Potential Role for Genetic Engineering - Literature review Example

?Introduction The occurrence of persistent organic pollutants and heavy metal contamination in both terrestrial and marine systems has not gone unnoticed over the past decades (Thevenon et al., 2011). Over the years, several reports have been made on the status and magnitude of soil contamination by heavy metals and other pollutants especially in areas near factories and industrial plants (Frangi and Richard, 1997; Agoramoorthy et al., 2006; Ezeh and Chukwu, 2011; Yuan et al., 2011). In fact, current studies seem to suggest that the extent of the areas being contaminated has been increasing along with the increase in mining activities, liquefaction residues from sewages, fertilizer use, and other anthropogenic activities related to industrialization (Costa, and Jesus-Rydin, 2001; Epochtimes, 2011; Yaylahi-Abanuz, 2011). For instance, a recent report by Wu and Yu (2011) indicated that the issue of environmental pollution in China has reached an alarming extent. Specifically, the article reported that as of 2011, one-sixth of China’s agricultural land has been contaminated with mercury, cadmium, copper, and other heavy metals (Wu and Yu, 2011). According to the Chinese Academy of Agricultural Engineering, this fraction of land is equivalent to approximately 20 million hectares, with the industrialized regions being more critically affected (Wu and Yu, 2011). Heavy metal-contaminated grounds have been shown to greatly affect the floral, faunal, and microbial communities (Lukkari et al., 2004; Agoramoorthy et al., 2006; Chen et al., 2007). In a study conducted by McGrath et al. (2001), it was found that exposure to toxic metals significantly reduced microbial diversity and other biologically-mediated soil activities. This alteration in the microbial composition, according to Elsgaard et al. (2001) may negatively affect recycling of plant nutrients, regulation of plant pest and plant growth, and maintenance of soil structure. Also, pollutants from the soil could accumulate in plants and can then be transferred to higher trophic levels in the food chain, posing health hazards to humans (Notten et al., 2005). Furthermore, the effects of soil pollution are not only confined to the terrestrial compartment of the ecosystem. It can also influence the integrity and dynamics of aquatic and marine systems (Cardellicchio et al., 2006; Vorosmarty et al., 2010). By virtue of geoweathering processes, hydrocarbons, organophosphates, and metal contaminants deposited in the ground could leach down the water table and redistribute in surrounding streams, lakes, and other bodies of water (Chen et al., 2007). Given the tremendous threats soil pollution poses to the environment and to human population, the need to control soil contamination or neutralize its toxic effects and the institution of rehabilitation in areas previously affected by this problem is indeed necessary. Bioremediation, a technology that makes use of living organisms to treat polluted areas, is one of the most commonly used strategies to rehabilitate contaminated soils (Nasu and Iwamoto, 2001; Obbard et al., 2005; Umrania, 2006; Luo et al., 2010). Technically, bioremediation focuses on enhancing the natural biodegradation process at a rate that significantly increases the removal of contaminants (Calvo et al., 2009). One commonly employed bioremediation strategy is to supply the polluted environment with nutrients like nitrogen to hasten the process of degradation (Nasu and Iwamoto, 2001; Calvo et al., 2009). Another popular method is to directly introduce organisms with desired capabilities to the contaminated areas in the hope of speeding up biodegradation (Nasu and Iwamoto, 2001). Hence, by exploiting the biological processes inherent to plants or bacteria, the clean-up of environmental pollutants like hydrocarbons, lead, cadmium, and the like could be made possible (Obbard et al., 2011; Luo et al., 2010­). However, the development of enhanced capabilities for degradation or accumulation of organic and heavy metal contaminants by organism native to contaminated environments are not always ensured. After all, the acquisition of enhanced degradation traits is a function of natural evolutionary process, more specifically by virtue of random mutation (Obbard et al., 2011). This process unfortunately occurs at an extremely slow rate (Obbard et al., 2011). In this regard, the use of genetic engineering, a technology that manipulates the genome of an organism, to hasten the process of acquiring desirable traits for better bioremediation strategies has become an attractive option. This paper aims to review the possible role of genetic engineering in the field of soil bioremediation. Specifically, this paper seeks to gather information on the progress and advances genetic engineering has contributed in the bioremediation of soil contamination. Literature Review Microorganisms and plants as targets for genetic engineering The principle behind genetic engineering revolves around the idea of modifying and redesigning the genetic make-up of organisms so that intrinsic physiologic and metabolic properties are enhanced (e.g. increase enzyme production) or novel characteristics are expressed in selected subjects (e.g. acquisition of resistance to harsh environmental conditions by some plants) (Chen et al., 1999; Singh et al., 2008; Arnold, 2009; FDA, 2011). With the advancement in the field of biotechnology, it has now become possible to isolate and create multiple copies of DNA fragments containing a specific trait, and insert these gene construct in an organism’s DNA to allow the expression of the selected phenotypic characteristics. These organisms which received recombinant DNA are called genetically modified organisms (GMO). The application of genetic engineering in the field of agriculture, livestock, medicine, and industry, among others has paved the way to the emergence of tomatoes with increased shelf life and enhanced resistance to rotting, crops that can thrive in hot conditions, and human insulin and other proteins derived from bacteria, to name a few (Arnold, 2009; Dixon 2012). Indeed, the contributions of genetic engineering have gone a long way since its birth way back in 1973 (Arnold, 2009). However, it is only until the mid 1980s-1990s when experiments on genetic manipulation to improve bioremediation techniques were noted to be gaining much interest (Menn et al., 2008). Among the many ground-dwelling organisms, current literature suggests that microorganisms and plants are the two major compost agents that breakdown waste chemicals of environmental concern (Balba et al., 1998; Pieper and Reineke, 2000; Venkateswarlu et al., 2011). In fact, it was in the early to mid 1970s when the isolation and identification of wild-type bacteria that can degrade polychlorinated biphenyl (PCBs) took place (Crawford and Crawford, 1996). The work of Ahmed and Focht (1973) provided the first evidence for PCB degradation by two Achromobacter species cultured in two different media supplemented with PCBs. Even more interesting is that the byproducts of PCB degradation were different in the two cultures, indicating a difference in the metabolic pathways employed by the two bacteria. In 1975, Baxter et al. discovered a Nocardia and Pseudomonas species that exhibited PCB degradative capacities. Within a period of 100 days, the Nocardia and Pseudomonas species were found to degrade 95% and 85% of the original PCB concentration, respectively species. Finally, during the late 1970s, the research on microbial genetics of PCB-degrading bacteria finally took place, although the enhancement of PCB biodegradation by genetic engineering came about much later. On the other hand, the interest in plants as bioremediation agents was initially driven by the discovery of hyperaccumulator species that are capable of extracting potentially harmful chemicals in the soil to concentrations more than 100 times beyond what non-accumulator plants can (Chaney et al., 1997; Salt et al., 1998; Raskin and Ensley, 2000). As early as the 19th century, some plants capable of extracting zinc at unusually high levels have already been reported by Bauman (1885) (Yang et al., 2005). In a literature review conducted by Pulford and Watson (2003) on the potential use of trees for phytoremediation, it was discussed that these hyperaccumulating plant species actually developed novel metal sequestration mechanisms in response to increase metal concentration in the environment. Some of these plants were found to secrete bioactivators that could mobilized heavy metals and increase its bioavailability for uptake (Yang et al., 2005). For instance, root reductases released by some dicot species can chemically reduce Fe3+ to increase the accumulation of this metal by plants (Welch et al., 1993). In other instances, hyperaccumulating plants were found to express high-affinity metal binding proteins and transport proteins that mediate the transfer of metals across the plasma membrane such as CPx-type heavy metal ATPases, cation diffusion facilitator (CDF) family of proteins, and zinc-iron permease (ZIP) family of proteins (Bulow and Mejare, 2001; Yang et al., 2005). Metallothoneins (MTs), a family of low molecular weight, cysteine-rich peptides with strong affinity to cadmium, copper, and zinc ions have been isolated in heavy metal exposed Arabidopsis thaliana and Salicornia brachiata ((Murphy et al.,1997; Chaturvedi et al., 2012). Given these innate potentials, it should not come as a surprise that most of the genetically engineered organisms investigated for the purpose of improving bioremediation techniques revolve mostly around microorganisms and plants. Advances in bioremediation through genetic engineering Improved biosorbents for waste removal Several studies have provided evidence that one of the most common responses exhibited by microorganisms and plants in the presence of heavy metals is the expression of metal-binding proteins such as metallothioneins (MTs) (Chen et al., 1999). In a literature survey conducted by Lorenzo and Valls (2002), it was reported that the first genetic engineering experiments on bioadsorbent agents involved cloning MTs from eukaryotes and expressing these cloned genes intracellularly in Escherichia coli. This technique accounted for a 3-5 times increase in copper and cadmium metal binding in the transgenic host as reported by Romeyer and co-workers (1988). However, because of the conservative metal binding improvement and increased probability of cytotoxicity related to intracellular metal accumulation, intracellular expression recombinant MT was deemed to be not very efficient (Lorenzo and Valls, 2002). Hence, research efforts have been directed to circumvent this problem. These efforts eventually paid off when Sousa and colleagues were able to express MTs on the cell surface of Escherichia coli (Sousa et al., 1998). Specifically, the study involved insertion of MT gene fragment in the permissive site of LamB sequence, an outer membrane maltose protein (Sousa et al., 1998). The resulting recombinant protein exhibited increase in cadmium binding by as much as 15-20 times more than the native E. coli (Sousa et al., 1998). Aside from extracellular MT expression to improve metal uptake, genetic engineering has also paved the way for the simultaneously activation of metal transport system and metal-binding proteins (Lorenzo and Valls, 2002; Deng et al., 2005). This strategy was first demonstrated by Chen and Wilson (1997) when they express MTs along with mercury transport proteins MerP and MerT in E.coli. Metal transporters assist the translocation of metal ions from the environment across the plasma membrane and into the cell. Hence, co-expression of metal transporters with MTs would theoretically increase metal uptake. Indeed, Chen and Wilson (1997) reported a significant increase in mercury accumulation in the genetically-modified E. coli as compared to the wild-type species. In another study conducted by Krishnaswamy and Wilson (2000), a cloned pea MT incorporated to glutathione S-transferase in E. coli, co-expressed with a nickel transporter protein resulted in a three-fold rise in nickel uptake compared to those that expressed MTs only. Another viable strategy that has been investigated is the synthesis of new metal-binding peptides specifically designed to enhance metal binding properties or selectively accumulate certain species of heavy metals. Surprisingly, the earliest attempt to synthesize metal-binding proteins has occurred even before recombinant MT expression (Jackson et al., 1985). Jackson and colleagues (1985) sought to synthesize metal-binding peptides from Datura cell cultures but the end-product was far from perfect. Previous studies on MTs showed an abundance of cysteine residues in its amino acid sequence. It has been proposed that these cysteine moieties are responsible for MTs’ high cadmium binding. Furthermore, histidine-rich sequences have also been reported to exhibit increase in zinc, cobalt, nickel, and copper ion binding (Chen et al., 1999). In 1998, Pazirandeh and his team successfully created a metal-binding peptide with several repeats of cysteine through insertion of a pre-designed DNA fragment in LamB gene of E.coli. The resulting peptide which was targeted to be expressed on the surface of the host cell, resulted to a 10 fold rise in mercury binding (Pazirandeh et al., 1998). The following year, Kotrba et al. (1999) expressed metal-binding peptides containing histidine or cysteine-rich structure on the surface of E. coli and noted enhanced metal uptake. Increase in nickel and cadmium binding was also noted by Samuelson et al. (2000) when certain strains Staphylococcus species were engineered to express a poly-histidine protein. In plants, enhanced metal tolerance has been achieved by over-expressing detoxification pathways such as phytochelatins (PCs) and glutathione (GSH). For instance, Zhu et al. (1999) inserted genes expressing y-ECS and GSH synthase from E. coli into an Indian mustard, resulting in significant rise in cadmium accumulation compared to the non-recombinant species. In another report, the Hg-accumulating activity of properties of poplar trees was significantly increased when bacterial mercuric reductase genes, merA and merB were successfully expressed by the host tree (Yang et al., 2005). Enhancing biodegradation Polycyclic aromatic hydrocarbons (PAHs) are among the most common soil pollutant that remains in the environment for quite a long period of time (Rafin et al., 2004). PAHs are mostly derived from burning of wastes, liquefaction of coals, fossil fuel combustion, and wood treatment processing (Bamforth and Singleton, 2005; Venkateswarlu et al., 2011). Even so, several studies showed that these compounds are amenable to degradation by enzymes secreted by microorganisms and plants (Bogan et al., 2003; Rafin et al., 2004; Nam et al., 2006; Tam et al., 2010). The most studied among the enzyme systems are the cytochrome P450 monooxygenases (CYPs) and oxidoreductases like laccases (Obbard et al., 2005). Laccases are lignin-modifying enzymes with copper atoms situated at its catalytic site (Tyagi et al., 2010). Specifically, laccases use molecular oxygen for the oxidation of aromatic and non-aromatic hydrogen donors. Moreover, it has the capacity to utilize a wide range of substrate, making it an interesting candidate for bioremediation purposes (Tyagi et al., 2010). However, laccases are produced mostly in fungi and its application in bioremediation appears to be limited by its low-yield in small scale cultures (Tyagi et al., 2010). One strategy that has been explored to improve the yield of laccases and its bioactivity is through heterologous gene expression followed by directed evolution. Butler and colleagues (2003) elucidated this possibility when they cloned the laccase gene from Myceliophthora thermophia and incorporated the gene construct into the genome of a yeast (Saccharomyces cerevisiae). The genetically altered yeast was then subjected to directed evolution, which is a biomolecular engineering technique used to create a large number of gene variants through random mutagenesis such as DNA shuffling, error prone polymerase chain reaction (PCR), or staggered extension process recombination (Obbard et al., 2005) Results of the study indicated that after 10 generations of directed evolution, the recombinant yeast exhibited an 8-fold increase in laccase production (Butler et al., 2003). Its catalytic activity was also observed to increase by approximately 22 times and was noted to be maintained even at higher temperatures (Butler et al., 2003). However, testing the applicability of this enzyme in actual field studies is yet to be determined The findings of Butler et al (2003) that expressing enzyme pathways in foreign organisms can improve biodegradation are actually consistent with the results obtained by Izumi and co-workers (1994) on their sulfur degradation experiments. However, Izumi et al (1994) fused the sulfur-degrading gene construct into an organism that produces a biosurfactant instead of using directed evolution. The investigators identified and cloned dsz gene clusters which were found to be responsible for sulfur degradation in Rhodococcus erythropolis. The genes were then transferred in a P. aeruginosa strain that secretes a rhamnolipid biosurfactant. This produced a recombinant P. aeuroginosa that degraded dibenzothiophene four times better compared to the native Rhodococcus (Izumi et al., 1994). This study established the utility of genetic engineering in combining different traits into a single organism to improve over-all biodegration. A more recent study conducted by Zumarraga et al. (2007) revealed that directed evolution can also be used to also enhance the stability of laccases. S. cerevisiae containing a recombinant laccase DNA from M. thermophila expressed a more significantly stable enzyme after being subjected to two rounds of directed evolution (Zumarrage et al., 2007). The investigators reported that the mutant laccase was able to maintain its structural and functional integrity upon exposure to high concentrations of organic solvents that were previously shown to cause denaturation of the unmodified enzyme (Zumarraga et al., 2007). Because of better understanding of enzyme structure-function relationships, protein engineering has paved the way for the synthesis of proteins directed towards a specific contaminant or a range of substrates (Chen et al., 1999). With the advent of computer-guided three-dimensional modeling of protein structures, it is now possible to increase the binding site of dehalogenases or increase the range of pollutants dioxygenases can degrade (Chen et al., 1999). Improved Precipitation of Metals Precipitation of heavy metal contaminants is another viable approach for bioremediation (Lorenzo and Valls, 2002; Gadd 2004; Subramanian et al., 2006; Field et al., 2010). By chemically reducing metal ions into a lower redox potential, these metal species are effectively immobilized and rendered harmless (Lorenzo and Valls, 2002). In bacteria, metal precipitation is achieved through dissimilatory metal reduction. This process precipitates metal ions without assimilating the end-products, keeping the resulting compounds trapped in the external environment (Lovley, 1993) The production of sulfides by sulfate-reducing bacteria (SRBs) is a natural precipitation reaction that has caught the interest of many researchers and in fact, has already gone through genetic enhancement (Krumholz et al., 2009; Lorenzo and Valls, 2002). Laboratory studies on SRB revealed that it could immobilize different species of heavy metals such as zinc, chromium(IV) and iron (III) (Lovley, 1993; Perez-Lopez et al., 2012). In a study conducted by Gadd and White (1993), SRBs have been found to effectively treat metal contaminated waters in bioreactors. This is also consistent with a more recent report of Bouallagui et al. (2009) that Desulfovibrio spp., an SRB has maintained chemical removal yields at an average of 90% even at increasing wastewater loading rates in anaerobic fixed bed bioreactors. However, it must be noted that bioprecipitation of heavy metals is not limited to SRBs. Iron (III)-reducing bacteria has gained considerable interest over the years due to its potential for bioremediation (Llyod and Lovley, 2001). However, the utility of SRBs in metal precipitation is hampered by its sensitivity to metals such as Ni(II), Cd(II), and Zn(II) even at very low concentrations. Since then, attempts to endow sulfide-mediated metal precipitation to other bacterial species with better heavy metal tolerance have been conducted (Lorenzo and Valls, 2002; Keller et al., 2011). For instance, in a study conducted by Bang et al. (2000), a thiosulfate reductase gene from Salmonella enterica was transformed in E. coli. Results indicated that the engineered E. coli exhibited significant increase in sulfide production in both anaerobic and aerobic environments (Bang et al., 2000). Another progress genetic engineering has contributed to bioprecipitation of metals is the expression of sulfate-reducing pathways under aerobic conditions. SRBs are known to have restricted viability in the presence of oxygen, limiting its application to anaerobic environments. Hence, Wang et al. (2000) attempted to engineer an aerobic sulfate reducing pathway in E.coli by simultaneously expessing serine acetyltransferase and cysteine desulfhydrase. This approach resulted in production of sulfide with concomitant precipitation of cadmium in the presence of oxygen (Wang et al., 2000). Conclusion Genetic engineering indeed has increased the inherent potential of microorganisms and plants for bioremediation purposes. It has improved the efficiency of microorganisms to degrade, immobilize, accumulate or bind soil pollutants several times better compared to the wild types. Not only it has boosted the efficiency of metabolic pathways naturally present in an organism, it has also demonstrated the possibility of conferring foreign traits to more robust organisms, transfer and express bacterial genes to plants, or combine desired characteristics in a single bacterial strain to improve over-all bioactivity. Genetic engineering has also allowed simultaneous expression and activation of certain traits to enhance removal of pollutants. Taking these things into consideration, it can be said that genetic engineering has already elucidated its application in designing viable soil bioremediation strategies in the laboratory setting. The next big step to be undertaken is perhaps to take these genetically-enhanced organisms in the field and evaluate their feasibility for actual environment clean-up. As of present, only a few studies have been reported to have done field release experiments on genetically engineered organisms. One example of an in situ experiment using GMO was conducted by Cox et al. (2000) on Pseudomonas fluorescens HK4 which is capable of enhanced napththalene degradation. However, this field experiment was done in an artificial environment where lysimeter structures made of removable stainless steels were filled with PAH contaminated soil. 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