The most important direct effects involved in plant growth promoting include biological nitrogen fixation, phytohormone production, nutrient solubilization, siderophore production, and ACC deaminase activity.

Atmospheric N is reduced to plant available form though natural or artificial means. When done artificially, N₂ is reduced to ammonia via the Haber–Bosch process (Rubio and Ludden, 2008), in which natural gas (CH₄) and N₂ are converted to reduced forms of N at high temperature and pressure. In nature, N₂ reduction is performed by N-fixing microorganisms that use the nitrogenase enzyme to reduce N₂ to ammonia (Kim and Rees, 1994). This biological nitrogen fixation (BNF) is responsible for two-thirds of the total fixed N worldwide.

The microbes performing BNF can be generally categorized as symbiotic, associative symbiotic, and free-living. However, a number of free-living N fixing bacteria, such as Azotobacter, Gluconoacetobacter, and Azospirillum spp. are present in nature and fix N for plants (Bashan and Levanony, 1990). The highest proportion of BNF is performed by symbiotic N₂ fixers, i.e., rhizobia, which make symbiotic associations with the roots of leguminous plants (Zahran, 2001). The establishment of symbiotic association involves a complex mechanism and exchange of chemical signals between host plant and symbionts i.e., rhizobia (Giordano and Hirsch, 2004), leading to the formation of root knots, also called nodules. These develop from the swelling of cortical cells that host rhizobia as intracellular symbionts.

PGPR other than rhizobia also have the nitrogenase enzyme and can fix N in non-leguminous plants, such as diazotrophs, which are capable of forming non-obligate interactions with host plants (Glick et al., 1998) other than legumes. Nitrogenase is a two-component metallo-enzyme (Dean and Jacobson, 1992) that consists of an iron (Fe) protein (dinitrogenase reductase) and molybdenum (Mo)-Fe protein (dinitrogenase). For nitrogenase complex to function, both components should be present. During N fixation, the Fe protein receives electrons with high reducing power from a low redox donor, such as reduced ferredoxin (Fd) or flavodoxin, and is reduced itself. The reduced Fe protein passes its electrons to Mo-Fe protein and becomes oxidized. This complex chain of reactions requires significant metabolic energy to reduce N₂. The genes involved are nif genes, which are present both in symbiotic and free-living microorganisms (Kim and Rees, 1994). Inoculation of legumes with rhizobia of a specific cross inoculation group can be helpful to improving the nodulating capability of crops under field conditions (Ahmad et al., 2013b).

Plants produce plant growth regulators/phytohormones, complex organic compounds that control plant growth and productivity. Due to their complexity, plants need a considerable amount of energy and nutrients to synthesize them. Bacteria synthesize significant quantities of phytohormones, and release them into the plant, resulting in pronounced positive effects on plant growth and development. It is reported that bacteria can produce up to 60 times more plant growth regulators than plants themselves (Camerini et al., 2008).

Phytohormones, such as indole acetic acid (IAA), ethylene, abscisic acid, cytokinins, and gibberellins production by PGPR help to improve crop growth and performance. Phytohormones are involved in plant growth at different scales, such as cell division, cell enlargement, seed germination, root formation, and stem elongation (Taiz and Zeiger, 2000; Khalid et al., 2006). Microbially-produced phytohormones have direct influence on plants' internal physiological processes and are involved in plant growth (Kang et al., 2010). The effectiveness of these microbially produced phytohormones to improve crop productivity has been well-documented (Zahir et al., 2007, 2010; Jamil et al., 2018). Microbes can meet the plant's hormonal requirements, saving the plant's metabolic energy for growth and reproduction. Microbially produced phytohormones play an effective role both under normal and stress conditions.

Auxins can alleviate the adverse effects of stress on plant growth. Some plants produce enough auxins to cope with adverse conditions, while others produce insufficient amounts, resulting in an inability to alleviate stress conditions. To meet the plant's auxin requirements, exogenous application of auxins or inoculation with microbes capable of producing auxins can be helpful and allow for resumption of normal metabolic functions. Ahmad et al. (2013c) evaluated the potential of auxin-producing Pseudomonas and Rhizobium strains to improve osmotic stress tolerance in Vigna radiata reporting an increase in total dry matter and salt tolerance index. In another study, Jamil et al. (2018) exogenously applied L-tryptophan in combination with Pseudomonas fluorescens under drought conditions, and reported a significant increase in growth, physiology, and yield. Abscisic acid (ABA) also improves plant development under stress conditions (Zhang et al., 2006) and plays an important role in photoperiodic induction of flowering (Wilmowicz et al., 2008). Patten and Glick (2002) inoculated canola plants with IAA-producing bacterial strains and reported increase in root length in comparison to IAA-deficient mutant and control plants. Similarly, the production of auxins, cytokinin and gibberellins by many strains of Bacillus, Paenibacillus, Pseudomonas, and Azospirillum has been reported (Ahmad et al., 2011; Gamalero and Glick, 2011; Mumtaz et al., 2017). Moreover, phytohormone-producing strains improved the growth and productivity of onion (Allium cepa) (Ahmad et al., 2016), cucumber (Cucumis sativus) (Ahamd et al., 2015) and maize (Zea mays) (Mumtaz et al., 2018). Steenhoudt and Vanderleyden (2000) showed that the production of phytohormones is the main mechanism used by Azospirillum strains to improve plant growth. Commercially available phytohormones can be used for improving plant growth (Jamil et al., 2018), although microbially-produced phytohormones are often more effective and economical under field conditions (Khalid et al., 2006).

Soil microorganisms are solely responsible for nutrient cycling. Around 50% of soil organic matter is composed of carbon, while the rest consists of N, P, S, and other nutrients. In addition to the decomposition of soil organic matter, microbes also make chemically fixed nutrients, such as phosphorus (P), zinc (Zn), potassium (K), and iron (Fe) available. The main mechanism in the solubilization of P, K, Fe, and Zn is the lowering of pH from the production of organic acids (Jennings, 1994). The P solubilizing soil bacteria include free living rhizobacteria, such as Pseudomonas, the symbiotic nitrogen fixers (rhizobia), and asymbiotic nitrogen fixers (Azotobacter). In addition to release bound P through phosphatase production and rhizosphere acidification, these bacteria also provide phytohormones to crop plants and protect plants from various diseases through synthesizing siderophores, antibiotics, cyanogenic compounds, etc. (Khan et al., 2013). To solubilize Zn and phosphate, bacteria produce gluconic acid and its derivatives (Gadd and Sayer, 2000; Saravanan et al., 2007). Derivatives of gluconic acid e.g., 5-keto-gluconic acid and 2-keto-gluconic acid have also been reported in solubilization of Zn by Saravanan et al. (2007). Soil bacteria and fungi have ability to reinstate soil fertility of degraded lands by improving nutrient bioavailability through nitrogen fixation and solubilization of P, K, and Fe, and aggregate satbility (Rashid et al., 2016).

Bacillus spp. has the potential to solubilize insoluble Zn resources. Ramesh et al. (2014) demonstrated the Zn solubilization ability of Bacillus aryabhattai on Tris-minimal medium separately amended with different Zn compounds. Similarly, Hussain et al. (2015) reported the formation of a halo zone on ZnO amended Bunt and Revira medium by Bacillus sp. (AZ6). Mumtaz et al. (2017) also demonstated Zn solubilization by screening 70 isolates and identified four best ZSB strains, Bacillus sp. ZM20, Bacillus aryabhattai ZM31, Bacillus subtilis ZM63, and Bacillus aryabhattai S10. Gandhi and Muralidharan (2016) reported Zn solubilization in ZnO amended broth inoculated with Acinetobacter sp. (AGM3). Shaikh and Saraf (2017) reported a 6-fold increase in Zn and Fe content in grains of wheat inoculated with Exiguobacterium aurantiacum strain in comparison to the control. Solubilization of Zn in ZnO amended broth inoculated with Bacillus aryabhattai and Gluconaacetobacter diazotrophicus (Ramesh et al., 2014; Mumtaz et al., 2017) was also reported. Bacillus strains solubilize unavailable Zn by producing and secreting chelating ligands, organic acids, amino acids, vitamins, phytohormones, and through oxidoreductive systems and proton extrusion (Wakatsuki, 1995; Saravanan et al., 2003).

However, the processes involved in P solubilization depend on the organic and inorganic nature of phosphate complexes. Various biochemical processes, comprising of organic acids secretion and proton discharge, achieve phosphate solubilization. The extent and forms of fixed P in soil depend on the soil's pH. At lower pH, Fe, and aluminum (Al) phosphate complexes are abundant, while at higher pH, calcium phosphate minerals are formed (Goldstein, 2000). Solubilization of Fe/Al-phosphate complexes by PSB mainly takes place through proton release, altering the negative charge at exchange sites, to facilitate the release of phosphate ions. The decrease in adsorption of phosphates enhances availability of primary and secondary orthophosphates (Henri et al., 2008). Furthermore, PSB secrete carboxylic acid that releases carboxyl ions, which replace P in precipitated complexes through ligand exchange. In alkaline conditions, P is precipitated with calcium compounds (Goldstein, 2000). The PSB solubilize calcium phosphate complexes by releasing organic acids, acidifying the surrounding environment. The decrease in pH releases utilizable P and increases the availability of other macro and micronutrients. The calcareousness of soils increases buffering capacity, but reduces the efficiency of PSB in releasing P (Stephen and Jisha, 2009). To solubilize the organic proportion of immobilized P, PSB use alkaline and acidic phosphatases, which liberate bioavailable inorganic forms of P. Phosphatases are also exudated by plant roots, but the largest quantity is secreted by PSB (Dodor and Tabatabai, 2003).

Fe is an essential mineral nutrient in plants and microorganisms, except some species of lactobacilli (Neilands, 1995). Under aerobic conditions, Fe exists in its oxidized form, i.e., ferric Fe, which forms insoluble complexes, such as hydroxides and oxyhydroxides, which are unavailable for uptake by plants and microorganisms (Rajkumar et al., 2010). PGPR secrete low molecular mass iron chelating compounds, siderophores, which solubilize iron and increase its availability for microbes and plants (Machuca et al., 2007).

Microbially-released siderophores increase plant Fe uptake through different mechanisms, such as chelation and release of Fe, direct uptake of Fe complexed siderophores, or by ligand exchange reactions (Schmidt, 1999). As a result, the soluble metal concentration increases through binding with siderophores. Inoculation with siderophore-producing rhizobacteria improved plant growth and nutrient assimilation (Rajkumar et al., 2010). Under low Fe conditions, siderophores can solubilize Fe from minerals and organic compounds (Khan et al., 2009a). Siderophores also form stable complexes with heavy metals, alleviating the toxic effect of heavy metals (Rajkumar et al., 2010). Marathe et al. (2015) inoculated Glycine max seeds with siderophore-producing Pseudomonas spp., and reported enhanced growth. Inoculation also showed antifungal activity against Aspergilus spp. In another study, Sharma et al. (2003) inoculated mung bean with Pseudomonas (GRP3) strain having ability to produce siderophores, and reported enhanced Fe contents in inoculated plants. Rajkumar et al. (2010) reported the uptake of Fe in cereal grains through production of siderophores by rhizobacteria. Pseudomonas fluorescens synthesizes Fe-pyoverdine, which can increase the Fe uptake in Arabidopsis thaliana tissues, enhancing plant growth (Vansuyt et al., 2007). Microbes with the ability to produce siderophores could be helpful in chelation of Fe from mineral and organic compounds to make it bioavailable. They also can promote uptake of Fe and other minerals in nutrient-deficient soils.

Ethylene is an endogenously produced phytohormone with a specific role in determining plant maturity. Lower levels of ethylene are essential for plant metabolism during normal growth and development (Khalid et al., 2006). It is a stress hormone (Saleem et al., 2007) that helps plants to cope with biotic and abiotic stress. Ethylene negatively affects normal metabolic processes in plants leading to decrease in root and shoot growth. For example, Ahmad et al. (2011) reported a decrease in root and shoot length and increased stem diameter due to salinity stress and linked it to increased concentrations of ethylene. It has been well-documented that PGPR strains belonging to genera Bacillus, Enterobacter, and Pseudomonas isolated from stress conditions contain ACC deaminase enzyme (Nadeem et al., 2009; Ahmad et al., 2011) and improve plant growth under biotic and abiotic stresses (Mayak et al., 2004; Zahir et al., 2010; Ahmad et al., 2012, 2013b; Glick, 2012). ACC is the immediate precursor of ethylene and cleaves it into α-ketobutyrate and NH₃ (Glick et al., 1998). Therefore, these bacterial strains can increase stress tolerance in plants by decreasing ethylene levels, allowing increased plant growth even under stress (Zahir et al., 2008; Ahmad et al., 2013b). Consequently, the use of these bacteria as biofertilizers and biopesticides can be helpful in reducing the dependence on chemical fertilizers and pesticides.

Examples of indirect effects of using soil microorganisms for plant growth promotion are biological control of phytopathogens through competition for nutrients, production of antibiotics, hydrolytic enzymes, and siderophores, along with the triggering of ISR in plants (Lugtenberg and Kamilova, 2009; Glick, 2012). PGPR can be used as a tool for biocontrol of plant pathogens as they indirectly improve plant growth by suppressing pathogenic microorganisms. The antibiotics produced are mostly effective against fungal pathogens (Glick, 2012). PGPR produce a number of antibiotics and antifungal metabolites, such as viscosinamide, tensin, pyrrolnitrin, phenazines, 2,4-diacetyphlaroglucinol, pyoluteorin and hydrogen cyanide (Raaijmakers et al., 2002; Haas and Keel, 2003; Compant et al., 2005; Mazurier et al., 2009; Bhattacharyya and Jha, 2012; Glick, 2012). Hydrogen cyanide works synergistically with antibiotics to improve their efficiency (Glick, 2012).

The production of hydrolytic enzymes, such as lipases, proteases, glucanases and chitinases by PGPR is also an effective mechanism of biocontrol. Hydrolytic enzymes dissolve the cell wall of fungal pathogens, suppressing their growth. Studies have shown PGPR strains to be successful in controlling pathogenic fungi from genera including Fusarium, Sclerotium, Botrytis, Rhizoctonia, Phytophthora, and Pythium (Frankowski et al., 2001; Kim et al., 2008; Glick, 2012).

In addition to improvement in Fe availability in crop plants, siderophores produced by PGPR strains can limit phytopathogen proliferation by suppressing Fe availability to pathogenic fungi (Kloepper et al., 1980). High affinity siderophores that bind Fe, limit its availability to fungal pathogens, therefore suppressing their growth (Glick, 2012). Siderophore producing bacteria can be used as a biocontrol agent, as plants need much lower Fe concentrations than most microorganisms (O'Sullivan and O'Gara, 1992) Moreover, many plants can take up siderophore-complexed Fe (Wang et al., 1993) while not the pathogenic microorganisms.

PGPR have the ability to effectively colonize plant roots to better use root exudates. Inoculation with PGPR can lead to higher proliferation and ultimately competition with pathogenic microorganism for nutrients and space, suppressing pathogenic microorganism growth. For example, Innerebner et al. (2011) tested bacterial strains from the genera Methylobacterium and Sphingomonas for their biocontrol activity against Pseudomonas syringae pv. tomato DC3000 in Arabidopsis thaliana. They reported that strains from the genus Sphingomonas diminished disease symptoms, suppressed pathogenic growth, and protected A. thaliana plants from developing severe disease symptoms. However, Methylobacterium strains were ineffective in controlling pathogens.

PGPR interact with plant roots and induce resistance in plants against pathogenic microorganisms, leading to ISR. ISR is phenotypically similar to the systemic acquired resistance (SAR), the plant's internal mechanism for responding to infection by pathogens (Pieterse et al., 2009). In plants, ISR involves jasmonate and ethylene signaling, which stimulate the host plant's defense responses against a range of pathogens (Verhagen et al., 2004). In addition to phytohormones, individual bacterial components can induce ISR, such as homoserine, 2,4-diacetylphloroglucinol, cyclic lipopeptides, lipopolysaccharides (LPS), lactones, and some other volatile compounds (Lugtenberg and Kamilova, 2009).

Persistant organic pollutants (POPs) include dyes used in textile and other inductries, pesticides and polycyclic aromatic hydrocarbons. These are toxic chemicals that have adverse effects on human health and environment around the world. Most of the POPs are generated in one country and affect human and wildlife far from where they are generated, used or released. Chemically these are complex compounds which are classified as xenobiotics thus barely removed from the environment. Microbial strains have been screened and identified that can degrade these POPs through their enzyme systems.

Bacterial strains have been identified that efficiently degrade the xenobiotic/recalcitrant compounds in industrial wastewater (Khalid et al., 2008a,b). Fungal species are also reported to degrade/detoxify industrial effluents. Bacteria contain specific enzymatic/gene systems responsible for the degradation of toxic compounds found in industrial effluents. Bacteria use biosorption and enzymatic degradation, or a combination of both, to detoxify/decolorize industrial effluents containing azodyes (Wu et al., 2012). An azo-reductase enzyme has been isolated and characterized from bacterial species. Azo-dyes have strong bonding properties, but some bacteria have the ability to break them with the azo-reductase enzyme (Chen, 2006). Oxygenase and hydroxylase enzymes also degrade the intermediate products formed during decolorization (Khalid et al., 2009). In this process, different amounts and kinds of microbes that degrade recalcitrant/xenobiotic compounds are involved (Table 2).

Several studies demonstrated the ability of fungi, algae, and yeast to degrade industrial effluents in wastewater (Dresback et al., 2001; Olguin, 2003). However, carbon sources are required in the dye solution for microbial growth during the decolorization process. Factors, such as pH, temperature and presence of salts in contaminants can also affect the rate of the biodegradation of industrial effluents (Prasad and Rao, 2011). Many species of bacteria, fungi, algae, and yeast are reported to decolorize the dyes (Khalid et al., 2009).

Pandey et al. (2018) tested the potential of the lignolytic mushroom Lenzites elegans WDP2 to dicolorize synthetic dyes. Brilliant green, malachite green, and Congo red were decolorized by almost 93, 21, and 99%, respectively. The Congo red dye was completely bio-absorbed by fungal culture within 1 month. The fungal decolorized broth revealed an extracellular laccase activity in all the three cases, supporting the involvement of laccase enzyme in decolorization. Photomicrographs clearly showed the bio-sorption of the dyes by fungal culture into the mycelium/spores.

Paper and pulp industries also produce wastewater with toxic compounds. This is problematic because of the color of the water, toxic complexes, suspended solids, lignin, chlorinated compounds, chemical oxygen demand, and biological oxygen demand. Xenobiotics of paper, pulp, and mill effluents can be degraded by different microorganism including bacteria, fungi and actinomycetes (Hossain and Ismail, 2015).

Bacillus cereus and two strains of Pseudomonas aeruginosa have been found to decolorize bleached kraft paper-mill effluents (Tiku et al., 2010). Pseudomonas putida and Acienetobacter calcoaceticus degrade black liquor from a kraft pulp and paper mill. Their color removal efficiency was around 80% after 8 days (Abd El-Rahim and Zaki, 2005). A few algal species, such as Microcystis spp., can decolorize diluted bleached kraft paper-mill effluents. Microcystis spp. removed 70% color within 2 months (Sharma et al., 2014). Research is required to determine the pathways to degrade industrial effluents with microbial strains. Moreover, research focusing on the use of mixed cultures for the detoxification of these compounds should also be prioritized. The in situ application of this novel technology should also be further evaluated.

Polycyclic aromatic hydrocarbons (PAHs) are organic pollutants, mostly originating from anthropogenic activities. Most are carcinogens and mutagens. Their longevity in terms of their persistence in the environment may be due to their complex structure, hindering their bioavailability. PAHs also have stronger adsorption potential on solid particles due to high hydrophobicity and solid-water distribution ratios (Johnsen et al., 2005). As a result of these qualities, they are difficult to degrade and pose a severe and long-lasting hazard to the environment. Long term contamination of petroleum products affects the microbial population and diversity in the soil (Galazka et al., 2018). One effective strategy is bioremediation, or the use of microbes for decontamination of complex organic compounds. Studies on the use of microbes for degradation of PAHs are summarized in Table 3.

However, PAHs can serve as energy source for microorganisms and consequently be converted into harmless or less toxic compounds. Bioremediation of PAHs is well-documented as cost-effective, feasible, and practically applicable natural process for the degradation of polyaromatic hydrocarbons in soil and water environments (Anwar et al., 2016). A variety of microorganisms, such as bacteria, fungi and actinomycetes can be used for bioremediation. However, bacteria are the most reliable in degrading PAHs in aquatic environments (Johnsen et al., 2005).

In sediments or soils, the efficiency of bacteria may be lower because of reduced bioavailability (Yuan and Chang, 2001), or alternatively improved due to the involvement of solid-phase bacteria in the degradation of sorbed contaminants (Hwang et al., 2003). The latter argument is supported by the work of Poeton et al. (1999), which showed the enhanced biodegradation rates of PAHs (phenanthrene and fluoranthene) by marine bacteria in the presence of sediments. However, further studies need to be conducted in order to evaluate the efficiency of bacteria in bioremediation of PAHs under a range of conditions. The effect of high sediment contents on biodegradation of PAHs in natural water bodies was studied by Xia et al. (2006). An increase in the population of PAH degrading bacteria in a water system corresponded with increasing sediment content.

PAHs meant to be degraded by bacteria must also be bioavailable (Dandie et al., 2004; Fredslund et al., 2008). Dissolution and vaporization make them bioavailable, but sorption on soil particles makes it difficult for bacterial degradation due to the limitations of the microbes under these conditions (Kim et al., 2007). The bioavailability of PAHs is affected by a number of factors, including soil structure, moisture contents, and presence of bacterial species. Uyttebroek et al. (2007) reported that desorption and ultimately decomposition of PAHs is affected by their age in soil. The rate of decomposition is also affected by the presence of readily available nutrient and carbon sources. Decomposition of PAHs becomes more difficult if the enzymes involved are non-specific. The enzymes would prefer to attach to substrates, which are easier to degrade, allowing PAHs to persist in the environment (Wang et al., 2009). The presence of intermediate compounds can favor decomposition of PAHs. Sphingomonas, Nocardia, Beijerinckia, Paracoccus, and Rhodococcus were found by Teng et al. (2010) to decompose anthracene completely in the presence of dihydriol, an initial oxygenated intermediate compound.

PAHs can also be degraded by fungal enzymes. Various aerobic and anaerobic fungi can be used for the degradation of PAHs (Aydin et al., 2016). Bjerkandera adusta, Pleurotus ostreatus, and Phanerochaete chrysosporium have the ability to produce ligninolytic enzymes applicable to PAH degradation, such as laccase, Mn peroxidase and lignin peroxidase (Kadri et al., 2016). The rate of degradation depends on culture conditions, such as the presence of oxygen, temperature, availability of nutrients, and in an agitated or shallow culture. The efficiency of fungal enzymes is also affected by addition of biosurfactants. The in situ biodegradation of PAH is more complicated due to heterogenic nature of soils (Kadri et al., 2016). Therefore, all factors affecting the bioavailability of PAHs should be considered when using fungi for the degradation of these compounds.

The PGPR also have the potential to degrade PAHs. Pseudomonas sp. JPN2 is able to degrade pyrene and other aromatic contaminants (Jin et al., 2016). Its degradation potential increased with increasing incubation time, and maximum degradation (83%) was achieved after 25 days of incubation. The analysis of culture medium on GC-MS showed four metabolites (1-hydroxy-2-naphthoic acid, 4-phenanthrol, 4,5-dihydroxy-4,5-dihydropyrene and phthalate). Additionally, JPN2 showed increased growth at 100 mg L⁻¹ of pyrene, indicating its tolerance for pyrene (Jin et al., 2016). Meena et al. (2016) isolated pyrene degrading bacteria from industrial effluent contaminated sites which suggest these sites may be attractive niches for PAH degrading bacteria. From the industrial effluent, six bacterial isolates were isolated and purified following serial enrichment techniques, and were assessed for their pyrene degradation potential using modified mineral salt medium supplemented with pyrene. Through 16S rRNA sequencing, the isolated bacteria were found to belong to four genera, i.e., Ochrobactrum, Microbacterium, Bacillus, and Acinetobacter. The isolate Bacillus megaterium YB3 was further evaluated for its efficiency to degrade pyrene. Over 7 days of incubation, it degraded almost 73% of 500 mg L⁻¹ pyrene into two relatively less toxic intermediate metabolites, based on GC-MS analysis. A further characterization of B. megaterium YB3 showed that it was positive for aromatic-ring-hydroxylating dioxygenase indole-indigo and catechol 1,2-dioxygenase conversion assays.

Bacteria belonging to the genus Mycobacterium are among the most effective degraders of PAHs. Hennessee and Li (2016) studied four Mycobacterium species with respect to single and mixed culture PAH metabolism. Four PAH compounds (phenanthrene, fluoranthene, pyrene, and benzo[a]pyrene) were used as model compounds to characterize the degradation potential of bacteria in a strain- and mixture-dependent manner. The results indicated metabolic differences between single and mixed degradation of PAH, which could be helpful for risk assessment and bioremediation of PAH-contaminated sites. Chen et al. (2016) investigated functional microbial populations and processes involved in pyrene biodegradation in soils. Two soils were incubated with 60 mg kg⁻¹ of pyrene, more than 80% of the added pyrene degraded within 35 days in both soils. Thirty-five days after incubation, a significant enrichment of gram-positive bacteria harboring PAH-ring hydroxylation dioxygenase (PAH-RHD_α GP) genes was observed along with an increase in Mycobacterium. Both soils showed a large proportion of uncultured gram-positive bacteria along with Mycobacterium, suggesting they may be the important groups of pyrene degraders in soils. The genes, nidA and nidA3, were identified as the major genes involved in biodegradation following two different pyrene catabolic pathways (Chen et al., 2016).

The above studies point to the viability of microbes to degrade PAHs. Both bacteria and fungi are able to degrade various organic contaminants, including pesticides, azodyes, and petroleum hydrocarbons. In some cases, microbes use PAHs as energy and nutrient sources, converting them into harmless or less toxic compounds. In some cases, they rely on alternate energy sources and degrade pollutants without any direct benefits. Therefore, the addition of an alternate carbon source may accelerate the bioremediation process. Bioremediation is equally effective in liquid (water) and solid (soil) environments. However, it is affected by different biotic and abiotic factors, such as the type of microorganism involved, microbial population and diversity, structure of substrate, bioavailability of substrate, pH, aeration and alternate energy sources (Sihag et al., 2014). Microbes have different genes related to the degradation of PAHs, and consequently follow different pathways for decomposition of these compounds (Table 4). A number of intermediate compounds are produced depending upon the nature of the original pollutant and the type of microbe involved. These intermediate compounds can be even more toxic, therefore the bioremediation system should be continuously monitored. Although a lot of work has already been carried out in the field, more extensive studies using mixtures of contaminants and microbial consortia are needed. Further exploration of the specific mechanisms and genes involved needs more attention.

Long-term exposure of certain microorganisms to agrochemicals can result in resistance. Therefore, they can be used in bioremediation of pesticide-contaminated sites (Khan et al., 2009b). Their performance substantially increases in the presence of specific pesticides as they use them as a nutrient and energy source (Qiu et al., 2009; Reddy et al., 2016). Their ability to degrade pesticides is essential for eliminating harmful, toxic chemicals from the environment, controlling and reducing environmental pollution (Surekha et al., 2008).

Pesticide degradation is possible through bioremediation as an alternative to conventional methods, and is a more effective, versatile, environment-friendly, and economical strategy (Finley et al., 2010). Certain microorganisms have the ability to biodegrade pesticides through specific enzymes. Enzyme catalyzed biodegradation of pesticides is a complex process that involves a series of biochemical reactions. However, the rate and final fate of different pesticides in the environment varies greatly and depends upon biotic and abiotic factors. For example, due to structure, some pesticides, such as dichlorodiphenyltrichloroethane (DDT) became more persistent and remain in the environment for longer periods than others (Kannan et al., 1994). A variety of microbes, including fungi, bacteria and actinomycetes, are able to degrade and remove pesticides from the environment (Parte et al., 2017). Bioremediation can be carried out by indigenous microflora, or through the enrichment of microbial cultures.

Bioremediation is also affected by environmental factors, such as temperature, pH, nutrients, and moisture availability. Consequently, environmental conditions should be conducive for the growth and activity of microbes for accelerated bioremediation (Vidali, 2001). The long-term application of pesticides can also promote biodegrading enzymes in the indigenous microflora, as they served as a source of carbon and energy, making the remediation of pesticide contaminated sites easier (Qiu et al., 2009). Pesticide degrading microorganisms have been isolated from a wide variety of polluted environments. Some of them have potential to degrade more than one organic pollutant. Hence, scientists are working on microbial diversity and functioning to understand their physiology, diversity, ecology, and evolution for in situ biotransformation of organic contaminants (Mishra et al., 2001).

As of this point, most studies have used pure cultures of microbes to observe biodegradation of pesticides. This involves the isolation of microbes from contaminated sites, their purification and characterization for degradation of different rates of specific pesticides under laboratory conditions. For example, some stains of microbes for biodegradation of DDT have been isolated from soil, sewage, animal feces, activated sludge and sediments (Johnsen, 1976; Lal and Saxena, 1982; Rochkind-Dubinsky et al., 1986). The in situ use of microbes for degradation of organophosphates in the pioneering work of Matsumura et al. (1968) showed that Pseudomonas sp. was capable of degrading dieldrin in soil in natural environments. The degradation of certain pesticides requires more than one bacterial strain. The use of this strategy is known as co-metabolism. In this process, one bacterial strain converts the original compound to an intermediate that is ultimately converted in to a final product having less or even no potential hazards for the environment. For example, the biodegradation of DDT with an alternative carbon source, which microbes used as a nutrient or energy source to transform DDT, involves co-metabolism (Bollag et al., 1990).

Neurotoxic systemic insecticides not only persist for long periods in the soil, but also are non-selective and kill beneficial insects. They are water-soluble and persist in food chains and biogeochemical cycles. There are several bacterial strains with the ability to biodegrade neonicotinoids in soil and water systems (Hussain et al., 2016). Organochloride pesticides are recalcitrant and are more resistant to biodegradation (Díaz, 2004). Bacteria are the major group among microorganisms with the potential to degrade organochloride pesticides. The degradation ability of soil bacteria from the genera Arthrobacter, Pseudomonas, Bacillus, and Micrococcus was reported by Langlois et al. (1970).

Fungi are equally important in the degradation of organochlorine pesticides. Xiao et al. (2010) conducted a study on fungal degradation of heptachlor and reported that after 14 days of incubation P. acanthocystis, P. brevispora, and P. tremellosa removed about 90, 74, 71% heptachlor, respectively. They proposed that hydrolysis and hydroxylation are the dominant reactions involved. Similarly, Ozdal et al. (2016) reported that P. aeruginosa G1, Stenotrophomonas maltophilia G2, B. atrophaeus G3, Citrobacter amolonaticus G4, and Acinetobacter lowffii G5 have high biodegradation ability for the organochlorine, endosulfan. Ortega-Gonzalez et al. (2015) conducted an experiment with marine fungi extracted from marine sponges. The fungi Penicillium raistrickii, Trichoderma sp., Aspergillus sydowii, Penicillium miczynskii, Bionectria sp., and Aspergillus sydowii were tested in solid culture medium containing 5, 10 and 15 mg of dichlorodiphenyldichloroethane (DDD), and in a liquid medium, which had the same amounts of DDD per 100 mL of liquid medium. They reported that A. sydowii, Trichoderma sp., and P. miczynskii grow well in the presence of the pesticide; however Trichoderma sp. performed best and was selected for further studies.

The concentration and form of the alternate carbon sources may affect the degradation potential of microbes. Fang et al. (2010) tested glucose, yeast extract, sucrose, and fructose as additional carbon sources to foster the degradation of DDTs using Sphingobacterium sp. and found the half-lives of DDTs significantly decreased in the presence of additional carbon sources. Glucose as an additional carbon source resulted in the fastest biodegradation of DDTs. They confirmed these in vitro results through field studies. Results showed a significantly lower concentration in the samples inoculated with Sphingobacterium sp. as compared to uninoculated control after 90 days of incubation.

Microbes play a significant role in the degradation of the nematicide oxamyl. However, little information exists regarding the types of microbes involved in its biotransformation. Rousidou et al. (2016) isolated four oxamyl-degrading bacterial strains with the potential to enhance biodegradation of oxamyl in soil. Based on the multilocus sequence analysis (MLSA), the bacterial strains belong to different subgroups of the genus Pseudomonas. They hydrolyzed oxamyl to oxamyl oxime, but did not use it as a carbon source, instead utilizing methylamine as source of C and N. Three of the four strains contain methylamine dehydrogenase enzyme. Furthermore, all these strains also have a gene highly homologous to a carbamate-hydrolase gene, cehA, which has been found in carbaryl- and carbofuran-degrading bacterial strains. A number of bacterial strains are responsible for the degradation of carbamates, such as carbofuran and carbaryl (Bano and Musarrat, 2004; Yan et al., 2007) as well as methomyl (Mohamed, 2009; Xu et al., 2009). Aldicarb is degraded by Stenotrophomonas maltophilia (Saptanmasi et al., 2008) and Aminobacter and Mesorhizobium spp. are also considered to be oxamyl-degrading bacterial strains (Osborn et al., 2010).

The biodegradation of organophosphorus pesticides has been extensively studied (Singh, 2008). A variety of enzymatic systems in bacteria degrades them, for example Acinetobacter sp., Serratia sp., Proteus vulgaris, and Vibrio sp. have been reported to degrade dichlorovos (Agarry et al., 2013). Malghani et al. (2009) isolated bacteria efficient in the degradation of profenofos from the genus Pseudomonas. Rayu et al. (2017) demonstrated that Xanthomonas sp. 4R3-M3 and Pseudomonas sp. 4H1-M3 were able to use both chlorpyriphos and 3,5,6-trichloro-2-pyridinol as a sole carbon and nitrogen source under laboratory conditions. Similarly, degradation of cyhalothrin along with other pyrethroides by B. thuringiensis was reported (Chen et al., 2015). Chanika et al. (2011) identified two bacterial strains, Pseudomonas putida and Acinetobacter rhizosphaerae, which were able to rapidly degrade organophosphate fenamiphos (FEN). Both strains hydrolyzed FEN to fenamiphos-phenol and ethyl-hydrogen-isopropylphosphoramidate, although it was only further transformed by P. putida. The two strains use FEN as a C and N source.

Due to a decrease in effectiveness of individual pesticides, mixtures containing different active ingredient groups are being developed, especially the combination of pyrethroid and organophosphorus pesticides (Moreby et al., 2001). Genetically engineered microorganism (GEM) inoculants have the potential to degrade these complex pesticides. For example, Zhang et al. (2016) succefully engineered a multifunctional Pseudomonas putida X3 strain by introducing methyl parathion (MP)-degrading gene. They reported that genetically engineered strain X3 is a strong bioremediation agent that showed competitive advantage in complex environment contaminated with MP and Cd. Yuanfan et al. (2010) suggested the development of GEMs could be effective for bioremediation of contamination by multiple pesticides, as indicated by GEM with the methyl parathion hydrolase gene, mpd, which is responsible for hydrolyzing methyl parathion to p-nitrophenol and dimethyl phosphorothioate. The development of GEMs and dual-species consortia has the potential to be used for degrading different pesticides.

Wang et al. (2018a) studied endophyte Neurospora intermedia, isolated from sugarcane roots grown in a diuron-treated soil. Diuron is a broad-spectrum phenylurea herbicide for pre-emergence weed control in a wide bunch of crops. Their results indicated that the strain DP8-1 was capable of degrading up to 99% diuron within 3 days under optimal degrading conditions. The degradation spectrum of DP8-1 inlcuded fenuron, monuron, metobromuron, isoproturon, chlorbromuron, linuron, and chlortoluron. Its on-site applicability, however, needs further investigation.

Microbes that can degrade more than one group of pesticides would be more efficient and economic than those with specific traits. Above all, strains with multiple plant growth promoting traits, such as the ability to solubilize zinc, promote phosphate and chitinase activity, with a high root colonization potential, and biodegrade pesticides would be most effective due to their multi-purpose applicability. As these strains can efficiently colonize the plant roots and help plant roots to proliferate, phytoremediation is more feasible and makes inoculation with these microbes an economical and applicable strategy for the remediation of pesticide-contaminated sites.

Recently, the potential of GME to degrade or accumulate contaminants is also discussed. Their impact is much wider than that of their wild relatives, improving degradation or alteration of catabolic pathways, either to protect the host plant against phytotoxicity or to improve their overall efficiency of phytoremediation. This is especially suitable when hydrophilic compounds fail to be degraded by rhizospheric microbes due to the rapid uptake by plants (Ijaz et al., 2016).

Pollution to soil, water, and air is caused by release of inorganic chemical waste by industries, auomobiles, construction companies, and fertilizers. Inorganic pollutants mainly include heavy metals which may be detoxified by using microbes in the presence or absence of plant systems. Heavy metals may be beneficial or harmful for microbes, depending upon their nature and bioavailability (Ayangbenro and Babalola, 2017). For example, some heavy metals like manganese (Mn), Fe, nickel (Ni), Mg, copper (Cu), chromium (Cr), cobalt (Co), and Zn are essential micronutrients, required in a number of physiological processes, such as forming parts of enzyme complexes, redox reactions, and the stabilization of molecules through electrostatic interactions (Bruins et al., 2000). Other heavy metals, such as arsenic (As), antimony (Sb), lead (Pb), gold (Au), cadmium (Cd), Al, silver (Ag), and mercury (Hg), are not essential, and have no biological role in the microbial body (Bruins et al., 2000). In high concentrations, they can form various complexes in microbial bodies that are highly toxic. Even essential heavy metals like Zn and Ni can also be toxic at higher concentrations.

Some microbial strains develop resistance against these heavy metals and they have the ability to detoxify them. They could be a means to detoxify heavy metals at higher concentrations in the environment. Heavy metals and other ions must first enter the microbial cells for any indication of beneficial or harmful effects on microbial physiology (Nies, 1999). Many divalent heavy metals, e.g., Zn²⁺, Cu²⁺, Ni²⁺ Co²⁺, Fe²⁺, and Mn²⁺, are structurally similar, and anions of some heavy metals with oxygen resemble the anions of essential elements, for example chromate resembles sulfate, while arsenate resembles phosphate. As a result, the microbial uptake mechanisms need to be tightly controlled. Microbes use chemiosmosis, a gradient driven, very fast, and unspecific uptake systems for heavy metals (Nies, 1999) that increase their accumulation within the microbial body.

In microbial cells, toxicity may occur when heavy metals displace essential metals from their binding sites (Nies, 1999). They may also cause toxicity due to ligand interactions (Bruins et al., 2000). Heavy metals have the tendency to bind with sulfur-hydrogene (SH) groups in the microbial body and play a role in the inhibition of sensitive enzymes. The minimum concentration of heavy metals effective enough to bind with SH groups and inhibit enzyme activity is called the minimal inhibitory concentration (MIC). Some bacterial strains have an exceptionally high MIC, and therefore have a high resistance to heavy metals. For example, Yilmaz (2003) isolated and identified a bacterial strain Bacillus circulans EB1, with a high MIC for heavy metals. Bacterial strains with a higher MIC are preferable for bioremediation of heavy metal contaminated sites.

Other possible mechanisms of heavy metal resistance in microbes could be intra and extra-cellular sequestration, enzymatic reduction, biosorption, reduction in sensitivity of cellular targets to metal ions, and antioxidant defense system (Huang et al., 1990; Brady and Duncan, 1994; Liu et al., 2004; Xu et al., 2015). Microorganims release extracellular polymeric substances (EPS) which bind the heavy metals. Biosorption mechnisms used by EPS from Bacillus subtilis involve functional groups. Heavy metals, such as Cu(II) binds with anionic oxygen-bearing ligands and form inner-sphere complexes with the EPS functional groups as reported by Fang et al. (2014). These mechanisms can be useful in understanding the survival of microbes in this context.

Microbes remove heavy metal contaminants in different ways, such as biosorption, precipitation, biotransformation, bioaccumulation, complexation, enzymatic transformation of metals and phytoremediation (Liu et al., 2004; Ojuederie and Babalola, 2017; Xu et al., 2017). In bioaccumulation, microbes retain and concentrate heavy metals in their body. Bioaccumulating microbial strains can be strong candidates for decontamination of polluted soil and water as reported by Akhter et al. (2017). They isolated and identified three bacterial strains of B. cereus, i.e., BDBC01, AVP12, and NC7401, from rhizosphere of Tagetes minuta and reported that these strains have strong solubilization and accumulation potential for Cr(VI), Ni(II), and Cd(II) thus help in biosorption of these metals. Biosorption of heavy metals is the sequestration of positively charged metal ions by ionic groups on cell surfaces (Malik, 2004). Bacteria-clay mineral interactions are important in the context of metal immobilization and allocation of metals to mineral fraction. The adsorption-desorption mechanisms are affected by microbial composition and diversity, chemical behavior of metals, metal speciation and concentration, modeling method (Du et al., 2017a; Qu et al., 2017b), soil physico-chemical properties, such as clay minerals and soil pH, affecting the number of negative charge sites (Qu et al., 2017a) and plant species. The mobility of heavy metals in soils also depends upon type and concentration of ligands and sorbents, such as bacteria-mineral complexes (Du et al., 2016a). For example, citrate and humic acid enhanced Cd adsorption on P. putida–montmorilonite and P. putida–goethite composites while oxalate suppressed Cd adsorption. Phosphate ligand increased Cd sorption on P. putida–goethite while decreased on P. putida–montmorilonite composite. Recently, Qu et al. (2018) conducted a study on Pb sorption on montmorillonite-bacteria composites using a combination of atomic force microscope (AFM), X-ray diffraction (XRD), surface complexation modeling (SCM), Pb-LIII edge extended X-ray absorption fine structure (EXAFS) spectroscopy, and isothermal titration calorimetry (ITC). They observed that formation of montmorillonite, Pseudomonas putida complex promoted the allocation of Pb to mineral fraction and reported that SCM, EXAFS, and ITC may help in predicting the speciation and fate of Pb in soils and associated environments. For risk assessments in soils and associated environments, the heavy metal adsorption in complex systems is based on accurate modeling. In another study, Du et al. (2017a) investigated Cd adsorption on Gram-positive Bacillus subtilis, Gram-negative Pseudomonas putida and their binary mixtures with montmorillonite using SCM, Cd K-edge EXAFS, spectroscopy, and ITC. They reported that B. subtilis adsorbed more amounts of Cd than P. putida at pH < ~6 while P. putida was more efficient for Cd sorption to phosphate groups. This suggests that microbial composition and diversity along with biochemical behavior of trace metals are important for metal sorption in microbe-bearing environments. Qu et al. (2017a) established component additivity (CA) method with CA-site masking for Cd adsorption on goethite-Pseudomonas putida composites using different mass ratios. Both CA and CA-site masking models were in line with ICT data however, it was observed that CA method was excellent in simulating Cd adsorption on bacteria-iron oxides composites at low bacterial and Cd concentrations while wide deviation was observed at higher concentrations. Both models were supported by thermodynamic reaction data while these models are conditional to mineral/bacteria ratio and concentration and different models behave differently under different conditions and heavy metal concentrations (Wang et al., 2016). Moreover, interfacial complexation reations that occur between iron (hydr)oxides and bacteria should be taken into account for higher bacterial/metal concentrations.

Organic matter in complexation with iron minerals helps the adsorption of metals in soil environments. Soil microbes can further imrpove the adsorption when present in these multi-complexes. In soil environments, iron oxides make complexes with organic composites and help in the transformation trace metals. Du et al. (2017b) studied the copper adsorption on composites of synthetic goethite, cells of Pseudomonas putida, humic acid (HA), and their binary and ternary composites with batch adsorption experiments coupled with ITC. They reported that bacterial composites with goethite or humic acid separately and in combination help in adsorption and cycling of Cu but their affinity was less than binary composite of goethite and humic acid. Furthermore, binary and tertiary complexes of bacteria, iron oxide and humic substances affected the sequestration of heavy metals (Du et al., 2017a). During precipitation and transformation, microbes change the oxidation state of metals and metalloids to make them less harmful. The microbes used in bioremediation demonstrate a wide range of mechanisms, which change the bioavailability, transport properties, sorption characteristics and toxicity of heavy metals (Malik, 2004; Gupta et al., 2016). Metals show competitive adsorption for same type of adsorbents as they tend to bound on same types of adsorption sites on the adsorbent (Du et al., 2016b). Similarly, different microbial strains vary in their affinity for the sorption of heavy metals. For instance, fungal strain Paecilomyces lilacinus XLA was more efficient and eco-friendly than Mucoromycota sp. XLC for bioremediation of Cd²⁺ from wastewater (Xia et al., 2015).

Some studies aimed to isolate and screen metal-resistant microorganisms from polluted environments for bioremediation purposes (Pal and Paul, 2004; Abou-Shanab et al., 2007). For example, Akinbowale et al. (2007) isolated aeromonads and pseudomonads from fish, which were resistant to heavy metals. Srivastava et al. (2007) and Congeevaram et al. (2007) successfully used heavy metal resistant bacteria to detoxify heavy metal polluted sites. In the past decade, attention has been turned toward identification of bacterial strains with the potential to bioremediate polluted soils through the sequestration of toxic heavy metals and degradation of xenobiotic compounds (Braud et al., 2009; Hayat et al., 2010; Wani and Khan, 2010; Ahemad, 2012). Bioaugmentation of contaminated sites through efficient microbial strains can significantly reduce metal concentrations in polluted soil (Emenike et al., 2016; Fauziah et al., 2017). There are a number of bacterial strains belonging to the genera Pseudomonas, Bradyrhizobium, Psychrobacter, Ochrobactrum Lysinibacillus, Rhodococcus, and Bacillus, which have novel traits useful for heavy metal decontamination from polluted environments (Dary et al., 2010; Ma et al., 2010; Wani and Khan, 2010; Emenike et al., 2016). Fungi are also equally important in the remediation of metal polluted sites. For example, Xu et al. (2017) reported that Paecilomyces lilacinus XLA has the ability to reduce Cr⁶⁺ that can reduce over 90% of Cr⁶⁺ in growth media with Cr⁶⁺ concentration below 100 mg L⁻¹ at pH 6 after 14 days of incubation. They reported that XLA used biosorption, biotransformation, and bioaccumulation as the major mechanisms for reduction of chromium. The XLA might have detoxified the reactive oxygen species produced by external Cr⁶⁺ through its efficient antioxidant enzyme system. The efficiency of these fungal species can be affected by soil conditions, such as pollution level of metals, pH, EC, temperature, and nutrient status of soils.

Phytoremediation is based on hyper-accumulating plant species. The phytoremediation ability of plants depends on environmental conditions, the quantity of heavy metals present at the site, soil type, and microbial number and diversity (Ojuederie and Babalola, 2017). To accelerate the process, scientists are exploring plant–microbe interactions, combining the capabilities of rhizosphere bacteria to improve metal uptake by the plant. Kuffner et al. (2008) isolated ten rhizosphere isolates obtained from heavy metal accumulating willows (Salix sp.). They belonged to the genera Agromyces, Flavobacterium, Serratia, Pseudomonas, and Streptomyces. Plant growth, as well as Zn and Cd uptake potential of these strains were measured. The strain Agromyces AR33 almost doubled Zn and Cd extractability, which was attributed to the improvement of release of Zn and Cd specific ligands. Some other strains were helpful in improving Zn and Cd uptake of Salix caprea plantlets. However, they might have used different plant-microbe interactions to improve heavy metal uptake, except IAA production, ACC deaminase activity and siderophore production.

Some PGPR have the ability to improve heavy metal uptake in crops. For example, Ghasemi et al. (2018) conducted an experiment to study the effectiveness of five bacterial isolates with plant growth promoting traits on improving growth, health and Ni phytoextraction capacity of three Ni-hyperaccumulators, Odontarrhena inflate, O. bracteata, and O. serpyllifolia. Plants were inoculated with five rhizobacterial strains (previously isolated from O. serpyllifolia) and grown for 3 months. The bacterial inoculants enhanced Ni removal due to the stimulation in growth and/or increase in shoot Ni concentration, but the effectiveness of these strains varied with bacterial strain, plant species and soil type (Ghasemi et al., 2018).

Bacteria with ACC deaminase can potentially induce heavy metal stress tolerance in crop plants. Consequently, they could enhance the phytoextraction and phytoremediation potential of plants. Four bacterial strains from Ni-contaminated soils were isolated by Rodriguez et al. (2008) on the basis of ACC deaminase activity. They were identified as Pseudomonas putida Biovar B, and also have plant growth promoting traits, such as indole acetic acid and siderophores production, in addition to ACC deaminase activity. They were tolerant of up to 13.2 mM Ni in a culture medium. Based its effectiveness from the laboratory results, strain HS-2 was tested in pot experiments. It was observed that canola plants inoculated with HS-2 strain accumulated more biomass and had higher Ni contents in shoots and roots. According to these results, HS-2 could be a potential inoculant for the phytoremediation of Ni contaminated sites (Rodriguez et al., 2008).

Endophytic bacteria have also been shown to improve heavy metal stress tolerance in crop plants. For example, Sheng et al. (2008) isolated two Pb resistant endophytic bacteria, Pseudomonas fluorescens G10 and Microbacterium sp. G16, from the roots of canola plants grown in Pb contaminated fields. The isolated strains were resistant to heavy metals and successfully improved growth of canola plants in pot experiment. Zhang et al. (2011b) characterized three ACC deaminase producing endophytic bacteria, Ralstonia sp. J1-22-2, Pantoea agglomerans Jp3-3, and Pseudomonas thivervalensis Y1-3-9 out of 100 isolates isolated from copper-tolerant plants. They reported that these bacteria promoted plant growth and copper (Cu) accumulation in canola plants in a pot experiment This supports the endophytic bacterial-assisted phytoremediation strategy for Cu-contaminated environments.

As seen in the above studies, microbes can be part of an innovative strategy to remediate heavy metal contaminated soils. Under varying conditions and in different crops, bacterial strains showed the ability to improve plant growth, such as IAA production, ACC deaminase activity, siderophores production, and heavy metal uptake, through bioaccumulation, biotransformation, precipitation, and biosorption. However, there are still gaps in the understanding of specific plant-microbe interactions involved in the bioremediation of metal contaminated sites. There are bacterial strains with no IAA and siderospores production ability, and no ACC deaminase activity, which have enhanced metal uptake and accumulation in plant organs. Bioremediation technologies are necessary for the detoxification of metal-contaminants in polluted environments (Emenike et al., 2018a,b), to prevent their toxic effects on the environment and living organisms.

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.