Soil Microorganisms and Environmental Health

Sayyed & Patel, Int. J.Biotech & Biosci., Vol. 1 (1) : pp 41-66 (2011)

Review Paper

Soil Microorganisms and Environmental Health Sayyed R Z* & Patel P R PG Department of Microbiology, PSGVP Mandal‟s S I Patil Arts, G B Patel Science and STSKVS Commerce College, Shahada, Maharashtra, 425 409, India. E-mail:[email protected] Abstract Over the past century, mining, agriculture, manufacturing and urban activities have largely contributed to extensive contamination of soil and water with several range of pollutants. Rapid development of many industries such as mining, energy, fuel producing, fertilizer, pesticide, metallurgy, iron and steel electroplating, electrolysis, electro-osmosis, leather, photography, appliance manufacturing, metal surface treating and aerospace and atomic energy installations, heavy metal waste are directly discharged in biosphere causing serious environmental pollution and even threatening human life. Bacteria thriving in the environment are exposed to a range of physical and chemical signals that need to be processed to achieve a positive or negative physiological response. Microorganisms have evolved towards ecological fitness rather than biotechnological efficiency; thus, it would take a long time for bacteria capable of cleaning up anthropogenic pollution to evolve by natural selection. Hence, studying the physiology, biochemistry and genetics of the catabolic pathways becomes crucial to recreate and accelerate natural processes in the test tube as well as to accomplish their rational manipulation to design more efficient biocatalysts for different biotechnological applications. Mixed microbial communities have the most powerful biodegradative potential because the genetic information of more than one organism is necessary to degrade the complex mixtures of organic compounds present in contaminated areas. The genetic potential and certain environmental factors such as temperature, pH, and available nitrogen and phosphorus sources, therefore, seem to determine the rate and the extent of degradation. Most of the microorganisms which exhibit PGPR activities belong to the Gram-negative group and among these fluorescent Pseudomonads are the most widely studied. The siderophores of fluorescent Pseudomonads and their introduction into the rhizosphere have been widely explored. Siderophore production in iron stress conditions confers upon these organisms an added advantage, resulting in exclusion of pathogens due to iron starvation. However, use of antagonistic rhizobia has an added advantage in that they have also the ability to fix nitrogen. Different strains of rhizobia have now been reported to produce siderophores. Strains of root nodulating bacteria have also been reported to produce phytohormones like indole acetic acid (IAA), auxin, gibberellin, ethylene etc. and antibacterial compounds like rhizobiotoxin, bacteriocins, siderophores and HCN. Phosphate solubilizing bacteria (PSB) are beneficial microorganisms in the plant rhizosphere, as they solubilize bound phosphorus (P) and increase their availability for the plant This ability also confers upon nodule bacteria a selective advantage and may lead to both direct and indirect control of plant pathogens. Keywords : PGPR; Siderophores, EPS, Petroleum products; Aromatic hydrocarbon; Bioabsorption; Heavy Metal, bioabsorption

1. Introduction 2. Degradation of chemical compounds   Degradation of petroleum products   Degradation of n-alkanes

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Degradation of alkenes  Degradation of cycloaliphatic compounds  Degradation of aromatic hydrocarbon  Degradation of Benzene  Degradation of chlorinated organic compound  Dehalogenation of aromatic compounds   Oxidative dehalogenation   Hydrolytic dehalogenation   Reductive dehalogenation  Degradation of halogenated aromatic compounds  Degradation of halogenated benzoic acids  Degradation of halogenated phenols  Degradation of halogenated phenoxyacetic acid  Degradation of halogenated anilines  Degradation of halogenated benzenes  Degradation of halogenated biphenyls  Degradation of halogenated dibenzo-p-dioxins and dibenzofurans Bioremediation of heavy metal  Microbiology and biochemistry of heavy metal resistance  Mechanisms of bacterial metal resistance  Mechanism of uptake  Mechanism of efflux  Plasmid resistance  Metallothioneins and binding  Metal binding peptides  Genetic engineering for specific protein  Genetically engineered microorganisms and bioremediation of heavy metal Biosorption by bacteria  Biosorption by Fungi  Biosorption by algae and moss What are PGPR?  Secondary metabolites  Mechanisms of PGPR in Plant Growth Promotion  Phosphate Solubilizers  Mechanisms of phosphate solubilization    Plant growth regulators (hormones)  HCN production  Production of extacellular polymorphic substances (EPS)  Production of siderophores Conclusion Future strategy References 

1. Introduction Bioremediation can be defined as any process that uses microorganisms or their enzymes to return the environment altered by contaminants to its original condition. Bioremediation may

be employed in order to attack specific contaminants, such as chlorinated pesticides that are degraded by bacteria, or a more general approach may be taken, such as oil spills that are broken down using multiple techniques including the addition of fertilizer to facilitate the decomposition of crude oil by bacteria. Not all contaminants are readily treated through the use of bioremediation; for example, heavy metals such as cadmium and lead are not readily absorbed or captured by organisms. The integration of metals such as mercury into the food chain may make things worse as organism bioaccumulates these metals. Major sources of xenobiotic compounds enters into the environment are (i) chemical and pharmaceutical industries that produce a wide array of xenobiotics and synthetic polymers, (ii) pulp and paper bleaching, which are the main sources of natural and man made chlorinated organic compounds in the environment; (iii) mining, which releases heavy metals into biogeochemical cycles; (iv) fossil fuels (coal and petroleum), which may be accidentally released in large amounts into the ecosystem (oil spills) (v) intensive agriculture, which releases massive amounts of fertilizers, pesticides, and herbicides. These are some of the examples through which xenobiotic compounds enter into the environment. Due to their potential toxicity to both wildlife and humans, several persistent organic pollutants (POPs) have now been totally banned from production and use in many countries around the world. Since agricultural fields due to the uncontrolled use of chemical pesticides and fertilizers are most contaminated, search for Plant Growth Promoting Rhizobacteria (PGPR) having potential of adsorbing heavy metals from field will have twin advantage of bioremediation and plant growth promotion (Tripati & Srivastava 2007).However, considerable attention has been paid to PGPR, as the best alternative to chemicals, to facilitate eco-friendly biological control of soil and seed-borne pathogens (Whipps 2001). One of the most important components of bio-intensive pest management is the suppression of insect pest by employing pathogen like bacteria, virus, and fungus, which was designated as microbial pesticides (Chatterjee 2009). The biological control of plant diseases with bacterial antagonism is a potential alternative of chemical control as it is expensive and also result in accumulation of toxic compounds in soil biota (Ghai et al 2007). Every year, severe global economic losses to agricultural crops are encountered due to plant diseases caused by more than sixty pathogens leading to the loss of 30% crop yield amounting 416 Million US dollars (Nehl et al 1996). Indiscriminate use of synthetic insecticides to control insect pests of various crop plants has resulted in environmental pollution, contamination of ground water and the development of pesticide resistance in insect pests. Therefore the development and use of biodegradable alternate plant protection agents including Bacillus thuringiensis (Bt) has become a necessity in integrated pest management program (IPM) of all the major crop plants. The use of „cry‟ gene based bioinsecticide produced by B. thuringiensis (Bt) offers advantages over harmful chemical insecticides because they are target specific and economical. Further Bt formulations, being proteinaceous in nature, are easily degraded and persist very little in soil, leading to negligible effect on soil health. (Singh & Boora 2007). Work on insect viruses in India was initiated as early as1968 with the report of nuclear polyhedrosis virus (NPV) from Helicoverpa armigera, a pest of national importance and Spodoptera litura, a polyphagous pest attacking several crops. Since then studies on insect viruses have progressed rapidly and several

viruses were reported to occur in insect pests, mostof them from the order Lepidoptera. These comprises of nuclear polyhedrosis virus (NPV), granulosis virus (GV) and cytoplasmic polyhedrosis virus (CPV) (Saxena 2008). Worldwide contamination of soil with metals has posed a great threat to the human health as most of them are proven to be carcinogenic even at slightly higher concentration. It is estimated that about $3 billion are needed to remediate the metal contaminated sites alone in USA (Kamaludeen & Ramasamy 2008). However, there are a number of advantages to bioremediation, which may be employed in areas which cannot be reached easily without excavation. For example, hydrocarbon spills (or more specific: gasoline) may contaminate groundwater well below the surface of the ground; injecting the right organisms, in conjunction with oxygen-forming compounds, may significantly reduce concentrations after a period of time. This is much less expensive than excavation followed by burial elsewhere or incineration, and reduces or eliminates the need for pumping and treatment, which is a common practice at sites where hydrocarbons have contaminated groundwater. Generally, bioremediation technologies can be classified as in situ or ex situ. In situ bioremediation involves treating the contaminated material at the site while ex situ involves the removal of the contaminated material to be treated elsewhere. Some examples of bioremediation technologies are bioventing, land farming, bioreactor, composting, bioaugmentation and biostimulation. A substance foreign to biological system is known as xenobiotic compound. Most of the xenobiotic compounds are degraded by microorganism may be defined as weak xenobiotic, however, few of them may persist longer in the environment and not easily degraded is known as recalcitrant compound. A xenobiotic is a chemical which is found in an organism but which is not normally produced or expected to be present in it. It can also cover substances which are present in much higher concentrations than are usual. Xenobiotic compounds are often toxic to life and are also often hard for microorganisms to metabolize (because they contain molecular arrangements that not normally encountered in nature). Thus, many of these compounds can accumulate in the environment and continue to be a hazard for many years. Microbes (bacteria and fungi) found in natural waters and soils have a very broad ability to utilize (catabolise) virtually all naturally occurring compounds as their sources of carbon and energy, thus recycling the fixed organic carbon back into harmless biomass and carbon dioxide. This capability of microbes has evolved over 3 billion years of the planets history and is responsible for the balance between photosynthesis (by plants and algae), fixing carbon dioxide into biomass, and respiration (by animals and bacteria), and converting organic compounds back to carbon dioxide by oxidation through natural detoxification processes. Recent advances in genetics and molecular biology in the last 20-30 years has shown that bacteria are genetically extremely adaptable, and in addition to the advantage due to their rapid growth rates, have a range of mechanisms which enable them to adapt to new environments. When compounds are persistent in the environment, their biodegradation often proceeds through multiple steps utilizing different enzyme systems or different microbial populations (Thakur 2007). 2. Degradation of chemical compounds

There are three requirements for biodegradation such as capable organisms, synthesis of requisite enzymes, suitable environmental conditions. In presence of recalcitrant compounds, the degradation initiated by induction of enzyme synthesis can be a problem, therefore, coordinate induction (through consortia of organisms) and co-metabolism i.e. transformation/degradation of a non-growth substrate in the obligate presence of another growth substrate are used. Degradation

of petroleum products

The most common sources of petroleum contamination are leaks in piping, leaks from corroded tanks, various equipment failures, overfill and spill while filling tanks and sudden accidental leakage. The petroleum products contaminate soil, ground water, surface water and air. The main components present in petroleum products are hydrocarbons, a significant portion of compounds contains nitrogen, sulphur and oxygen (N S O) and aromatic compounds. Some metals are also present. In crude oil, chemical compounds may be divided into aliphatic hydrocarbon (e.g. alkanes, branched alkanes and alkenes), cycloaliphatic hydrocarbons (e.g. cyclohaxane), aromatic hydrocarbons and N S O compounds. These compounds entered in the environment not only from petroleum products but also agricultural, industrial, commercial, municipal, transportation and related activities. Some of them also formed unintentionally in the environment such as dioxin–like compounds formed in industrial and municipal sludge and in atmosphere. The compounds are degraded by microorganism contains catabolic enzymes and genes present either in plasmid or genomic DNA or both. Degradation

of n-alkanes

Alkanes originate from both biogenic and anthropogenic sources. Anthropogenic sources of nalkanes include incomplete fossil fuel combustion, lubricant oils and biomass burning. Wind erosion of leaf epicuticular waxes, direct suspension of pollen, vegetation debris and microbial degradation are considered as the most important natural sources of particulate nalkanes. Longchain n-alkanes (C10 to C24) are degraded most rapidly. Short-chain alkanes (less than C9) are toxic to many microorganisms, but they evaporate rapidly from petroleum contaminated sites. Pseudomonas maltophilia, P. putida, Pseudomonas sp. strain C12B, Burkholderia cepacia, Acinetobacter sp. strain ADP1 and Acinetobacter sp. strain M1 and Nocardiodes sp. strain CF8 are capable of efficient degradation of alkanes. Oxidation of alkanes is classified as being terminal or diterminal. The monoterminal oxidation is the main pathway. It proceeds via the formation of the corresponding alcohol, aldehyde, and fatty acid.ß-Oxidation of the fatty acids results in the formation of acetyl-CoA. n-Alkanes with an uneven number of carbon atoms are degraded to propionyl-CoA, which is in turn carboxylated to methylmalonyl-CoA and further converted to succinyl-CoA. Fatty acids of a physiological chain length may be directly incorporated into membrane lipids, but the majority of degradation products are introduced into the tricarboxylic acid cycle. The subterminal oxidation occurs with lower (C3 to C6) and longer alkanes with the formation of a secondary alcohol and subsequent ketone. n-Alkanes are readily degraded in both laboratory culture and the natural environment. The metabolic pathway initiated by a hydroxyloase (mono-oxygenose) enzyme to produce the corresponding alkane –1–ol:

R – CH3 + O2 + NAD(P)H + H+ ‟! R – CH2OH + NAD (P+) + H2O Degradation

of alkenes

Unsaturated 1- alkenes are oxidized at the saturated end of the chains. A minor pathway has been shown to proceed via an epoxide, which is converted to a fatty acid. Branching, in general, reduces the rate of biodegradation. Metabolic attack on saturated aliphatic hydrocarbons may be initiated either via attack on the double bond or by the same mecha-nism employed in nalkane metabolism. Four main patterns of initial attack may be recognized:

Fig. 1: Degradation of n-alkane by monooxygenase enzyme initially to alcohol, which is converted to related aldehyde and acid, and finally to carbon dioxide

Fig. 2: Basic metabolic pathways involved in the metabolism of alkenes 1. Oxygenase attack upon a terminal methyl group to produce corresponding alkene-1-ol, 2. Subterminal oxygenase attack to produce an alkenol with the hydroxyl group at a non terminal carbon, 3. Oxidation across the double bond to give an epoxide, 4. Oxidation across the double bond to produce a diol.

Many microorganisms display more than one of these pathways. Alkynes can also be degraded aerobically but have been little studies. It is thought that hydratases are responsible for the initial metabolic attack. Degradation

of cycloaliphatic compounds

Cycloaliphatic compounds are present in large proportion in petroleum product and some oils. Cometabolism of cycloaliphatics has been reported using whole cells and cell extracts of alkanes and other hydrocarbon degraders. The initiation of co-metabolic involves the conversion of cycloaliphatics to alcohol or ketones by low specificity monooxy-genase enzymes. The mechanism of cyclohexane degradation is shown in Figure 3. Degradation

of aromatic hydrocarbon

Aromatic hydrocarbons are ubiquitous in nature. The commercial, industrial, natural activities generate huge amount of aromatic hydrocarbons causes a great concern due to potential hazard to flora and fauna. Benzene, toluene, ethyl benzene, styrene and the xylenes are among the 50 largest-volume industrial chemicals produced with production figures of millions of tones per year. Diversity of aerobic and anaerobic biodegradation of aromatic compounds has several common features. Aromatic compounds degradation is initiated by conversion of complex aromatic compounds to a “starting substrate”. The initial reaction may involve the introduction of oxygen atoms by a monooxygenase and then formation of catechol. This phenomenon occurs in the aerobic catabolic funneling, and then most peripheral pathways involve oxygenation reactions carried out by monooxygenases and/or hydroxylating dioxygenases that generate dihydroxy aromatic compounds (catechol, protocatechuate, gentisate, homoprotocatechuate, homogentisate, hydroquinone, hydroxyquinol). These intermediate compounds are the substrates of ring-cleavage enzymes that use molecular oxygen to open the aromatic ring between the two hydroxyl groups (ortho cleavage, catalyzed by intradiol dioxygenases) or proximal to one of the two hydroxyl groups (meta cleavage, catalyzed by extradiol dioxygenases). A dioxygenase breaks open the aromatic ring of catechol, producing cis,cismuconate, an unsaturated dicarboxylic acid. This product is then oxidized to acetyl-CoAs by the aforementioned beta-oxidation path. Catabolic plasmids occur naturally in aromatics hydrocarbons such as camphor, naphthalene, salicylate and other compounds. Most of the catabolic plasmids are self-transmissible and have a broad host range. Degradation

of benzene

There are few reports on the bacterial degradation of benzene. The excellent studies carried out in the previous three decades elucidated the pathway identified the intermediates and characterized the enzyme systems. The first step of benzene oxidation is a hydroxylation catalyzed by a dioxygenase. The product, a diol, is then converted to catechol by a dehydrogenase. These initial reactions, hydroxylation and dehydrogenation, are also common to pathways of degradation of other aromatic hydrocarbons. The introduction of a substituent group onto the benzene ring renders alternative mechanisms possible to attack side chains or to oxidize the aromatic ring.

Fig. 3: Metabolic pathway involved in the degradation of cyclohexane. The hydroxylase enzyme responsible for the initial metabolic step was detected in Pseudo-monas sp.

Fig. 4 . Degradation of aromatic compounds by funneling pathway and central pathway

Fig 5. Monooxygenase and dioxygenase reaction in degradation of benzene 

Degradation of chlorinated organic compound

The recalcitrance of organic pollutants increases with increasing halogenation. Substitution of halogen as well as nitro and sulfo groups at the aromatic ring is accomplished by an increasing electrophilicity of the molecule. These compounds resist the electrophilic attack by oxygenases

of aerobic bacteria. Halogenated organic compounds constitute one of the largest groups of environ-mental chemicals including pesticides. The industrial production of new halogenated organic compounds has increased throughout the last century and these compounds are integral to a variety of industrial applications. A critical step in the degradation of organohalides is cleavage of the carbon-halogen bond, and microorganisms have evolved a variety of metabolic strategies for dehalogenation. The natural production and anthropogenic release of halogenated hydrocarbons into the environment has been the likely driving force for the evolution of an unexpectedly high microbial capacity to dehalogenate in presence of different classes of xenobiotic haloorganics.  Dehalogenation of aromatic compounds There are three classes of dehalogenation

Oxidative dehalogenation

In this process the halogen substituents are lost fortuitously during oxygenation of the ring. A number of proteobacterial pure culture isolates of the genera Thauera, Pseudomonas, and Ochrobacterium completely mineralized chlori-nated aromatic compound as a sole source of carbon and energy conditions has been repeatedly observed for mixed cultures.

Fig. 6. The biodegradation route of benzene. Benzene is converted into catechol via cis-Benzene dihydrodiol to catechol which is finally degraded by ortho (II) or meta ring cleavage (I)

Fig 7. Oxidative dehalogenation processes



Hydrolytic dehalogenation

In this case a hydroxyl group specifically replaces halogen substituent. The source of the oxygen atom in the hydroxyl group is water instead of oxygen. This reaction can occur under both aerobic and denitrifying conditions. e.g. dehalogenation of 4-chlorobenzoate to form 4hydroxybenzoate by Arthrobacter and Pseudomonas.

Figure 8: Hydrolytic dehalogenation processes 

Reductive Dehalogenation

Microbial means to dehalogenate organohalides under anaerobic conditions by a reductive mechanism can be largely divided into abiotic, or cometabolic, and metabolic conversion. While the latter is found in halorespiring bacteria (HRB), which couple the reductive dehalogenation reaction by specific, high-affinity biocatalysts to microbial growth, the former is proposed to be catalyzed mostly by metal ion-containing heatstable tetrapyrroles or enzymes, in which these compounds are incorporated as cofactors. More than a decade ago, the isolation and characterization of the - proteobacterium Desulfomonile tiedjei, able to couple the reductive dehalogenation of 3- chlorobenzoate to energy conservation, set the stage for the unraveling of this novel type of energy metabolism in anaerobic microorganisms.

Fig. 9. Reductive dehalogenation reaction Degradation

of halogenated aromatic compounds

Chlorinated aromatic compounds are widely used as pesticides, industrial applications, and produced unintentionally as trace contaminants during the industrial production of chlorinated compounds and incineration of chlorine-containing waste. Brominated aromatic compounds have found as flame-retardants, and florinated and iodinated aromatic compounds have pharmaceutical applications. The chemical inertness and hydrophobicity of these compounds has resulted in widely distribution in the environment. The important compounds are halogenated benzoic acid halogenated benzene, halogenated phenols, halogenated anilines,

halogenated phenoxyacetic acids, halogenated biphenyls and halogenated dibenzo-pdioxins and dibenzofurans. Degradation

of halogenated benzoic acids

Aerobically, halogenated benzoates are degraded by initial dioxygenation of the aromatic ring to yield halocatechol. Ring cleavage of these compounds takes place most efficiently by othro, or â-ketoa-dipate, route yielding halo-cis, cis-muconates. Meta, or extradiol, cleavage of halocatechol is formed by catechol- 2,3-dioxygenase to yield degradable metabolites. Meta cleavage of 3-halocatechols gives productive halogenated semialdehydes and 5-chloroformyl-2hydrooxypenta-2,4-dieonates. The latter compounds probably find irreversibly to basic groups of the dioxygenese, which results in inactivation of the enzyme. Under anaerobic conditions halobenzoates are reductively halogenated. Genes for the complete mineralization of 3-chlorobenzoate via the modified ortho cleavage pathway are known to be located on the plasmid pWR1 (pB13) from Pseudomonas sp. strain B13 of molecular size 11 kb, plasmid pAC25 or pAC27 from Pseudomonas putida AC and on the plasmid pJP4 from Alcaligenes eutrophus JMP134.

Fig. 10. Aerobic degradayion of chlorobenzoate to chlorocatechol. The enzymes benzoate dioxygenase (A) and benzoate dihydrodioldehydrogenase (B) involved in the reactions Degradation

of halogenated phenols

Chlorinated phenols are used on large scale as wood preservatives, fungicides, herbicides, and general biocides. They may be mono, di, tri, tetra, penta and hexa chlorophenols. The biodegradation of chlorophenols takes place by three main pathways. In general mono and dichlorophenols are converted into chlorocatechol by monooxygenase, whereas higher chlorinated phenols are hydroxylated to form chlorinated hydroquinones. Under anaerobic conditions, chlorophenols undergo initial reductive dechlorination. The first three enzymes of the pentachlorophenol (PCP) degradation pathway in Sphingobium chlorophenolicum (formerly Sphingo-monas chlorophenolica) have been characterized, and the corresponding genes, pcpA, pcpB, and pcpC, have been individually cloned and sequenced. To search for new genes involved in PCP degradation and map the physical locations of the pcp genes, a 24-kb fragment containing pcpA and pcpC was completely sequenced. The four gene products PcpB, PcpC, PcpA, and PcpE were responsible for the metabolism of PCP to 3- oxoadipate.

Fig. 11. Metabolic route for aerobic degradation of chlorocatechols. A, catechol-2,3-dioxygenase; B, catechol-1,2-dioxygenase; C, muconate cycloisomerase; D, dienelactone hyrolase, and E, maleylacetate reductase involved in degradation of chlorocatechol. Degradation

of halogenated phenoxyacetic acid

The herbicides 2, 4 dicholorphenoxy acetic acid (2,4 D) and 2, 4, 5- trichlorophenoxyacetic acid (2, 4, 5-T) have been used for more than 40 years. In general, biodegradation of 2,4-D takes place via initial cleavage of other bond, followed by hydroxylation of the resulting dichlorophenol to chlorocatechols. The chlorocatechol intermediates formed from the chlorinated compounds are degraded by the ortho cleavage. The genes encoded this pathway are located on a 80kb plasmid. The genes for 2,4-D degradation is present in three operon: tdf A gene coding for 2,4dicholophenoxyacetic acid monoxygenese, the tdf B gene coding for 2,4-dichlorophenol hydroxylase and the tdf CDEF genes coding for the modified ortho cleavage pathway. Degradation

of halogenated anilines

Chlorinated anilines are used as intermediates in the synthesis of pesticides. Degradation of these compounds proceeds by initial dioxygenation to form holocatechols catalyzed by an aniline oxygenase. Further degradation is via the modified ortho cleavage pathway. Reductive dehalogenation of chloroanilines takes place under anaerobic conditions. Degradation

of halogenated benzenes

Chlorobenzenes are used as solvents, fumigants and as intermediates in dye production and pesticides. Chlorobenzene is degraded via dioxygenation to form chlorocatechol that is further degraded by the ortho cleavage pathway. The genes involved in degradation of chlorobenzene in strain P51 are located in two clusters on a 110kb plasmid. The tcbA and tcbB genes are encoding the degradation of chlorocatechol present on a transposable element. The second cluster consists of tcbC, tcbD and tcbE genes, coding for type II catechol 1,2-dioxygenase, cycloisomenase and hydroxylase enzymes. The transposable elements may have played a role in the transfer of the first two genes and the evolution of this catabolic pathway.

Degradation

of halogenated biphenyls

Polychlorinated biphenyls are used as non-inflammable heat transfer finds as dielectric fluids in capacitors and transformers, hydraulic fluids and plasticizers in paints. The ability to degrade PCB is found in several genera of Gram positive and Gram negative aerobic bacteria. In most cases, chlorobenzoates accumulate as end products. Biphenyl degraded by meta cleavage of the resulting chlorocatechol can lead to toxic end products, only when enzymes for the ortho cleavage route of chlorocatechols are induced then it involved in the degradation. The enzymes involved in the degradation of chlorinated biphenyls to benzoate are encoded in genes located on both on chromosomes and on plasmids. At least four genes are involved in the degradation of PCBs. Gene bphA codes for the biphenyl dioxygenase, gene bphb encodes the dihydrodial dehydro-xygenase, gene bphc encodes the 2,3-dihydro-xybipheyl dioxygenase and gene bphd codes for the 2-hydroxy-6-oxo-6 phenyl hexa– 2,4-dieonate hydrolase. The evidences for anaerobic degradation of PCBs by dechlorination have been reported. Degradation

of halogenated dibenzo-p-diox-ins and dibenzofurans

Dioxin like compound is formed unintentionally in the environment that is in highly toxic and recalcitrant. Very slow oxidative degradation of 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin (2,3,7, 8TCDD) has been reported. Dibenzo-p-dioxin-utilizing bacteria, Sphingomonas sp. Strain RW1 was isolated from enrichment cultures inoculated with water samples from the river Elbe. The isolate grew with both the biaryl ethers dibenzo-p-dioxin and dibenzofuran (DF) as the sole sources of carbon and energy. Biodegradation of the two aromatic compounds initially proceeded after an oxygenolytic attack at the angular position adjacent to the ether bridge, producing 2,2',3- trihydroxydiphenyl ether or 2,2',3-trihydroxybiphenyl from the initially formed dihydrodiols, which represent extremely unstable hemiacetals. A key enzyme in the degradation pathways of dibenzo-p-dioxin and dibenzofuran, namely, 2,2',3trihydroxybiphenyl dioxygenase, which is responsible for meta cleavage of the first aromatic ring, has been genetically and biochemically analyzed. The dbfB gene of this enzyme has been cloned from a cosmid library of the dibenzo-p-dioxin- and dibenzofuran-degrading bacterium Sphingomonas sp. strain RW1. 3. Bioremediation of heavy metals Heavy metals are metals with a density above 5 g/cm3 (Nies 1999). There are 53 metals with a density above 4-5 g/cm3 reported as heavy metals, but it‟s better to consider them from their physiological effects and toxicity. At high concentrations, metals form unspecific complex compounds in the cell, which lead to various toxic effects depending on the metal and the micro-organism considered. Nevin et al (2003) have observed by studies with highly saline uranium contaminated aquifer sediment the demonstrated that the addition of acetate could stimulate the removal of uranium from the ground water; this removal was associated with enrichment in microorganism most closely related to Pseudomonas and Desulfosporosinus species. Kashefi and Lovley (1999) had proposed that reduction of toxic metals with hyperthermophilic microorganisms (Thermotoga maritime and Pyrobaculum islandicum) or their enzymes might be applied to the remediation of metal contaminated waters or waste streams. De et al (2008) isolated mercury resistant bacteria from various locations along the Indian coast. Seawater nutrient agar (SWNA, Himedia Laboratories Pvt. Ltd.,India) was used for their isolation and

enumeration. They observed the growth on SWNA plates with 25ppm mercury and were termed bacteria highly resistant to mercury. Microorganisms including bacteria, algae, fungi and yeast are found to be capable of efficiently accumulating heavy metals (Yan & Viraraghavan 2000). The solubilization of metals from soil, sludge, or solid waste can be done via autotrophic or heterotrophic leaching, the use of metallophores, or by chemical leaching followed by microbial treatments. Metal displaced in this way into water phase or metals already available in waste water can be desolubilized via biologically induced adsorption, precipitation, and transformation or complexation processes. On other hand, the degradation of organic pollutants can be inhibited by the presence of some heavy metal. This problem can be solved by an increase of the heavy metal resistance of the bioremediating system ( Diels 1997). The determination of trace metal concentration is a common procedure in environmental and biological sciences. Analytical methods available include atomic absorption spectroscopy (AAS), inductively coupled plasma spectroscopy (ICP), neutron activation analysis (NAA), emission spectroscopy, and X- ray fluorescence (O‟Halloran et al 1997). Microbiology and biochemistry of heavy metal resistance Metals play an integral role in the life processes of microorganisms. Some metals, such as calcium, cobalt, chromium, copper, iron, potassium, magnesium, manganese, sodium, nickel and zinc, are essential, serve as micronutrients and are used for redox-processes; to stabilize molecules through electrostatic interactions; as components of various enzymes; and for regulation of osmotic pressure. Many other metals have no biological role (e.g. silver, aluminium, cadmium, gold, lead and mercury), and are nonessential and potentially toxic to microorganisms. Toxicity of nonessential metals occurs through the displacement of essential metals from their native binding sites or through ligand interactions. For example, Hg 2+, Cd2+ and Ag2+ tend to bind to SH groups, and thus inhibit the activity of sensitive enzymes (Nies 1999). In addition, at high levels, both essential and nonessential metals can damage cell membranes; alter enzyme specificity; disrupt cellular functions; and damage the stucture of DNA. To have a physiological or toxic effect, most metal ions have to enter the microbial cell. Many divalent metal cations (e.g. Mn2+, Fe2+, Co2+, Ni2+, Cu2+ and Zn2+) are structurally very similar. Also, the structure of oxyanions such as chromate resembles that of sulfate, and the same is true for arsenate and phosphate. Thus, to be able to differentiate between structurally very similar metal ions, the microbial uptake systems have to be tightly regulated. Usually, microorganisms have solved this problem by using two types of uptake systems for metal ions. One is fast, unspecific, and driven by the chemiosmotic gradient across the cytoplasmic membrane of bacteria . Since this mechanism is used by a variety of substrates, it is constitutively expressed (Nies 1999). The second type of uptake system has a high substrate specificity, is slower, often uses ATP hydrolysis as the energy source and is only produced by the cell in times of need, starvation or a special metabolic situation. Even though microorganisms have specific uptake systems, high concentrations of nonessential metals may be transported into the cell by a constitutively expressed unspecific system. This “open gate” is the one reason why metal ions are toxic to microorganisms (Nies 1999). As a consequence, microorganisms have been forced to develop metal-ion homeostasis factors and metalresistance determinants. Because metal ions cannot be degraded or modified like toxic organic

compounds, there are six possible mechanisms for a metal resistance system: exclusion by permeability barrier; intra- and extra-cellular sequestration; active efflux pumps; enzymatic reduction; and reduction in the sensitivity of cellular targets to metal ions. One or more of these resistance mechanisms allows microorganisms to function in metal contaminated environments. Mechanisms of bacterial metal resistance There are 4 mechanisms of bacterial metal resistance 1. Keeping the toxic ion out of the cell (reduced uptake) 2.

Highly-specific efflux pumping (i.e. removing toxic ions that entered the cell by means of transport system evolved for nutrient cations or anions). Efflux pumps can be either ATPases or cheiosmotic driven. ATPases are enzymes that use the chemical energy from cleavage of the high-energy phosphodiester bond of ATP to drive the formation of concentration gradients.

3.

Intra or extracellular sequestration by specific mineral-ion binding components (e.g. metallo-thioneins) and/or segregation into complex compounds.

4.

Enzymatic detoxification (oxydoreductions) which convert a more toxic ion to a less toxic one.

The first two mechanisms can be grouped under the term avoidance, whereas the last two are known as sequestration mechanisms. It is important to notice that mechanisms of metal tolerance might be adaptations of the processes of normal homeostasis. Homeostasis describes the fact that cells have processes to monitor and maintain intracellular concentrations of metals. So, metal homeostasis must involve uptake of sufficient essential metals while providing protection against their toxicity. Mechanism of uptake The first barrier to penetration is the wall, which provides some protection for the cytoplasmic membrane. Cell walls, especially those of fungi, can be used as biosorbents. However, walls cannot act as a perfect barrier to entry of some ions that are essential trace elements for micro-organisms. So, the metal ion is first transported into the cytoplasm in spite of its high concentration, which is the first reason why metal ions are toxic. Mechanism of efflux Efflux pumps reduce the intracellular concentration of metals by means of transport systems, without any enzymatic transformation. This mechanism is more widespread than enzymatic detoxification. Uptake and efflux mechanisms can be classified in 8 protein families approximately: the most important are the ABC family (ATP Binding Cassette), the P- and Atype ATPases family, the RND family (Resistance, Nodulation and cell Division) and the MIT family (Metal Inorganic Transport). Plasmid resistance

Bacterial plasmids contain genes that provide extra functions to the cells, among which resistances to toxic metals. Plasmids are small circular DNA molecular that can move from one cell to another. Thus, the transfer of toxic metal resistance from one cell to another is facilitated. This is why, most of the time, resistance systems are found on these plasmids, but some systems are determined by chromosomal genes in other organisms. Plasmid-determined resistance systems are very specific. The genetic material involved can be exchanged using in vivo as well as in vitro techniques. In vivo genetic exchange techniques make use of the natural genetic engineering capacities of the bacteria and of the fact that degradation and metal resistance markers are often encoded by transposable elements and transferable plasmids. In vivo techniques have the advantage that they are not considered as genetic manipulation senso strictu (no use of recombinant DNA technology) and that the legislation concerning genetically engineered organisms need not be taken into account. The genetic tools provided by the bacteria may, however, be limited. In vitro techniques make use of metal resistance, organic xenobiotic degradation determinants, and other markers cloned on specialized vehicles (cloning vectors, mini-transposable elements) for genetic exchange and stable inheritance of the markers. Moreover, the marker may have been genetically manipulated to ameliorate its expression and to extend its expression host range. When the genetic exchange of markers is considered, the expression host range of the markers and the host and transfer range of the replicon involved have to be taken into account. In most cases, not much data exist and one has to work emipirically. The phylogenetic distance between the donar bacterium and the recipient bacterium bearing the marker can be considered but may not be sufficient. Moreover, under selective pressure, genetic rearrangements and mutations can occur in the recipient strains that may alter the expression range of marker. Plasmids currently used in conjugation procedures to construct strains for bioremediation belong mainly to IncP and IncQ groups. IncP plasmids are best known and the most used broad host-range conjugative plasmids. They are remarkable for very high frequency of self transfer that often can rise up to 0.1-1 transconjugation/donar cell even in interspecific matings. Of special interest are the natural IncP plasmids carrying catabolic genes(degradation of 2,4-D, biphenyl, 3-chlorobenzoate). IncP plasmids display also the property of capturing genes from the mating partner into the plasmid-bearing donor. This process implies the formation of mating aggregates, occurs at very substantial frequencies, and may have important ecological significance. IncQ plasmids are much smaller (near about 9 kb) and are not self transferable, but are easily mobilized by IncP plasmids in a variety of recipients far exceeding the limits of Proteobacteriae. Their replication range is very wide. (Mergeay & Springael 1997) Metallothioneins and binding Bacterial metallothioneins (MT) are commonly grouped in 3 classes. Class I and II are gene encoded, whereas class III is not. Metallothioneins are low molecular mass, cysteine-rich metalbinding proteins, homologous to mammalian metal binding proteins. They have bind divalent cations, such as Ca2+, Cu2+ and Zn2+ so tightly that harm is avoided (sequestration). Metal binding peptides Enhance bioabsorbents has been developed by creating a repetitive metal-binding motif consisting of (Glu-Cys) Gly. These peptides emulate the structure of phytochelatin, metal chelating molecules that play a major role in metal detoxifications. The phytochelatin analogues

were presented on bacterial surfaces, enhancing Cd2+ and Hg2+ bioaccumulation by 12 fold to 20 fold respectively. Genetic engineering for specific protein The highly specific metal absorbing proteins is the result of a clearly designed genetic circuit that is tightly under the control of gene which are regulatory proteins used for controlling the expression of enzyme. The high affinity and selectivity of MerR toward mercury has been exploited for the construction of microbial biosorbent specific for mercury removal. Presence of surface-exposed MerR on an engineered strain enabled 6 fold higher Mg2+ binding. Similarly, cell overexpressing ArsR accumulated 5- and 60-fold higher level of arsenate and arsenite. Genetically engineered microorganisms and bioremediation of heavy metal During the past 20 years, recombinant DNA techniques have been studied intensively to improve the degradation of hazardous wastes under laboratory conditions. Tripathi and Srivastava (2007) had have developed stable mutants Aspergillus nidulans. Resistance to 1mM Ni were developed by step-by-step repeated culturing of the fungus on the Czapeck Dox agar containing increasing concentrations of nickel chloride. Only one field test has been successfully implemented. Recombinant bacteria can be obtained by genetic engineering techniques or by natural genetic exchange between bacteria. Applications for genetically engineered microorganisms (GEM) in bioremediation have received a great deal of attention, but have largely been confined to the laboratory environment. This has been due to regulatory risk assessment concerns and to a large extent the uncertainty of their practical impact and delivery under field conditions. There are at least four principal approaches to GEM development for bioremediation application.These include: (1) modification of enzyme specificity and affinity, (2) pathway construction and regulation, (3) bioprocess development, monitoring, and control, and (4) bioaffinity bioreporter sensor applications for chemical sensing, toxicity reduction, and end point analysis. There are many reports on the degradation of environmental pollutants by different bacteria. However, only potential bioreme-diation related to GEM will be reviewed, and the construction and application of GEM will be the focus of this review. These genetically engineered microorganisms have higher degradative capacity and have been demonstrated successfully for the degradation of various pollutants under defined conditions. However, ecological and environmental concerns and regulatory constraints are major obstacles for testing GEM in the field. These problems must be overcome before GEM can provide an effective clean-up process at lower cost. The use of genetically engineered microorganisms has been applied to bioremediation. The genes responsible for microbial metal resistance mechanism are organized in operons and are usually found in plasmids carried by the resistant bacteria. The expression of the resistance genes is tightly regulated and induced by the presence of specific metals in the cellular environment. Because of the specificity of this regulation, the promoters and regulatory genes from these resistance operons can be used to construct metal-specific biosensors (promoterreportergene fusions). By using metal specific bacterial sensors in addition to chemical analyses it is possible to distinguish the bioavailable metal concentration from the total metal concentration of the samples. Recently, various metal-specific sensor strains have been developed and applied in many laboratories. These sensor strains are all based on the same

concept: a metal responsive regulation unit regulates the expression of a sensitive reporter gene. Reporter genes include those that code for bioluminescent proteins, such as bacterial luciferase (luxAB) and firefly luciferase (lucFF) or for â-galactosidase, which can be detected electrochemically or by using chemiluminescent substrates. The light produced can be measured by a variety of instruments, including luminometers, photometers and liquidscintillation counters. In the construction of biosensors, it is important to know how metal-resistance mechanisms work and how the genes coding for them are regulated. For example, the resistance for mercury is encoded by genes of the mer operon. This operon encodes proteins that are involved in the transport of mercury (Hg2+) into the cell and for the transformation of Hg 2+ to elemental mercury Hg2+, which is volatile and evaporates out of the cell (Nies 1999). Three separate biosensors have been constructed they contain different parts of the mer operon fused to the bacterial luciferase genes, using E. coli as a host strain. Two of the strains constructed did not contain the merA gene, which reduces Hg2+ to elemental mercury. Thus, these strains were not resistant to mercury and they could be used to detect only low concentrations of mercury. In contrast, the third strain contained the whole mer operon. This sensor was resistant to high concentrations of mercury but the response was not as sensitive as in the first two strains. On the other hand, the ars operon of pR773, present in certain strains of E. coli, consists of five genes: arsA, arsB, arsC, arsD and arsR, which are responsible for arsenic resistance. The arsA and arsB genes encode for the proteins that form the ion pump, which is capable of transporting As3+ out of the cell. As5+ cannot be transported by this pump, but arsC encodes for an enzyme, arsenate reductase, that reduces As5+ to As3+, which can then be removed by the pump. In addition, the arsR gene encodes for proteins that regulate the expression of the ars operon. In the arsenic biosensor developed by, the sensing of As3+ is based on controlling the expression of the firefly luciferase (lucFF) reporter gene by the regulatory unit of the ars operon of plasmid pR733 in recombinant plasmid pTOO31, with E. coli as a host strain. The regulatory unit of the ars operon consists of the ars promoter and the repressor protein, ArsR. In the absence of As3+, the expression of lucFF is repressed, while in the presence of arsenic, transcription of the promoter is induced, and luciferase is produced at a level corresponding to the concentration of arsenic. Recent advances in molecular fingerprinting methods using signature biomarkers, such as lipids and nucleic acids, provide a qualitative and quantitative measurement of microbial diversity and community composition in undisturbed and polluted soils. For example, phospholipid fatty acid analysis (PLFA) can be used to quantify microbial community structure and biomass without relying upon the cultivaltion of microorganisms. Unfortunately, this approach does not have the capability of identifying microorganisms at the species or strain level, but rather produces descriptions of microbial communities based on functional groupings of fatty acid profiles. However, the identification of particular species, contributing to the bacterial community, by cell fatty acid methyl ester profiles (FAME) can be determined from cultured isolates. Changes in the microbial community structure in response to soil metal contamination has been monitored by PLFA analysis in various studies. For example, accumulation of copper in soil as a concequence of fungicide application resulted in the development of a microbial community with markedly different PLFA patterns when

compared to uncontaminated soil. In metal contaminated soils, the increase of monounsaturated fatty acids, lower concentrations of branched-chain and methyl-branched fatty acids indicated an increase in numbers of gram-negative bacteria and a decrease of the proportion of grampositive bacteria and actinomycetes, respectively. Also, a decrease in several iso- and anteiso-branched PLFAs and an increase in cy17:0, which is considered to be typical for gram-negative bacteria, indicated a dominance of gram-negative over gram-positive bacteria in metal contaminated soils. However, found that many branched PLFAs, like br17:0 and br18:0, or iso- and anteiso-branched PLFAs, commonly found in gram-positive bacteria, increased in the metal contaminated soil. Similar results were found also in coniferous forest humus due to Ni-Cu pollution and acidification. In coniferous forest soil, also fungal markers (18:2 omega 6,9 and 20:4) decreased in response to long-term heavy metal deposition, even though various studies have found that fungi are more resistant to metals than bacteria. On the other hand, the methyl-branched PLFAs 10Me16:0, 10Me17:0 and 10Me18:0 increased in metal-polluted forest soil, indicating an increase in the proportion of actinomycetes. However, in arable soils, a decrease was observed for 10Me16:0 and 10Me18:0 in response to most metals. Also, according to one of the report relative decrease in 10Me16:0 in zinc contaminated soil when compared to undisturbed soil, suggesting that different members of the actinomycete population responded differently to the elevated metal cocentrations. A molecular approach based on 16S rDNA is useful in detecting bacterial community structure changes, because these genes are conserved and present in all bacteria. Microbial community analyses using nucleic acids, such as 16S rDNA, can detect and identify community members with high specificity to the species and strain level, and can also detect and suggest phylogenetic relationships of uncultured organisms. For example, using amplified ribosomal DNA restriction analysis (ARDRA) it was found distinct differences in microbial community structure in soil contaminated with heavy metals compared to uncontaminated soil. An alternative 16S rDNA-based method is terminal restriction fragment analysis (t-RFLP), which measures only the terminal restriction fragment of each 16S rRNA gene. Thereby, the complexity of the RFLP pattern is reduced (compared to that of ARDRA) and every visible band (fragment) is representative of a single ribotype or operational taxonomic unit. Biosorption by bacteria Biosorption is one of the most important biological mechanisms which involves the ability of microorganisms to accumulate heavy metals from contaminated site through metabolically mediated pathway (Azmat et al 2007). Bacteria may carry determinants of resistance to a number of heavy metals. Bacterial resistance to heavy metals is conferred by specific resistance determinants, which are often, but not always, carried on plasmids or transposons. Resistance is specific to one or a few metals and the mechanisms of resistance include efflux of the metal, modification of the speciation of the metal, sequestration of the metal, or a combination of these mechanisms. Cell walls of bacteria and cynobacteria are principally composed of peptidoglycans, N acetylglucasamine, â 1-4 acetyl muramic acid with peptide chains. Cell wall of gramnegative bacteria is not heavily crosslinked. They have an outer membrane which is composed of an outer layer of lipopolysaccharides (LPS), phospholipids and proteins. Gram negative bacteria are more widespread in metal contaminated soils then gram-positive bacteria. The anionic nature of

bacterial surface enables them to bind metal cations through electrostatic interactions. Because of their thickness and anionic character which is mainly due to peptidoglycan, teiochoic acid and teichuronic acids the cell wall of gram positive bacteria has high capacity for metal binding. Bacillus subtilis, B. licheniformis, Pseudomonas sp., Serratia mercascens, Pseudomonas aeruginosa, Zooglea ramigera and Streptomyces sp. are widely used for metal removal from effluent. Biosorption by fungi Among microorganisms, fungal biomass offers the advantage of having a high percentage of cell wall material which shows excellent metal binding properties. Polysaccharides, in association with lipids and proteins, represent the main constituent of fungal cell wall. In filamentous fungi outer cell wall layers mainly contain neutral polysaccharides (glucans and mannans). While the inner layers contain more of glucosamines (chitin and chitosan) in a microfibrillar structure, ligands within these matrices include carboxylate, amine phosphate, hydroxyl, sulphydral and other functional groups. Proteins are also found to be associated with metal binding. Rhizopus, Aspergillus, Streptoverticillum and Saccharomyces are important fungi used for metal biosorption. Biosorption by algae and moss Photoautotrophs marine algae have bulk availability of their biomass from water bodies. Special polysaccharides are present in the algae cell wall contained potential metal ion binding sites. The number and kind of binding sites depend on the chemical composition of the cell wall. In Pheophycean members, algin is present and contributes significantly to metal binding. It was been suggested that the polysaccharides of cell wall could provide amino and carboxyl group as well as the sulphate. The amino, carboxyl group and the nitrogen and oxygen based moieties could also form coordinated band with metal ion. Metal ion could also be electrostatically bonded to unprotonated carboxyl oxygen and sulphate covalent bonding between divalent cation and algae cell wall proteins has also been reported. Mechanisms such as entrapment of metal both in the form of insoluble micro deposits in the inter and intra-fibrillar capillaries and paracrystalline regions of polysaccharides and the binding to other biopolymers (RNA, Polyphos-phates) can contribute to the metal binding. The photoautotrophs, eukaryotic algae cell wall are mainly cellulosic and potential metal binding groups are carboxylate, amine, imidazole, phosphate, sulfhydryl, sulfate and hydroxyl. Of these amine and imidazoles are positively charged when protonated and build negatively charged metal complexes. The amino and carboxyl groups and nitrogen and oxygen of the peptide bonds are also available for coordination bonding with metal ions such as lead (II), copper (II) and chromium (IV). 5. What are PGPR? A group of microorganisms which are beneficial to crops is bacteria that colonize roots or rhizosphere soil of crop plants. These bacteria are referred to as plant growth promoting rhizobacteria (PGPR). The science of PGPR is thus relatively young in comparison to nitrogen fixing bacteria and momentarily applications to crop production are limited. The science is developing rapidly and producers and the crop production industry wise to keep abreast of developments as they may reach the dealer level in years to come. Secondary metabolites

Offensive PGPR colonization and defensive retention of rhizosphere niches are enabled by production of bacterial allelochemicals, including iron-chelating siderophores, exopolysacchrides, antibiotics, biocidal volatiles, lytic enzymes, and detoxification enzymes (Barka et al 2005). According to Thakur (2007) microbial cells can accumulate heavy metals by a variety of physicochemical and biological processes. These includes siderophores, EPS, phytoharmones, HCN, organic acids etc (Surya Devara et al 2009). Mechanisms of PGPR in plant growth promotion Microbes (bacteria and fungi) found in nature associated with rhizhosphere have a very broad ability to utilize (catabolise) virtually all naturally occurring compounds as their sources of carbon and energy, thus recycling the fixed organic carbon back into harmless biomass and carbon dioxide. This capability of microbes has evolved over 3 billion years of the planets history and is responsible for the balance between photosynthesis (by plants and algae), fixing carbon dioxide into biomass, and respiration (by animals and bacteria), and converting organic compounds back to carbon dioxide by oxidation through natural detoxification processes. Recent advances in genetics and molecular biology in the last 20-30 years has shown that bacteria are genetically extremely adaptable, and in addition to the advantage due to their rapid growth rates, have a range of mechanisms which enable them to adapt to new environments. When compounds are persistent in the environment, their biodegradation often proceeds through multiple steps utilizing different enzyme systems or different microbial populations. The mechanisms by which PGPR increase crop performance is not well understood. There are several PGPR inoculants currently commercialized that seem to promote growth through at least one mechanism; suppression of plant disease (termed Bioprotectants), improved nutrient acquisition (termed Biofertilizers), or phytohormone production (termed Biostimulants). Inoculant development has been most successful to deliver biological control agents of plant disease, that is organisms capable of killing other organisms pathogenic or disease causing to crops. Bacteria in the genera Bacillus, Pseudomonas, Arthrobacter, Azospirillum, Enterobacte and, Serratia have been found to have enormous potential as plant growth promoting rhizobacteria (PGPR) and are now used in agriculture as bioinoculants. Earlier workers have reported increase in crop yield as high as 160% using PGPR are the biological control agents predominantly studied and increasingly marketed (Saxena & Matta 2005). They suppress plant disease through at least one mechanism; induction of systemic resistance, and production of siderophores or antibiotics. Paenibacillus polymyxa (formely Bacillus polmyxa), a non pathogenic and endospore-forming Bacillus, is one of the most industrially significant facultative anaerobic bacterium. It occurs naturally in soil, rhizosphere and roots of crop plants and in marine sediments. During the last two decades, there has been a growing interest for their ecological and biotechnological importance, despite their limited genomic information. Paenibacillus polymyxa have wide range of properties as that of PGPR. It also helps in bioflocculation and in the enhancement of soil porosity. In addition, it is known to produce optically active 2,3butanediol (BDL), a potentially valuable chemical compound from a variety of carbohydrates (Lal & Tabacchioni 2009). Exposure to the PGPR triggers a defense response by the crop as if attacked by pathogenic organisms. The crop is thus armed and prepared to mount a successful defense against eventual challenge by a pathogenic organism. Siderphores produced by some PGPR scavenge heavy metal micronutrients in the rhizsophere (e.g. iron) starving pathogenic

organisms of proper nutrition to mount an attack of the crop. Interestingly, plants seem capable to still acquire adequate micro-nutrient nutrition in the presence of these PGPR. Antibiotic producing PGPR release compounds that prevent the growth of pathogens. The compounds produced are not unlike antibiotics we take to rid of pathogens of human. Biofertilizers are also available for increasing crop nutrient uptake of nitrogen from nitrogen fixing bacteria associated with roots (Azospirillium), iron uptake from siderophore producing bacteria (Pseudomonas), sulfur uptake from sulfur-oxidizing bacteria (Thiobacillus), and phosphorus uptake from phosphate-mineral solubilizing bacteria (Bacillus, Pseudomonas). Nitrogen fixing biofertilizers provide only a modest increase in crop nitrogen uptake (at best an increase of 20 lbs N per acre). The popular inoculants presently commercialized for increasing phosphorus uptake through phosphorus solubilization (Penicillium and Aspergillus) and phosphorus transfer directly to roots (mycorrhizae) are not bacteria but fungi. Species of Pseudomonas and Bacillus can produce as yet not well characterized phytohormones or growth regulators that cause crops to have greater amounts of fine roots which have the effect of increasing the absorptive surface of plant roots for uptake of water and nutrients. These PGPR are referred to as biostimulants and the phytohormones they produce include indole-acetic acid, cytokinins, gibberellins and inhibitors of ethylene production. Agrochemicals like fertilizers and pesticides have greatly influenced natural PGPR microbes in agro systems. The economic and ecological problems of today have re-invigorated the idea of using biofertilizers and biocontrol agents in order to reduce the application of costly and environmentally-polluting agrochemicals to a minimum. Plant beneficial microbial resources promise to replace and/or supplement many such destructive, high intensity practices and support ecofriendly crop production. Plant growth promoting rhizobacteria (PGPR) may promote growth directly, e.g. by fixation of atmospheric nitrogen, solubilization of minerals such as phosphorous, production of siderophores that solubilize and sequester iron, or production of plant growth regulators (hormones), HCN, ACC deaminase production, EPS production, lytic enzymes, competition and by inducing systemic resistance (Sharma et al 2010) . Some bacteria support plant growth indirectly, by improving and/or eliminating the growthrestricting conditions either via production of antagonistic substances or by inducing resistance against plant pathogens. Apart from rhizobial symbionts, the rhizosphere-associated beneficial bacteria (RABB) consist of either biocontrol agents such as Pseudomonas and Bacillus group, which antagonize pathogenic or deleterious microorganisms and/or bioenhancer agents such as Azospirillum, Herbaspirillum, Enterobacter, Acetobacter, Azotobater, and Pseudomonas group which directly enhance plant growth. The extensive and enduring challenges in soil microbiology depend on the development of efficient methods to be acquainted with the types of microbes present in soil, and to determine the functional performance of the overall microbial groups in situ. It is an interesting topic to investigate the combined uses of species richness and diversity as well as to estimate the combinatorial effect of species richness and diversity in order to understand their role and distribution in their habitat (Jha et al 2010). Phosphate solubilizers The ability of a few soil microorganisms to convert insoluble forms of phosphorus to an accessible form is an important trait in plant growth-promoting bacteria for increasing plant

yields (Chen et al 2006). An adequate supply of phosphorus during early stage of phosphorus during early phases of plant development is important for laying down the primordia of plant reproductive parts (Sayyed et al 2001). Phosphate solubilizing bacteria application has promoted P-uptake as well as the yields in several crops. The plant growth promoting effects of phosphate solubilizers is considered to be related to their ability to synthesize plant growth regulating substances (Vikram et al 2007). Phosphorus deficiency in soil can severely limit plant growth productivity, particularly in legumes, where both the plants and their symbiotic bacteria are affected, and this may have a deleterious effect on nodule formation, development and function (Sridevi & Mallaiah 2007). The concentration of soluble in soil is very low which leads to deficiency of soluble phosphate and make it a limiting factor in plant nutrient. Lower the quantiy of phosphate in the medium, greater the solubilization (Geethalaksmi & Panneerselvam et al 2009). The use of phosphate solubilizing bacteria as inoculants increases the P uptake by plants. Phosphorus is second only to nitrogen in mineral nutrients most commonly limiting the growth of crops. Phosphorus is an essential element for plant development and growth making up about 0.2% of plant dry weight. Plants acquire P from soil solution as phosphate anions. However, phosphate anions are extremely reactive and may be immobilized through precipitation with cations such as Ca2+, Mg2+, Fe3+ and Al3+, depending on the particular properties of a soil. In these forms, P is highly insoluble and unavailable to plants. As the results, the amount available to plants is usually a small proportion of this total. Several scientists have reported the ability of different bacterial species to solubilize insoluble inorganic phosphate compounds, such as tricalcium phosphate, dicalcium phosphate, hydroxyapatite, and rock phosphate. Most of the Indian rock phosphates are not suitable for phosphatic fertilizer production due to low reactivity and impurities present in them and could be utilized as direct application fertilizers with or without modifications. Sustainable p supply in agro ecosystem can be achieved by using rock phosphate in conjugation with PSM. Microbial volatilization of rock phosphate especially low grade and its use in agriculture is receiving greater attention. Owing to escalating cost of P fertilizers, there is a need to switch over to cheaper source of rock phosphate which indigenously available by using P solubilizers, organic and chemical amendments (Narsian et al 2010). The concentration of soluble P in soil is usually 1 ppm or less. Therefore, the application of phosphatic fertilizers is essential for enhanced crop yield. However, more than two-thirds of the applied phosphatic fertilizers every year. Since crops require 10 to 100 kg P ha-1, the ability of microorganisms to solubilize and mineralize P in soil is vital. Pseudomonas, Micrococcus, Bacillus, Aerobacter, Xanthomonas, Brevibacterium, Alcaligenes, Rhizobium, Rhodotorula, Penicillium, Fusarium, Sclerotium, Aspergillus, Pythium, Phoma and Cladosporium have been reported to be active in phosphate solubilization (Nasreen et al 2005). Mechanisms of phosphate solubilization The principal mechanism for mineral phosphate solubilization is the production of organic acids, and acid phosphatases play a major role in the mineralization of organic phosphorus in soil. It is generally accepted that the major mechanism of mineral phosphate solubilization is the action of organic acids synthesized by soil microorganisms. Production of organic acids results in acidification of the microbial cell and its surroundings. The production of organic acids by

phosphate solubilizing bacteria has been well documented. Gluconic acid seems to be the most frequent agent of mineral phosphate solubilization. Also, 2-ketogluconic acid is another organic acid identified in strains with phosphate solubilizing ability. Strains of Bacillus were found to produce mixtures of lactic, isovaleric, isobutyric and acetic acids. The capacity of bacterial isolates to to solubilize phosphate depends upon the zone/site of their origin. Those derived from rhizoplanes have the highest capacity; rhizosphere organisms have intermediate capacity while those from bulk soils have the least phosphate solubilizing activity. The same is true for fungi. The „climosquence‟ as well as „chemosequence‟ of soils determine the survival of phosphate solubilizers (Narsian et al 2009). Plant growth regulators (hormones) Plant-growth promotion by PGPR include bacterial synthesis of the plant hormones indole-3acetic acid, cytokinin, and gibberellin; breakdown of plant produced ethylene by bacterial production of 1-aminocyclopropane-1- carboxylate deaminase; and increased mineral and N availability in the soil. Although low-molecular-weight plant volatiles such as terpenes, jasmonates, and green leaf components have been identified as potential signal molecules for plants and organisms of other trophic levels, the role volatile emissions from bacteria play in plant development is unknown. Ethylene is a gaseous plant growth hormone produced endogenously by almost all plants. It is also produced in soil through a variety of biotic and abiotic mechanisms, and plays a key role in inducing multifarious physiological changes in plants at molecular level. Apart from being a plant growth regulator, ethylene has also been established as a stress hormone. Under stress conditions like those generated by salinity, drought, waterlogging, heavy metals and pathogenicity, the endogenous production of ethylene is accelerated substantially which adversely affects the root growth and consequently the growth of the plant as a whole. Certain plant growth promoting rhizobacteria (PGPR) contain a vital enzyme, 1-aminocyclopropane-1-carboxylate (ACC) deaminase, which regulates ethylene production by metabolizing ACC (an immediate precursor of ethylene biosynthesis in higher plants) into ketobutyrate and ammonia. Inoculation with PGPR containing ACC deaminase activity could be helpful in sustaining plant growth and development under stress conditions by reducing stress-induced ethylene production. Lately, efforts have been made to introduce ACC deaminase genes into plants to regulate ethylene level in the plants for optimum growth, particularly under stressed conditions. In this review, the primary focus is on giving account of all aspects of PGPR containing ACC deaminase regarding alleviation of impact of both biotic and abiotic stresses onto plants and of recent trends in terms of introduction of ACC deaminase genes into plant and microbial species (Saleem et al 2009). The cyanobacterial strains belonging to the genera Nostoc and Anabaena comprised 80% of the rhizhosphere isolates, which were also efficient in enhancing the germination and growth of wheat seeds and exhibited significantly high protein accumulation and IAA production (Prasanna et al 2009). HCN production Weeds cause more economic losses in agricultural lands. Currently, the most effective means of managing weeds are herbicides. Some widely used herbicides have been implicated in contamination of groundwater, soils, and food products, which may threaten public health and

safety. Today, biological methods have known as effective and appropriate ways in weed control. A major group of PGPR with potential for biological control is the Pseudomonads. A secondary metabolite produced commonly by rhizosphere Pseudomonads is Hydrogen Cyanide (HCN), a gas known to negatively affect root metabolism and root growth. In studying the volatiles produced by microorganisms isolated from spoiled chicken it was noted that strains of Pseudomonas fluorescens produced hydrogen cyanide when grown on either sterile chicken or on Trypticase soy agar supplemented with 0.5% yeast extract (Freeman et al 1975). Volatiles produced by fluorescent Pseudomonas Psfl-1 inhibited sclerotia formation of pathogen up to 58.0%, principle compound was HCN (Bandopadhyay 2009). Cyanogenesis was predominantly associated with pseudomonads and was enhanced when glycine was provided in the culture medium. Production of extacellular polymorphic substances (EPS) Microorganisms with a capability of secreting array of secondary metabolites like (EPS), mucilage, capsule, etc. have the potential of adsorbing variety of heavy metals. There are different types of exopolysaccharides which form a biofilm around the cells facilitating attachment of the cells to surface, colonization and providing protection against unfavourable conditions. Xanthan, alginate, pullulan, dextran, alternan, levan and inulan of the example (Purama & Goyal 2005). EPS contains ionizable functional groups such as carboxyl, phosphoric, amine and hydroxyl groups which enables EPS to sequester heavy metals. Ion exchange, complexation with negatively charged functional groups, adsorption and precipitation are the mechanisms involved in metal biosorption onto EPS (martin et al 2008). This ability of EPS to bind and make possible to reduce heavy metals may represent a novel feature in some microorganisms (Pal and Paul 2008). Sayyed et al (2008) have studied the effect of Ca 2+ and iron on EPS synthesis in Alcaligens fecalis. Production of siderophores Siderophores are small molecular weight extracellular organic compounds secreted by microorganisms under iron-starved conditions, used by them to chelate and solubilize iron. Siderophore was found to complex with heavy metals like Cadmium, Lead, Nickel, Arsenic (III, V), Aluminium, Magnesium Zinc, Copper, Cobalt, and Strontium other than iron (Nair et al 2006). Siderophore producing PGPR function as a biocontrol agent, by depriving the pathogen from iron nutrition thus resulting in increased yields of crop (Sayyed et al 2010). Siderophore based Biological Control Agents (BCAs) are gaining commercial significance as they are safer, do not lead to biomagnification, their self-replication circumvents repeated application and target organisms do not develop pesticide resistance (Sayyed et al 2005). Fluorescent pseudomonads have been recognized as biocontrol agents against certain soil-borne plant pathogens. They are characterized by the production of yellow-green pigments termed pyoverdines, which fluoresce under UV light and function as „siderophores‟. Pyoverdines chelate iron in the rhizosphere and deprive pathogens of iron which is required for their growth and pathogenesis. Involvement of iron competition in biological control has been inferred from experiments in which mutants deficient in siderophore production were compared with the wild type with respect to biocontrol activity. A number of fluorescent Pseudomonads with biocontrol activity against Ralstonia solanaceraum causing bacterial wilt disease in tomato, have been reported. Plant-associated Pseudomonas lives as saprophytes and parasites on plant

surfaces and inside plant tissues. Many plant-associated Pseudomonas promote plant growth by suppressing pathogenic micro-organisms, synthesizing growth-stimulating plant hormones and promoting increased plant disease resistance. Others inhibit plant growth and cause disease symptoms ranging from rot and necrosis through to developmental dystrophies such as galls. It is not easy to draw a clear distinction between pathogenic and plant growth-promoting Pseudomonas. They colonize the same ecological niches and possess similar mechanisms for plant colonization. Pathogenic, saprophytic and plant growth-promoting strains are often found within the same species, and the incidence and severity of Pseudomonas diseases are affected by environmental factors and host-specific interactions. Plants are faced with the challenge of how to recognize and exclude pathogens that pose a genuine threat, while tolerating more benign organisms. A better understanding of the molecular basis of plant- plant growthpromoting Pseudomonas (PGPP) interactions and of the key differences between pathogens and PGPP will enable researchers to make more informed decisions in designing integrated diseasecontrol strategies and in selecting, modifying and using PGPP for plant growth promotion, bioremediation and biocontrol (Jagadeesh et al 2001). Growth and siderophore production by PGPR is attributed to organic acids, sugars, amino acids, minerals, enzymes and several other components of root exudates. Any factor influencing either the growth or siderophore production by PGPR would greatly influence the efficacy of that PGPR in the plant growth promotion and disease suppression (Sayyed et al 2009). 6. Conclusion The advent of modern chemical industry has resulted in the release of huge amounts of novel organic compounds, as industrial by-products, pesticides, other agrochemicals etc into the environment. These new compounds tend to persist in the environment and have possibility to bioaccumulate in food chains. The presence of „artificial‟ groups such as chloro-, nitro- or sulfonate- in many synthetic chemicals makes them resistant to decomposition, as they are no longer recognized by the degrading microbes. The compounds are highly resistant to biodegradation is known as recalcitrant compounds. Xenobiotic compounds are often toxic to life and are also often hard for microorganisms to metabolize (because they contain molecular arrangements that not normally encountered in nature). Thus, many of these compounds can accumulate in the environment and continue to be a hazard for many years. Xenobiotic substances are becoming an increasingly large problem in sewage treatment systems, since they are relatively new substances and are very difficult to remove from the environment. Human activities have brought about widespread pollution of the natural environment not only in the form of heavy metals but also in the form of inorganic and organic compounds. A number of organic pollutants, such as polycyclic aromatic hydrocarbons, polychlorinated aromatic compounds and nitrogen containing aromatic compounds are resistant to degradation and represent an ongoing toxicological threat to both wildlife and human beings. Over recent years, a growing number of potential hazards linked to the ubiquitous presence of POPs in the environment have been reported. Bioremediation is an attractive alternative to traditional physicochemical techniques for the remediation of these POPs at a contaminated site, as it can be more cost-effective and it can selectively degrade the pollutants without damaging the site or its indigenous flora and fauna. However, despite being hailed as a panacea to the safe and effective solution to contaminated environmental media, bioremediation technologies, to date,

have had limited applications due to the challenges of substrate and environmental variability, as well as the limited biodegradative potential and viability of naturally occurring microorganisms. Microorganism with suitable and stable genetic traits, and efficient and effective biodegradation processes would be helpful for clean and green environment. 7. Future strategy Microbes play a key role in controlling the speciation and cycling of metals in soil. Because bioavaila-bility, toxicity and reactivity of metals is greatly influenced by chemical speciation, it is important to have a better understanding of the major factors that link microbial activity to the biogeochemistry of metals. In addition, plants can considerably alter the mobility and bioavailability of metals in soil. Therefore, understanding the roles of microorganisms and plants in cycling of metals may lead to improved processes for bioremediation of contaminated sites. Studies indicated near isogenic PGPR strains consortium with biofertilizer and growth regulator effective beneficial traits for pathogen suppression, plant nutrient supply and growth promotion. Molecular characterization of functional genes and biopesticide from genetically improved PGPR may necessitate sustainable agriculture. There is no doubt that bio-fertilizers are the potential tools for sustainable agriculture not only in India but also around the world. The use of bio-fertilizer in preferences to chemical fertilizer is always welcome taking into consideration the suitability of agriculture. It is beneficial always in terms of soil fertility, ecological health etc. As, the use of bio-fertilizer, till so far, is grossly inadequate in India, more emphasis on its production, consumption and also proper distribution need to be taken into consideration. The Government of Indian has made substantial investments in biotechnology research. Still substantial investments require to achieve the target. The problem related to adverse climatic situation, soil condition, production technologies, storage, awareness among the farmers are also some of the important areas to be concentrated towards this direction. Bacterial consortium influences plant growth positively by a multitude of synergistic mechanisms under adverse environmental conditionds when compared to single strain inoculation. . This finding has significant implications in mitigation of adverse effects of climate change on crop growth and may lead to development of microorganism-based products for alleviation of such effects. However, further studies are required under green house and field conditions and the actual mechanism of protection has to be elucidated. Acknowledgment Authos are thankfull to BRNS, DAE, goverment of India for providing financial assisatnce under the scheme of YSRA. References Azmat R, Uzma & Uddin F (2007). Biosorption of toxic metals from solid sewage by marine green algae,. Asian J of plant sciences,. 6 (1) 42-45.

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