Introduction heavy metal pollution

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Pesticides have long history since the emergence of agriculture. Human beings are facing the development of pests (including weeds, insects and pathogenic agents) causing considerable agricultural losses. If these pests are not controlled, they diminish the quality and quantity of crop production (Richardson, 1998). In the beginning, either some inorganic chemicals or compounds extracted from plants were used as pesticides. The pyrethrine was extracted from Chrysanthemum flowers and used to control the pest development during winter storage of crop. This was reported by the Greek civilization and authorization of this compound is still going on. However, agricultural revolution in the 19th century has lead to the intensive and diversified use of the pesticides corresponding to compounds derived from minerals and plants. As an example, the development of Bouillie Bordelaise (Bordeaux mixture) in 1880, consisting of copper sulphate and lime allowed better control of cryptogamic diseases in Bordeaux and French vineyard. It is still in use for vineyard and fruit tree protection. Development and application of pesticides for the control of various insectivorous and herbivorous pests is considered as fundamental contributor to this “Green Revolution”. The use of synthetic organic pesticides began during the early decades of 20th century and increased tremendously after the World War II, with the introduction of synthetic organic molecules such as DDT, aldrin (two insecticides) and the herbicide 2,4-D in the agricultural market. Due to their advantages of being effective and cheap, use of synthetic pesticides is continously increasing in the whole world.

Although pesticide application ensures better yield in agricultural production, however, when they contaminate the soil and water resources, became harmful for the environment and living beings through the food chain (Briceno et al. 2007). Due to their intensive and repeated application, and their relative recalcitrance to biodegradation, pesticide residues are persistent in the environment where they have often been detected beyond the permissible limits in different compartments of the environment as well as in food chain. In many parts of the world, particularly in developing countries, clean drinking water is a limited resource and, in this context, intensive agricultural production is a major environmental and health problem because pesticide residues accumulate in surface and under ground water (Rasmussen et al. 2005). Contamination with pesticides is restricted not only to developing countries but also in Europeon countries, where pesticide residues have often been detected in surface and ground water resources (Gooddy et al. 2002). As a result, the use of pesticides in conventional agriculture has attracted much attention in recent years due to rising public and governmental concerns about their impact not only on environmental contamination but also on human and animal health.

Pesticide exposure to environment is dependent on various factors like production, formulation processing and application doses. A pesticide enter in to the environment via (1) direct intentional application to soil to control pre emergent weeds, plant pathogens, soil insects /pests, and/or (2) indirect unintentional entry followed by foliar application for post emergent weeds and insects/pests (Mathews, 2006; Brieceno et al. 2007). In addition to this a certain portion of pesticides may undergo spillage from formulation plants during processing and waste disposal process as well. Adverse impact of pesticides on soil biology and ecosystem have been described by many researchers (Sacki and Toyota, 2004).

These recalcitrant compounds build up regularly in the environment, as they are not at all biodegradable, and even if degraded, very slowly. Owing to low water solubility, pesticides have strong affinity for particulate matter and consequently enter in to water sediments (Giri et al. 2014). For instance Lindane, the most commonly used isomer of HCH is known to accumulate in food chains, causing toxicity in wild/domestic animals and human beings. Apart from food contamination, human beings are exposed to lindane by inhalation, polluted water and dermal contact (Giri et al. 2014).

Repeated applications of haloginated insecticide endosulfan causes its accumulation in the soil and water environment. Consequent upon accumulation, it is extremely toxic to aquatic fauna, while provoking chronic symptoms like testicular and prostate cance, breast cancer, sexual abnormality, genotoxicity and neurotoxicity in various mammalian species (Giri and Rai, 2012).
Bioremediation of Pesticides
Bioremediation is the use of microorganisms for degradation of hazardous chemicals in soil, sediments, water, or other contaminated materials. Often the microorganisms metabolize the chemicals to produce carbon dioxide or methane, water and biomass. Alternatively, the contaminants may be enzymatically transformed to metabolites that are less toxic or environmentally innocuous. It should be noted that in some instances, the metabolites formed are more toxic than the parent compound. For example, perchloroethylene and trichloroethylene may degrade to vinyl chloride which is highly toxic in nature than parent compound (Sacki and Toyota, 2004).
There are a number of possible pesticide degradation pathways in the soil and water environment including chemical treatment, volatilization, photodecomposition and incineration. However, most of them are not applicable for the diffused contamination with low concentration because of being expensive, less efficient and environmental friendly. Thus, keeping in view the environmental concerns associated with pesticides/recalcitrant compounds, there is a need to develop safe, convenient and economically viable methods for its remediation. In this context, several researchers have focused their attention to study the microbial biodegradation which has been reported as a primary mechanism of pesticide dissipation from the environment (Cox et al. 1996, Pieuchot et al. 1996). Although bioremediation strategies are more acceptable to the society because of their reduced impact on the natural ecosystem (Zhang and Quiao, 2002). However, complexity of the mechanisms responsible for pesticide degradation has made it slow to emerge as an economically viable remediation method (Nerud et al. 2003). It is noteworthy that bioremediation strategies have been developed extensively for taking care of sites heavily contaminated with organic pollutants, however, up to now, the sites diffusely contaminated are only monitored and natural attenuation is the process of interest leading to contaminant abatement.

Microbial biodegradation occurs mostly in the soil solution. Pesticide microbial biodegradation is carried out by soil microorganisms like bacteria fungi and actinomycetes possessing a large set of enzymes susceptible to transform these pesticides. It is the principal mechanism for diminishing the persistence of pesticides in soil environment (Arbeli and Fuentes, 2007). Soil serves as a potential habitat for variety of microorganisms which have the ability to interact not only with other living components but also the physical elements including pesticides for the fulfilment of their energy requirement. When pesticides are applied in the soils, enzyme-driven biochemical reactions carried out by the indigenous soil microorganisms result in modification of the structure and toxicological properties of pesticides leading to their complete conversion into harmless inorganic end products (Hussain et al. 2009a). Pesticides degradation by soil microbial communities has been reported by several researchers (Fenlon et al. 2007; Hussain et al. 2007a; Shi and Bending, 2007; Hussain et al. 2009b; Sun et al. 2009) and it has been described as a primary mechanism of pesticide dissipation from the environment (Fournier et al. 1975; Cox et al. 1996; Pieuchot et al. 1996). The efficiency of pesticide biodegradation varies considerably between different groups of the microorganisms and even between the different members belonging to the same group of microorganisms. Although a strong diversity of microbial species is found in the soil, however, the adaptability of these different degrading microbial species in the contaminated soils assures the continuity of biodegradation process. Microbial biodegradation of pesticides in the soil can be categorized into two principal types based on the mode and pathway of degradation i.e. metabolic and co-metabolic.

Metabolic degradation
Metabolic pesticides degradation is carried out by soil microbial population harbouring specific catabolic enzymes allowing complete mineralization of target compound. A large number of pesticide degrading fungal and bacterial strains have been isolated and characterized from the soil environment (Hussain et al. 2009b). Although often metabolic biodegradation can leads to incomplete degradation resulting in the formation of metabolites (Turnbull et al. 2001; Hangler et al. 2007; Badawi et al. 2009). However, up to now, the full mineralization of the pesticides has been found to be taking place only by the soil bacteria. The enzymes required for metabolic degradation of pesticides are either harboured by a single microorganism or scattered in various microbial populations working as a cooperative consortium, jointly involved in the degradation of the pesticides (Fournier et al. 1996).
Co-metabolic degradation
The co-metabolic degradation corresponds to the non specific degradation of xenobiotic molecule by microorganisms. In most of the cases, this is a non-inducible phenomenon occurring because of the presence of detoxifying enzymes able to degrade xenobiotics depicting homologies with their substrate. In this case, the target pesticides do not contribute to the growth of the degrading organisms (Dalton and Stirling, 1982; Novick and Alexander, 1985). For this reason, the degradation rate of pesticide in a given environment depends primarily on the size of microbial biomass and on the competitiveness of the degrading microbial population towards sources of energy and nutrients in the soil. In other words, pesticide degradation rate is dependent on size of the biomass (Fournier et al. 1996). In general, co-metabolism does not yield in extensive degradation of the molecule but rather causes incomplete transformation such as oxidation, hydroxylation, reduction, N-dealkylation or hydrolysis (Fournier et al. 1996) which may lead to the formation of metabolites that may prove even more toxic and recalcitrant than the parent compound (De Schrijver and De Mot, 1999).

Some compounds can only be partially metabolized by microbial populations and transformed into metabolites that may either accumulate in the environment or be metabolized further by other microbial species. These metabolic reactions do not provide benefit to the responsible organism because they do not gain either carbon or energy. These processes are typically fortuitous and occur because the responsible population produces one or more enzymes that are comparatively nonspecific and can react with structural analogues compounds of the “normal” substrate for enzyme(s). Co-metabolism is important for the degradation of many environmental contaminants particularly chlorinated pesticides solvents (e.g. trichloroethylene,), polychlorinated biphenyls, and many polyaromatic hydrocarbons (Fournier et al. 1996). Giri and Rai (2012), studied Biodegradation of endosulfan isomers in broth culture and soil microcosm by Pseudomonas fluorescens. After 15 days incubation, maximum 92.80% α and 79.35% β endosulfan isomers were degraded in shake flask culture at 20 mg/L concentration, followed by 50 and 100 mg/L, while the corresponding values in static condition were 69.15 and 51.39%, respectively.

Advantages and Limitations of Bioremediation
The use of intrinsic or engineered bioremediation processes offers several potential advantages that are attractive to site owners, regulatory agencies, and the public. These include:

1. Lower cost than conventional technologies.

2. Contaminants usually converted to innocuous products.

3. Contaminants are destroyed, not simply transferred to different environmental media.

4. Nonintrusive, potentially allowing for continued site use.

5. Relative ease of implementation.

However, there are some limitations to bioremediation as well, these include:

1. Difficult to control the laboratory optimized conditions in the field.

2. Amendments introduced into the environment to enhance bioremediation may cause other contamination problems.

3. May not reduce concentration of contaminants to the required levels.

4. Requires more time.

5. May require more extensive monitoring.


Analytical Techniques
The development of new technologies and their implementation in the analysis of pesticides in environmental samples has greatly affected the way we perceive and use pesticides. In the 1940’s, pesticides were perceived as miracle chemicals that gave tremendous gain in crop yields and they were used without adequate regard to health and the environment. At the time, thin layer chromatography (TLC) with semi-quantitative detection was the primary means of analysis. Gas liquid chromatography (GLC or GC) with packed columns became the method of choice as commercial instruments improved and selective quantitative detectors were developed in the late 1950’s, to mid 1960’s. By the time of publication of Rachel Carson’s Silent Spring in 1962, GC was predominant method of pesticide analysis (Hawthorne et al. 1994). When the environmental and ecological impacts of pesticides were come in to lime light, the perception of pesticides began to change. Laws that established regulatory controls on the use of pesticides and their presence in the environment required residue analysis using state-of-the-art instrumentation (Hopper, 1996). With the development of improved capillary columns for GC in the 1970’s, tremendous gains in separation power were achieved and the capabilities of multi-residue methods improved accordingly. During the same time frame, high performance liquid chromatography (HPLC) was commercialized and its implementation in pesticide residue analysis permitted detection of many compounds that were not analysed easily. Through the complementary nature of GC and HPLC, a wide range of pesticides could be analysed and many environmentally safer pesticides were developed and registered using these sophisticated technologies (Richter et al. 1996). Presently potentially advanced and sophisticated pesticide analytical methods such Gas chromatography mass spectrometry (GC-MS) Liquid chromatography mass spectrometry (LC-MS), etc, have been developed and commercialized. Figure 2 represent a schematic representation of pesticide analysis method in environmental samples.
Preparation Sample
Extraction Organic Solvents

Liquid-Liquid Partitioning SPE GPC Concentration

Separation/ Analysis GC HPLC

Use of selective detectors and

multiple analyses are common

(SPE= solid phase extraction, GPC= gel permeation chromatography, GC= gas chromatography, HPLC= high performance liquid chromatography)

Figure 2. Pesticide residue analysis method in environmental samples

(Source: Hawthorne et al. 1994)

Recent Advances in Pesticide Bioremediation

Pesticide degrading catabolic gene and their respective enzymes of microorganisms have been isolated and identified by several researchers. For example lindane (Kumari et al. 2002), endosulfan (Sutherland et al. 2002; Hussain et al. 2007), DDT (Barraga et al. 2007) and monocrotophos degrading microbial genes and enzymes have been (Subhas and Singh 2003; Das and Singh, 2006), isolated and identified. Genetic studies revealed that plasmids are the main place to harbour pesticide catabolising genes in microbial community. Sutherland et al. (2002) had reported Esd gene having sequence homology to monooygenase family which uses reduced flavin, provided by a separate flavin reductase enzyme, as co-substrates in Mycobacterium smegmatis. Esd catalyzes the oxygenation of β-endosulfan to endosulfan monoaldehyde to endosulfan hydroxyether. Esd did not degrade either α-endosulfan or the metabolites of endosulfan and endosulfan sulphate. Wier et al. (2006) have reported that Ese gene of Arthrobacter sp. encoding enzyme from monooxygenase family is capable of degrading both the isomers of endosulfan. After, understanding the gene of interest and enzyme involved, the Superbugs can be created to achieve the desired result at fast rate in short time frame. Lal et al. [25] has reviewed the degradation of HCH and distribution of lin gene in Sphingomonads. S. indicum B90A was found to contain two non-identical linA genes (designated as linA1 and linA2). The linA-encoded HCH dehydrochlorinase (LinA) mediates the first two steps of dehydrochlorination of γ-HCH (Singh, 2008). Besides, genetically modified microbes are used to enhance the capability of pesticide degradation. However, genetically engineered technology for environment use is still controversial because an adverse genotype can be readily mobilized in the environment. In a development of technology following points should be taken care i.e. (i) heterogeneity of contaminant. (ii) concentration of contaminant and its effect on biodegradative microbe, (iii) persistence and toxicity of contaminant, (iv) behaviour of contaminant in soil environment and (v) conditions favourable for biodegradative microbe or microbial population (Singh, 2008). The degradation of persistent chemical compounds by microorganisms in the natural environment has revealed a larger number of enzymatic reactions with high bioremediation potential (Finley et al., 2010). These biocatalysts can be obtained in quantities by recombinant DNA technology, expression of enzymes, or indigenous organisms, which are employed in the field for removing pesticides from polluted sites. The microorganisms contribute significantly for the removal of toxic pesticides used in agriculture and in the absence of enzymatic reactions many cultivable areas would be impracticable for agriculture (Abramowicz, 1995).

Although, significant advances have been made in understanding the roles of plant associated microbial pesticide degradation and application of these processes in field scale bioremediation (Joshi and Juwarkar, 2009; Li et al., 2010; Shi et al., 2011). An exciting alternative to the use of plant-associated bacteria to degrade toxic organic compounds in soil is the use of recombinant DNA technology to generate transgenic plants expressing bacterial enzymes resulting in improved plant tolerance and metabolism of toxic organic compounds in soil. Transgenic plants have been produced for phytoremediation of both heavy metals and organic pollutants (Eapen et al. 2007). Transgenic poplar plantlets expressing bacterial mercuric reductase were shown to germinate and grow in the presence of toxic levels mercury. Arabidopsis thaliana was engineered to express a modified organomercurial lyase (Rugh et al. 1992) and those transgenic plants grew vigorously on a wide range of concentrations of highly toxic organomercurials, probably by forming ionic mercury which should accumulate in the disposable plant tissues. The first report of genetically modified plant for the transformation of xenobiotic contaminants to nontoxic material was reported (French et al. 1999). They previously reported that Enterobacter cloacae PB2 is capable of growth with trinitrotoluene (TNT) as a nitrogen source (Bhatia and Malik, 2011).
Ecosystem degradation resulting from resource extraction, land-use change, shifting cultivation, invasion by exotic species, forest fire and subsequent biodiversity loss alters the functions and services provided by forest ecosystems. Mineral mining exerts a long lasting impact on landscape, ecosystem and socio-cultural economic considerations. Mining and its subsequent activities have been found to degrade the land to a significant extent. Overburden removal from the coal field results in significant forest and top soil loss (Figure 3). Most of the mining wastes are inert solid materials and toxic in nature (Guha Roy, 1991). These toxic substances are inherently present in the ore, e.g. heavy metals such as iron, mercury, arsenic, lead, zinc, cadmium, etc (Giri et al. 2014). These heavy metals leach out of the stored waste piles and contaminate immediate environment. However, some toxic chemicals are also found in waste, as they are added intentionally during extraction and processing. The major environmental impacts due to coal mining are changes in soil stratification, reduced biotic diversity, and alteration of structure and functioning of ecosystems; these changes ultimately influence water and nutrient dynamics as well as trophic interactions (Giri et al. 2014). Land degradation due to forest clearance, shifting cultivation and mining activities is the cumulative effect of air and water pollution, soil quality degradation and biodiversity loss (Sankar et al. 1993). This process works through a cycle known as land degradation cycle. The magnitude and impact of mining on environment varies from mineral to mineral and also depends on the potential of the surrounding environment to attenuate the negative effects of mining, geographical disposition of mineral deposits and size of mining operations. A list of minerals has been prepared by Department of Environment, which is supposed to have serious impact on environment. These minerals include coal, iron ore, zinc, lead, copper, gold, pyrite, manganese, bauxite, chromite, dolomite, limestone, apatite and rock phosphate, fireclay, silica sand, kaolin, barytes. Mineral production generates enormous quantities of waste/ overburden and tailings / slimes (Rai, 1996, Giri et al. 2014).

Figure 3 (a) Open cast coal mining (b) acid mine drainage (c &d) coal dumping in Margherita Assam, India

Acid mine drainage is a serious environmental issue of coal/mineral mining activities. This occurs when sulphide ores are exposed to the atmosphere, which can be enhanced through mining and milling processes where oxidation reactions are initiated. Mining increases the exposed surface area of sulfur-bearing rocks allowing for excess acid generation beyond natural buffering capabilities found in host rock and water resources (Caruccio, 1975). Once acid drainage is created, metals are released into the surrounding environment, and become readily available to biological organisms. When fishes are exposed directly to metals and H+ ions through their gills, impaired respiration may result chronic and acute toxicity. Fishes are also exposed indirectly to metals through ingestion of contaminated sediments and food materials. A common weathering product of sulfide oxidation is the formation of iron hydroxide (Fe (OH)3), a red/orange coloured precipitate found in thousands of miles of streams affected by acid mine drainage (Figure 3). Iron hydroxides and oxyhydroxides may physically coat the surface of stream sediments and streambeds destroying habitat, diminishing availability of clean gravels used for spawning, and reducing fish food items such as benthic macro invertebrates. Acid mine drainage, characterized by acidic metalliferous conditions in water, is responsible for physical, chemical, and biological degradation of aquatic ecosystems (Ashraf et al. 2010). Acidic water adversely affects the soil environment by way of making the soil acidic and rich in inorganic component and poor in organic content. Deterioration of soil quality has severely affects the crop growth and yield in the area mainly due to high concentrations of hydrogen ions, which inactivate most enzyme systems, restrict respiration, and root uptake of salts and water by plants. It also leads to deficiency of nitrogen, phosphorous, calcium, magnesium, molybdenum and boron as well as iron and manganese toxicity. Solubilisation and transport of phosphorus from soil to the water environment due to acidity is an important issue associated with decreased agriculture productivity (Giri et al. 2014). Open cast coal mining and other mineral mining activities resulted forest degradation, biodiversity loss and severe environmental pollution in mining areas. These mineral mining activities are being carried out in various parts of the country such as Madhya Pradesh, Jharkhand, Chhattisgarh, Orissa, Assam, Meghalaya, Arunachal Pradesh and Nagaland.

Shifting cultivation also called slash and burn agriculture is the clearing of forested land for raising or growing the crops until the soil nutrients are exhausted and/or the site is overtaken by weeds and then moving on to clear more forest. It has been often reported as the main cause of deforestation and land degradation (Dick, 1991; Barbier et al., 1994 and Ross, 1996). Mostly all reports indicate shifting agriculture is responsible for about one half of tropical deforestation in the world (Figure 4 a & b). In India shifting cultivation/jhom cultivation is predominant in Northeast part of the country, particularly in Assam, Nagaland, Meghalaya and Mizoram. Shifting cultivation has been considered one of the major causes for ecological degradation and deforestation in the country, which has become a serious environmental issue.

Figure 4 Shifting Cultivation and Ecosystem Degradation in Mizoram

Ecological restoration embraces a broad suite of goals, ranging from amelioration of highly degraded abiotic conditions (e.g., toxic pollutant levels and the absence of topsoil on old mine sites), to the reinstatement or enhancement of key ecosystem functions (e.g., production, erosion control, water flow and quality), to the reestablishment of a target biotic community (e.g., rare species, native species, high diversity, eradication of invasive species). In terrestrial ecosystems, plant–soil interactions are the foundation for effective and sustained achievement of any of these goals. Soil conditions constrain plant performance and community composition (Grime 2001; Pywell et al. 2003), and attempts to restore plant communities are likely to fail if they do not consider the limitations imposed by soil conditions. In contrast, plant composition can impact almost every aspect of soil structure and function Ecological restoration is the practice of restoring ecosystems as performed by practitioners at specific project sites, whereas restoration ecology is the science upon which the practice is based (Eviner and Haukes, 2008).

Restoration ecology ideally provides clear concepts, models, methodologies and tools for practitioners in support of their practice. Sometimes the practitioner and the restoration ecologist are the same person the nexus of practice and theory. The field of restoration ecology is not limited to the direct service of restoration practice. Restoration ecologists can advance ecological theory by using restoration project sites as experimental areas. For example, information derived from project sites could be useful in resolving questions pertaining to assembly rules of biotic communities. Further, restored ecosystems can serve as references for set-aside areas designated for nature conservation (SER, 2004). Ecological restoration is one of several activities that strive to alter the biota and physical conditions at a site. These activities include reclamation, rehabilitation, mitigation, ecological engineering and various kinds of resource management, including wildlife, fisheries and range management, agro forestry, and forestry. At the heart of plant–soil interactions lies the microbial community. Microbial communities (Eviner et al. 2008):

1. are ultimately responsible for most biogeochemical transformations in soil,

2. can play a significant role in impacting soil structure, and

3. Can have strong effects on plant growth and competitive dynamics.

Success in eco-restoration studies requires the presence of key microbial groups, particularly those microbes that are obligate or facultative symbionts with plant roots. Plant seedlings grow substantially better when planted into a community with established mycorrhizal connections than in disturbed sites or in isolation Eviner et al. 2008. In some cases, such as with pine trees, establishment requires simultaneous introduction of plants and ectomycorrhizal fungi if these root symbionts are not already present. Addition of symbiont inoculum can also facilitate restoration efforts when microbial communities have been disturbed or altered (Eviner et al. 2008). For example mycorrhizal inoculations, have been shown to increase plant establishment and growth (Cuenca & Lovera 1992); soil organic matter, nitrogen, aggregation (Requena et al. 2001), and alter succession by shifting competitive interactions between plants (Allen & Allen 1990). In addition, inhibiting microbial symbiont establishment can be used as a tool to reduce establishment and growth of undesirable species. For example, in a study, absence of arbuscular mycorrhizal fungi (AMF) and actinorhizal Frankia, native oleaster shrub growth decreased by 4-fold (Visser et al. 1991), whereas growth of an invasive leguminous shrub decreased by 5-fold in the absence of specific Bradyrhizobium strains (Parker et al. 2006; Eviner et al. 2008).

Isolation of various plants associated microbes and characterization of its beneficial metabolites/processes are time consuming, since it requires the analysis of more than thousands of isolates. Thus strong molecular research effort is required in order to find specific biomarker associated with the beneficial microbes for efficient microbe assisted bioremediation. Although promising results have been reported under laboratory conditions, showing that inoculation of beneficial microbes particularly plant growth promoting bacteria and/or mycorrhizae may stimulate heavy phytoextraction or phytostabilization. Only a few studies have demonstrated the effectiveness of the microbial assisted plant bioremediation of pesticides and toxic metals in field conditions (Brunetti et al., 2011; Juwarkar and Jambhulkar, 2008; Wu et al., 2011a; Yang et al., 2012). Genetically engineered organisms with novel pathways will to generate new or improved activities hold a great potential for enhanced bioremediation. Using genes encoding the biosynthetic pathway of bio-surfactants can enhance biodegradation rates by improving the bioavailability of the substrates and genes encoding resistance to critical stress factors may enhance both the survival and the performance of designed catalysts. Thus, genetic engineering of indigenous microflora, well adapted to local environmental conditions, may offer more efficient bioremediation of contaminated sites and making the bioremediation more viable and eco-friendly technology. Complete genome sequences for several environmentally relevant microorganisms, mechanism of pesticide solubility, uptake and availability of nutrients/ pesticides, signalling processes that occur between plant roots and microbes, these types of analysis will surely prove useful for exploring the mechanism of pollutants-microbes-plant interactions. Moreover, such knowledge may enable us to improve the performance and use of beneficial microbes as inoculants for microbial bioremediation.

Emphasis should be placed when developing bioremediation systems using plant-associated bacteria, to choose wild type bacteria, or bacteria enhanced using natural gene transfer, to avoid the complications of national and international legislation restricting and monitoring the use of genetically modified microbes (GMMs). However, with a global political shift towards sustainable and green bioremediation technologies, the use of plant-associated bacteria to degrade toxic synthetic organic compounds in environmental soil may provide an efficient, economic, and sustainable green remediation technology for future environment (Bhatia and Malik, 2011).

Much is still unknown about tolerances, degradative capacities and ecological interactions of organisms that have potential use in eco-restoration. However, it is clear, that plants and microbes act cooperatively to improve the rates of biodegradation and biostabilization of environmental contaminants as well as improve nutrient contents in degraded lands.. Knowledge of the microbial community structure resident to the rhizosphere of plants that are resistant to a given contaminant will improve the chances of successfully increasing biodegradation rates when co-inoculating plants and microbes into contaminated environments. In designing a eco-restoration program the oxidative capacity of a plant should be considered in terms of its action on the contaminant itself and for its potential to support rhizospheric microbes with the capacity to enhance biodegradation. Additional basic biological and ecological information in these areas will allow us to make better informed decisions on how to widen bottlenecks in bioremediation/eco-restoration processes (Cohen et al. 2004).
Since the plant-associated microbes possess the capability of plant growth promotion and/or metal mobilization/immobilization. There has been increasing interest in the possibility of manipulating plant microbe interactions in contaminated soils (Aafi et al., 2012; Azcón et al., 2010; Braud et al., 2009b; Dimkpa et al., 2008, 2009a,b; Hrynkiewicz et al., 2012; Kuffner et al., 2010; Luo et al., 2011; Luo et al., 2012; Maria et al., 2011; Mastretta et al., 2009; Orłowska et al., 2011; Sheng et al., 2008a,b). Bioremediation is cost effective, faster than natural attenuation, high public acceptance and generates less secondary wastes and emerged as an integrated tool for environmental cleanup as well as ecosystem service provider. (Dickinson et al. 2009). The potential role of plants and associated rhizomicrobial population in facilitating microbial degradation for in situ bioremediation of surface soils contaminated with hazardous organic compounds is substantial. Support for this concept comes from the fundamental microbial ecology of the rhizosphere, documented acceleration of microbial degradation of agricultural chemicals and mobilization/immobilization of metals in the root zone.

Further understanding of the critical factors influencing the plant-microbe-toxicant interaction in soils will permit more rapid realization of this new approach to in situ bioremediation (Dubey and Fulekar, 2013). To effectively restore an ecosystem or ecological community, it is often critical to consider multiple species, multiple functions, and their interactions. Furthermore, the restoration of self-maintaining systems is increasingly requiring the consideration of human-induced local- to global-scale environmental changes. Studies on plant–soil interactions vis-à-vis plant microbe interaction provide an important foundation for eco-restoration. In order to help managers with the challenge of designing successful restoration techniques at a specific site, we need to embrace the variability of ecological studies and develop frameworks to understand this variability (Bever, 2002; Eviner and Hawkes, 2008). Bioremediation is not a Panacea to restore all the contaminated environmental sites, however, in comparison to other remediation processes i.e. incineration, thermal disposition, land farming etc. it has a better future in development of technology for removal of contaminants from actual site and restoration of degraded lands (Singh, 2008). With a global political shift towards sustainable and green technologies, the use of plant-associated microorganisms to degrade toxic synthetic organic and inorganic pollutants in environmental soil may provide an efficient, economic, and sustainable green remediation technology for future environment (Bhatiya and Malik, 2011).

Aafi, N. E., Brhada, F., Dary, M., Maltouf, A.F., & Pajuelo, E. (2012). Rhizostabilization of metals in soils using Lupinus luteus inoculated with the metal resistant rhizobacterium Serratia sp. MSMC 541. International Journal of Phytoremediation, (14), 261–274.

Abramowicz, D. A. (1995). Aerobic and Anaerobic PCB Biodegradation in the Environment. Proceeding from Conference on Biodegradation: Its Role in Reducing Toxicity and Exposure to Environmental Contaminants, Triangle Park, North Carolina, June, 1995.

Aislabie, J., & Lloyd-Jones, G. (1995). A Review of Bacterial Degradation of Pesticides, Australian Journal of Soil Research, 33 (6), 925-942.

Allen, E. B., Allen, M. E., Egerton-Warburton, L., Corkidi, L., & Gomez-Pompa, A. (2003). Impacts of early- and late-seral mycorrhizae during restoration in seasonal tropical forest, Mexico. Ecological Applications, (13), 1701-1717.

Anderson, C. W. N., Brooks, R. R., Chiarucci, A., LaCoste, C. J., Leblanc, M., Robinson, B. H., Simcock, R., & Stewart, R. B. (1999). Phytomining for nickel, thallium and gold. Journal of Geochemical Exploration, 67(1-3), 407–415.

Arbeli, Z., & Fuentes, C. L. (2007). Accelerated biodegradation of pesticides: An overview of the phenomenon, its basis and possible solutions and a discussion on the tropical dimension. Crop Protection, (26), 1733-1746.

Auger, C., Han, S, Appanna, V. P., Thomas, S. C., Ulibarri, G., & Appanna, V. D. (2013). Metabolic reengineering invoked by microbial systems to decontaminate aluminium: Implications for bioremediation technologies. Biotechnology Advances, (31), 266–273.

Azcon, R., Peralvarez, M. D. C., Roldan, A., & Barea, J. M. (2010). Arbuscular mycorrhizal fungi, Bacillus cereus, and Candida parapsilosis from a multi-contaminated soil alleviate metal toxicity in plants. Microbial Ecology, (59), 668–677.

Azcon-Aguilar, C., & Barea, J. M. (1992). Interactions between mycorrhizal fungi and other rhizosphere microorganisms. In M. F. Allen (Ed.), Mycorrhizal functioning, an integrative plant-fungal process (pp. 163–198). New York: Routledge, Chapman & Hall Inc.

Badawi, N., Ronhede, S., Olsson, S., Kragelund, B. B., Johnsen, A. H., Jacobsen, O. S., & Aamand, J. (2009). Metabolites of the phenylurea herbicides chlorotoluron, diuron, isoproturon and linuron produced by the soil fungus Mortierella sp. Environmental Pollution, (157), 2806-2812.

Barbier, E. B., Burgess, J. C. & Folke, C. (1994). Paradise lost? The ecological economics of biodiversity. Earthscan.

Barraga, n-Huerta B. E., Costa-Perez. C., Peralta-Cruz, J., & Barrera-Corte, J. (2007) Biodegration of organochlorine pesticides by bacteria grown in microniches of the porus structure of green bean coffee. International Biodeterioration Biodegradation, (59), 239-244.

Bernhoft, R. A. (2012). Mercury toxicity and treatment: a review of the literature. Journal of Environmental Public Health.

Bever, J. D. (2002). Host-specificity of AM fungal population growth rates can generate feedback on plant growth. Plant and Soil, (244) 281–290.

Bhatiya, D., & Malik, D. K. (2011). Plant-Microbe Interaction with Enhanced Bioremediation. Research Journal of Biotechnology, 6 (4), 1-8.

Boominathan, R., Saha-Chaudhury, N. M., Sahajwalla, V., & Doran, P. M. (2004). Production of nickel bio-ore from hyperaccumulator plant biomass: applications in phytomining. Biotechnology Bioengineering, 86(3), 243–250.

Brewer, E. P., Saunders, J. A., Angle, J. S., Chaney, R. L., & McIntosh, M. S. (1999). Somatic hybridization between the zinc accumulator Thlaspi caerulescens and Brassica napus. Theoretical and Applied Genetics, 99(5), 761–771.

Briceno, G., Palma, G., & Duran, N. (2007). Influence of organic amendment on the biodegradation and movement of pesticides. Critical Reviews in Environmental Science and Technology, (37), 233-241.

Brooks, R. R., Chambers, M. F., Nicks, L. J., & Robinson, B. H. (1998). Phytomining. Perspectives, 3(9), 359–361.

Brunetti, G., Farrag, K., Rovira, P. S., Nigro, F., & Senesi, N. Greenhouse and field studies on Cr, Cu, Pb and Zn phytoextraction by Brassica napus from contaminated soils in the Apulia region, Southern Italy. Geoderma, (160), 517–523.

Byung, T. O., Patrick, J., Shea R.A., Drijber, G., Vasilyeva, K. & Gautam, S. (2003). TNT biotransformation and detoxification by a Pseudomonas aeruginosa strain. Biodegradation, (14), 309–319.

Chaudhry, Q., Blom-Zandstra, M., Gupta, S., & Joner, E. J. (2005). Utilizing the synergy between plants and rhizosphere microorganisms to enhance breakdown of organic pollutants in the environment, Environmental Science and Pollution Research, (12), 34-48.

Clemens, S., Palmgren, M. G., & Kranmer, U. A. (2002). A long way ahead: understanding and engineering plant metal accumulation. Trends in Plant Science, 7(7), 309–314.

Cohen, M. F., Yamasaki, H., & Mazzola, M. (2004). Bioremediation of soil by plants microbe system. International Journal of Green Energy, 1(3), 301-312.

Cox, L., Walker, A., & Welch, S. J. (1996). Evidence for the accelerated degradation of isoproturon in soils. Pesticide Science, (48), 253-260.

Cuenca, G., & Lovera, M. (1992). Vesicular-arbuscular mycorrhizae in disturbed and revegetated sites from La Gran Sabana, Venezuela. Canadian Journal of Botany, (70), 73-79.

Dalton, H. & Stirling, D. I. (1982). Co-metabolism, Philosophical transactions of the Royal Society. Series B, Biological Science, (297), 481-495.

Das, S., & Singh, D. K. (2006) Purification and characterization of phosphotriesterases from Pseudomonas aeruginosa F10B and Clavibacter michiganense subsp. insidiosum SBL11. Canadian Journal of Microbiology, (52), 157-168

De Schrijver, A, & De Mot, R. (1999). Degradation of pesticides by Actinomycetes. Critical Reviews in Microbiology, (25), 85-119.

Dick, J. (1991). Forest land use, forest use zonation and deforestation in Indonesia: a summary and interpretation of existing information. Background paper to UNCED for the state Ministry for Population and Environment (KLH) and the Environmental Impact Management Agency (BAPEDAL).

Dickinson, N. M., Baker, A. J. M., Doronila, A., Laidlaw, S., & Reeves, R. D. (2009) Phytoremediation of inorganics: realism and synergies. International Journal of Phytoremediation, (11), 97-114.

Dubey, K. ., & Fulekar, M. H. (2013) Rhizoremediation of pesticides: mechanism of microbial interaction in mycorrhizosphere. International Journal of Advancements in Research & Technology, 2 (7), 2278-7763.

Eapen, S., & D’Souza, S. F. (2005). Prospects of genetic engineering of plants for phytoremediation of toxic metals. Biotechnology Advances, 23(2), 97–114.

Eapen, S., Singh, S., & D'Souza, S. F. (2007). Advances in development of transgenic plants for remediation of xenobiotic pollutants. Biotechnology Advances, (25), 442–451.

Eviner, T., & Hawkes, C. V. (2008). Embracing Variability in the Application of Plant–Soil Interactions to the Restoration of Communities and Ecosystems Valerie. Restoration Ecology, 16 (4), 713–729.

Fenlon, K. A., Jones, K. C., & Semple, K. T. (2007). Development of microbial degradation of cypermethrin and diazinon in organically and conventionally managed soils. Journal of Environmental Monitoring, (9), 510-515.

Finley, S. D., Broadbelt, L. J., & Hatzimanikatis, V. (2010). In Silico Feasibility of Novel Biodegradation Pathways for 1,2,4-Trichlorobenzene, BMC Systems Biology, 4.(7), 4-14.

Floriane, S. S., Nicolau, E., Gregory, F., Jouanneau, Y., & Marchal, R. (2009). Biodegradability of 2-ethylhexyl nitrate (2-EHN), a cetane improver of diesel oil. Biodegradation, (20), 85-94.

Fournier, J. C., Soulas, G., & Parekh, N. R. (1996). Main microbial mechanisms of pesticide degradation in soils In: J., Tarradellas, G., Bitoon, & D.L. Rossel, (Eds). Soil ecotoxicology. Pp. 85-115.

French, C. E., Rosser, S. J., Davies, G. J., Nicklin, S. & Bruce, N. C. (1999). Biodegradation of explosives by transgenic plants expressing pentaerythritol tetranitrate reductase, Nature Biotechnology, (17), 491–494.

García, R., & Báez, A. P. (2012). Atomic absorption spectrometry (AAS). In M. A. Farrukh (Ed.), Atomic Absorption Spectroscopy (pp. 1-12).In Tech. absorption-spectrometry-aas.

Gerhardt, K. E., Huang, X. D., Glick, B. R. & Greenberg, B. M. (2009). Phytoremediation and rhizoremediation of organic soil contaminants: Potential and challenges. Plant Science, (176), 20-30.

Giri, K ., & Rai, J.P.N. (2012). Biodegradation of endosulfan isomers in broth culture and soil microcosm by Pseudomonas fluorescens isolated from soil. International Journal of Environmental Studies, 69(5), 729-742.

Giri, K., Mishra, G., Pandey, S., Verma, P. K. Kumar, R., & Bisht, N. S. (2014). Ecological degradation in Northeastern coal fields: Margherita Assam. International Journal of Science, Environment and Technology, 3 (3), (In Press).

Giri, K., Rawat A.P., Rawat, M., & Rai, J.P.N. (2014). Biodegradation of Hexachlorocyclohexane by Two Species of Bacillus Isolated from Contaminated Soil. Chemistry and Ecology, 30 (2), 97-109.

Gleba, D., Borisjuk, N. V., Borisjuk, L. G., Kneer, R., Poulev, A., Skarzhinskaya, M., Dushenkov, S., Logendra, S., Gleba, Y. Y., & Raskin, I. (1999). Use of plant roots for phytoremediation and molecular farming, Proceedings of the National Academy of Sciences U.S.A., 96(11), 5973–5977.

Godt, J., Scheidig, F., Grosse-Siestrup, C., Esche, V., Brandenburg, P., Reich, A., et al. (2006). The toxicity of cadmium and resulting hazards for human health. Journal of Occupational Medicine and Toxicology, (1), 1-22.

Gooddy, D. C., Chilton, P. J. & Harrison, I. (2002). A field study to assess the degradation and transport of diuron and its metabolites in a calcareous soil. Science of Total Environment, (297), 67-83.

Greenberg, B. M. (2006). Development and field tests of a multi-process phytoremediation system for decontamination of soils. Canadian Reclamation, (1), 27-29.

Grime, J. P. (2001). Plant strategies, vegetation processes, and ecosystem Properties. 2nd edition. John Wiley & Sons, New York.

Guerinot, M. L., & Salt, D. E. (2001). Fortified foods and phytoremediation. two sides of the same coin. Plant Physiology, 125(1), 164–167.

Hangler, M., Jensen, B., Ronhede, S., & Sorensen, S. R. (2007). Inducible hydroxylation and demethylation of the herbicide isoproturon by Cunninghamella elegans. FEMS Microbiology Letters, (268), 254-260.

Hawthorne, S. B., Yang, Y. & Miller, D. J. (1994). Extraction of organic pollutants from environmental solids with sub and supercritical water. Analytical Chemistry, (66), 2912-2920.

Hopper, M. L. (1996). Solid phase partition column technology in Emerging strategies for pesticide analysis. In. T. Cairns and J. Sherma (Ed.), CRC Press, Boca Raton, FL, pp. 39-50.

Hrynkiewicz, K., & Baum, C. (2012). The potential of rhizosphere microorganisms to promote the plant growth. In A. Malik, & E. Grohmann (Eds.), Environmental Protection Strategies for Sustainable Development, Strategies for Sustainability (pp. 35-64). Springer Science + Business Media B.V. DOI 10.1007/978-94-007-1591-2-2

Hrynkiewicz, K., Dabrowska, G., Baum, C., Niedojadlo, K., & Leinweber, P. (2012). Interactive and single effects of ectomycorrhiza formation and Bacillus cereus on metallothionein mt1 expression and phytoextraction of Cd and Zn by willows. Water Air Soil Pollution, (223), 957–68.

Hseu, Z. Y., Chen, Z. S., Tsai, C. C., Tsai, C. C., Cheng, S.-F., Liu, C.-L., & Lin, H. T. (2002). Digestion methods for total heavy metals in sediments and soils. Water Air and Soil Pollution, 141(1-4), 189–205.

Huang, X. D., El-Alawi, Y. S., Gurska, J., Glick, B. R., & Greenberg, B. M. (2005). A multi-process phytoremediation system for decontamination of persistent total petroleum hydrocarbons (TPHs) from soils. Microchemistry Journal, (81), 139-147.

Hussain, S., Arshad, M., Saleem, M., & Khalid, A. (2007a). Biodegradation of α and β-endosulfan by soil bacteria. Biodegradation, (18), 731-740.

Hussain, S., Siddique, T., Arshad, M., & Saleem, M. (2009a). Bioremediation and phytoremediation of pesticides: recent advances. Critical Reviews in Environmental Science and Technology, (39), 843-907.

Hussain, S., Sorensen, S. R., Devers-Lamrani, M., El-Sebai, T., & Martin-Laurent, F. (2009b). Characterization of an isoproturon mineralizing bacterial culture enriched from a French agricultural soil. Chemosphere, (77), 1052-1059.

INSA, (2011). Hazardous metals and minerals pollution in india. Indian National Science Academy, Bahadurshah Zafar Marg, New Delhi, Angkor Publishers (P) Ltd., Noida. pp. 1- 24.

Jeneper, M. L., & Hayao, S. (2005). Comparison of the acid combinations in icrowave-assisted digestion of marine sediments for heavy metal analyses. Analytical Science, 21(10), 1181-1184.

Jomova, K., Jenisova, Z., Feszterova, M., Baros, S., Liska, J., Hudecova, D.,et al. (2011). Arsenic: toxicity, oxidative stress and human disease. Journal of Applied Toxicology, (31), 95-107.

Joshi, P. M., & Juwarkar, A. A. (2009). In vivo studies to elucidate the role of extracellular polymeric substances from Azotobacter in immobilization of heavy metals. Environmental Science and Technology, (43), 5884–5889.

Juwarkar, A. A., & Jambhulkar, H. P. (2008). Phytoremediation of coal mine spoil dump through integrated biotechnological approach. Bioresource Technology, (99), 4732–4741.

Khan, A. G. (2005). Role of soil microbes in the rhizospheres of plants growing on trace metal contaminated soils in phytoremediation. Journal of Trace Elements in Medicine and Biology, 18(4), 355–364.

Kuiper, I., Lagendijk, E. L., Bloemberg, G. V., & Lugtenberg, B. J. J. (2004). Rhizoremediation: a beneficial plant–microbe interaction. Molecular Plant Microbe Interaction, (17), 6-12.

Kumari, R., Subudhi, S., Suar, M., Dhingra, G., Raina, V., Dogra, C., Lal, S., Meer, J. R., Holliger, C., & Lal, R. (2002) Cloning and Characterization of lin Genes Responsible for the Degradation of Hexachlorocyclohexane Isomers by Sphingomonas paucimobilis Strain B90. Applied and Environmental Microbiology, (68), 6021-6028,

Lal, R., Dogra, C., Malhotra, S., Sharma, P., and & Pal, R. (2006) Diversity, Distribution and Divergence of lin genes in hexachlorocyclohexane degrading sphingomonads. Trends in Biotechnology, (24), 121-129.

Leisinger, T. Hutter, R. Cook, A. M., & Nuesch, J. (1981). Microbial degradation of xenobiotics and recalcitrant compounds. Academic Press, New York.

Li, W. C., Ye, Z. H., & Wong, M. H. (2010). Metal mobilization and production of short-chain organic acids by rhizosphere bacteria associated with a Cd/Zn hyperaccumulating plant Sedum alfredii. Plant Soil, (326), 453–467.

Lombi, E., Zhao, F. J., Dunham, S. J., & McGrath, S. P. (2001). Phytoremediation of heavy metal contaminated soils: natural hyperaccumulation versus chemically enhanced phytoextraction, Journal of Environmental Quality, 30(6), 1919–1926.

Luo, S. L., Chen, L., Chen, J. I., Xiao, X., Xu, T. Y., Wan, Y., et al. (2011). Analysis and characterization of cultivable heavy metal-resistant bacterial endophytes isolated from Cd-hyperaccumulator Solanum nigrum L. and their potential use for phytoremediation. Chemosphere, (85), 1130-8.

Luo, S., Xu, T., Chen, L., Chen, J., Rao, C., Xiao, X., et al. (2012). Endophyte-assisted promotion of biomass production and metal-uptake of energy crop sweet sorghum by plant-growth-promoting endophyte Bacillus sp. SLS18. Applied Microbiology and Biotechnology, (93), 1745–1753.

Macek, T., Mackova, M., & Kas, J. (2000). Exploitation of plants for the removal of organics in environmental remediation, Biotechnology Advances, (18), 23–34.

Maria, S. D., Rivelli, A. R., Kuffner, M., Sessitsch, A., Wenzel, W. W., Gorfer, M., et al. (2011). Interactions between accumulation of trace elements and macronutrients in Salix caprea after inoculation with rhizosphere microorganisms. Chemosphere, (84), 1256–1261.

Marques, A. P. G. C., Rangel, A. O. S. S., & Castro, P. M. L. (2009). Remediation of heavy metal contaminated soils: phytoremediation as a potentially promising clean-up technology. Critical Reviews in Environmental Science and Technology, 39(8), 622–654.

Mathew, G. A. (2006). Pesticides: Health. Safety and the Environment. Blackwell Publishing.

Meagher, R. B. (2000). Phytoremediation of toxic elemental and organic pollutants, Current Opinion in Plant Biology, (3), 153-162.

Meirer, F., Singh, A., Pepponi, G., Streli, C., Homma T., & Pianetta. (2010). Synchrotron radiation-induced total reflection X-ray fluorescence analysis. Trends in Analytical Chemistry, 29(6), 479–496.

Mohammed, A. S., Kapri, A., & Goel, R. (2011). Heavy metal pollution: source, impact, and remedies. In M. S. Khan, A. Zaidi, R. Goel & J. Musarrat (Eds.), Biomanagement of Metal-Contaminated Soils (pp 1-28). Springer Netherlands.

Nerud, F., Baldrian, P., Gabriel, J. & Ogbeifun, D. (2003). Nonenzymic degradation and decolorization of recalcitrant compounds. In, V., Sasek, J.A., Glaser, P. Baveye, (Eds). Utilization of bioremediation to reduce soil contamination: Problems and solutions. Springer, Dordrecht, Pp. 127-133.

Nieuwenhuize, J., Poley-Vos, C. H., Van den, Akker, A. H., & Van Delft, W. (1991). Comparison of microwave and conventional extraction techniques for the determination of metals in soil, sediment, and sludge samples by atomic spectrometry. Analyst, 116(4), 347–51.

Novick, N. J. & Alexander, M. (1985). Cometabolism of low concentrations of propachlor, alachlor and cycloate in sewage and lake water. Applied and Environmental Microbiology, (49), 737-743.

Pal, R., & Rai, J. P. N. (2010). Phytochelatins: Peptides Involved in Heavy Metal Detoxification. Applied Biochemistry and Biotechnology, 160(3), 945–963.

Parker, M. A., Malek, W., & Parker, I. M. (2006). Growth of an invasive legume is symbiont limited in newly occupied habitats. Diversity and Distributions, (12), 563-571.

Patrick, L. (2006). Lead toxicity part II: the role of free radical damage and the use of antioxidants in the pathology and treatment of lead toxicity. Alternate Medical Revives, (11), 114–27.

Pieuchot, M., PerrinGanier, C., Portal, J. M. & Schiavon, M. (1996). Study on the mineralization and degradation of isoproturon in three soils. Chemosphere, (33), 467-478.

Pilon-Smits, E. (2005). Phytoremediation. Annual Reviews in Plant Biology, (56), 15-39.

Pilon-Smits, E. A. H., Hwang, S., Lytle, C. M., Zhu, Y., Tai, J. C., Bravo, R. C., Chen, Y., Leustek, T. & Terry, N. (1999). Over-expression of ATP sulfurylase in Brassica juncea leads to increased selenate uptake, reduction and tolerance. Plant Physiology, 119(1), 123–132.

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