after the closure of connection between these areas and the Indo-Pacific ocean (15 MYA). The endemic species of the Mediterranean (Raja asterias, R. radula, R. polystigma and Leucoraja melitensis) originated after the Messinian salinity crisis (7-5 MYA), when the recolonization of the basin occurred, and are still maintained in allopatric distribution by the presence of biogeographic barriers. Moreover from 4 to 2.6 MYA we can observe the formation of sister species for Raja, Leucoraja and Rajella, one of which has a Northern distribution, and the other has a Southern distribution (R. clavata vs R. straeleni, L. wallacei vs L. naevus, R. fyllae vs R. caudaspinosa and R. kukujevi vs R. leopardus + R. barnardi). The Quaternary and present oceanographic discontinuities that occur along the western African continental shelf (e.g., Cape Blanc and the Angola–Benguela Front) might contribute to the maintenance of low or null levels of gene flow between these closely related siblings species. Also sympatric speciation must be invoked to explain the evolution of skates, for example for the division between R. leopardus and R. barnardi. The speciation processes followed a south-to-north pathways for Dipturus and a north-to-south pathways for Raja, Leucoraja and Rajella underling that the evolution of the genera occurred independently. In the end, it is conceivable that the evolutionary pathways of the tribes followed the costal line during the gondwana fragmentation. The results demonstrate that the evolution of this family is characterized by a series of parallel and independent speciation events, strictly correlated to the tectonic movement of continental masses and paleogeographic and paleoclimatic events and so can be explained by a panbiogeographical (vicariance) model.