When coal is combusted for electricity generation in a thermal power plant’s boiler; the carbon is consumed leaving behind molten particles rich in silica, alumina, and calcium that are driven out of coal-fired boilers together with the flue gases. This fine gray colored powder that makes up more than half of the coal left over is called fly ash (FA) as they “fly” up into the exhaust stacks of the power plant until get captured by electrostatic precipitators. A heavier portion of coal ash that settles on the ground of the boiler is known as the bottom ash (10% of total coal ash) (Hower et al. 2017). According to annual report on FA generation-utilization of Central Electricity Authority of India, approximately 160 million tones of FA is presently generated in India annually with utilization rate of approximately 60.97% (CEA 2014). Presence of trace amounts of heavy metals (HM) including arsenic, beryllium, cadmium, barium, chromium, copper, lead, mercury, molybdenum, nickel, radium, selenium, thorium, uranium, vanadium, and zinc make FA unhealthy for human. FA is disposed of either by dry or wet methods. In dry disposal, dry FA is dumped in landfills and FA basins. In the wet method, the FA is mixed with water and dumped in slurry form into artificial lagoons as pond ash. Both the methods ultimately lead to dumping of FA on open lands, which not only destroys the land but also contaminates surrounding water bodies, ground water, and agricultural lands due to leaching of the hazardous HMs (Pandey et al. 2011). Among the different HMs, leaching of As is particularly important in Indian context as various parts of India, including this Gangetic basin is already suffering from groundwater As pollution from the geogenic origin. Arsenic known as the silent killer is a semi metallic element commonly found in two major speciation forms: pentavalent arsenate [As(V)] and trivalent arsenite [As(III)]. [As(III)] is more mobile, biologically available and toxic to living organisms than [As(V)] (Roy et al. 2015). Long-term exposure of As contaminated food or water may lead to various diseases like conjunctivitis, hyperkeratosis, hyperpigmentation, cardiovascular diseases, and skin cancer (Smedley and Kinniburgh 2002). Leaching studies of 18 elements from the Indian coal FA by Praharaj et al. (2002) showed that As is one of the elements that leaches out of the ashes in excess of World Health Organization’s (WHO) limit values. Arsenic amount exceeding permissible limits has been reported in soil, tube-well and surface water near thermal power plants (Pandey et al. 2011). Burn et al. (2006) studied natural attenuation of As in soil coming from FA landfills and found that [As(V)] exhibits greater adsorption than [As(III)]. Another physiochemical attenuation study of As in soil coming from FA landfills showed that most of the As was strongly sorbed (60%–90%) or irreversibly bound (10%–40%) with < 5% readily desirable (Hyun and Lee 2013). Soil chemicals play important roles in the sorption process. Another way of combating HM pollution by nature is by employing metal resistant microbes that can detoxify the surrounding environment by transforming toxic metals to a less toxic form or bio-absorb them (Gadd 2004). Although extensive literature reviews are present regarding the bioremediation ability of the microbes towards As (Roy et al. 2015), none of them studied natural As bioremediation in ash landfill sites. Arsenic resistant aerobic microbes possessing [As(III)] oxidizing genes are widespread in nature and can oxidize the [As(III)] to less toxic [As(V)] and help in bioremediation of As impacted environments (Mukhopadhyay et al. 2002). Since As leaching from FA dumps is deleterious to human health, the purpose of this paper was to investigate whether FA dumping sites undergo natural attenuation of As by As resistant bacteria. The study made an attempt to understand number of As resistant microbes among the total heterotrophs of FA dump sites and to isolate As resistant strains and characterized them to understand their mechanism of As tolerance. A microcosm study was conducted to (i) understand the impact of As spiking on fly ash microbes under natural multi metal environment, (ii) examine the leachate for As content to get an indication on the As immobilization/accumulation ability of the indigenous microbes and (iii) evaluate the possibility to extrapolate data from stable laboratory incubations to field applications.

Materials and Methods

Dry ash and pond ash were collected in sterile containers from DVC-Mejia Thermal Power Station located at Durlovpur, Bankura, WB, India (23.28°N and 87.08°E) (Fig. 1).

Fig. 1
figure 1

Collection of dry ash (a) and pond ash (b)

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The pH, electrical conductivity, and salinity of the dry ash and pond ash were analyzed. Arsenic contents in the two FA samples were analyzed by graphite furnace atomic absorption spectroscopy (AAS). For As analysis, 1 mL of sample was first pre-reduced by the addition of 2 mL reducing solution composed of 1 mL of concentrated HCl and 1 mL of 10% (w/v) potassium iodide. The solution was allowed to stand for 30–45 min at room temperature for complete conversion of [As(V)] to [As(III)]. In comparison to As(V), unstable [As(III)] form is more amenable to volatile hybrid generation and so this form is required in AAS for As quantification. Detection parameters were as follows: wavelength, 193.7 nm; slit width, 1.2 nm; lamp intensity, 6.0 mA. The optimized parameters for the ashing and atomization steps were kept as below: temperatures of 500–1200 and 2150°C, heating rates of 100 and 500°C s−1, hold times of 90 and 7 s, and medium then maximum and medium argon gas flows, respectively. The calibration plots were linear with a correlation coefficient of 0.9799.

For microbial analysis, an aliquot of 1.0 g dry FA sample and 1.0 g pond ash sample were added separately in two 500 mL flasks containing 200 mL of LB broths. 3 µg mL−1 nystatin was added to inhibit fungal growth. Dilution spread plating in LA plates was performed to get isolated colonies. Heterotrophic counting was performed. For isolation of As resistant strains, sodium arsenite (m-NaAsO2) was added at a concentration of 2 mM for enrichment in LB broth. Subculturing was made two times. Finally dilution spread plating was made onto 2 mM of [As(III)] containing and 5 mM [As(V)] containing agar plates. From there they were replica plated in 20 mM [As(III)] containing LA plates and 30 mM [As(V)] containing LA plates. Then strains that appeared in both of the two replica plates were again replica plated in multimetal containing plates having 1 mM of Cu, Fe, Zn, Pb, Ba, Cr, Co, Cd, Mn. Colonies that appeared in all these replica plates and differed morphologically among themselves were picked up. Colonies were checked for purity. With these sequential steps of filtering, finally 10 strains were sorted out from the initial ~ 100 starting colonies. The selected strains showed resistances to both forms of As and were tolerant to multiple HMs.

The pure colonies were studied for colony morphologies after Gram staining. Molecular identification was done by 16S rDNA analyzes with 10 ng of genomic DNA. PCR was done using primers 27f and 1492r, dNTPs and 1U of Taq polymerase (Roychowdhury et al. 2016). The PCR products were purified and send to SCI GENOME lab for Sanger sequencing. The 16 s sequence data obtained was submitted to GenBank. Consensus sequences of the strains were compared with others by using the BLAST program. For phylogenetic analysis, multiple sequence alignment was performed using ClustalW and evolutionary tree was constructed with MEGA 3.1 software package (Kumar et al. 2004). Resulting tree topology was evaluated by bootstrap analysis of 1000 re-samplings.

The optimum pH, temperature, and salt tolerance of the isolates were determined. All the experiments were performed in triplicates. For temperature optimization, 100 µL of overnight grown culture of each isolate was dispensed into the test tube containing 5 mL LB medium and incubated at 4°C, 25°C, 30°C, 37°C and 42°C for 24–48 h. For pH optimization, 5 mL of LB broth was dispensed into different test tubes and the pH was adjusted from 4 to 11. For investigation on salt tolerance, the strains were grown in LB medium amended with 2%–15% of NaCl. For pH optimization and salt tolerance studies, cultures were incubated for 24–48 h at 30°C for Bacillus and Brevibacillus and 37°C for Staphylococcus, Kytococcus and Micrococcus. To determine the effect of As and other toxic metals on the growth patterns of the strains, two strains that showed maximum resistances to As were selected. For each strain, three 100 mL cultures were set up. One culture was kept as control. 70 mM of [As(V)] and 25 mM of [As(III)] were added in the 2nd flask. Multiple HMs (2 mM concentration each for [As(III)], [As(V)], Cd, Cu, Cr, Ni, Pb, Mn, Ba, Zn, Fe) were added in the third flask. 0.25 mL inoculums were added and the cultures were incubated at 37°C for 72 h. The optical density of the culture, as a measure of microbial growth, was detected at a wavelength of 600 nm by an UV–Visible spectrophotometer.

[As(III)] and [As(V)] resistances in the isolated strains were evaluated by the minimum inhibitory concentration (MIC) tests by plate based and test tube based methods. The MIC values of the isolates were determined by gradually increasing the concentration of As from 20 mM to 40 mM for [As(III)] and 30 mM to 90 mM for [As(V)] on the nutrient agar plates or tubes. MIC to other HMs [CuSO4·5H2O (0–25 mM); CdCl2·5H2O (0–5 mmol); ZnSO4·6H2O (0–25 mM); Co(NO3)2·6H2O (2–15 mM); Pb(NO3)2 (0–15 mM); AgSO4 (0–2 mM); HgCl2 (0–2 mM), K2Cr2O7 (2–15 mM); FeSO4 (2–15 mM); BaCl2 (10–30 mM); MnSO4 (15–30 mM)] were measured individually in test tubes and LA plates. Tubes/plates were incubated at 30°C/37°C and observations for growth/colonies were monitored for 4 days. Growth was calculated by measuring OD value at 600 nm.

To determine the presence of oxidase enzyme which causes oxidation of arsenite to arsenate, agar plates were flooded with a solution of 0.1 M AgNO3 (Simeonova et al. 2004). In the case of arsenite oxidizing bacteria, brownish precipitate forms and in the case of [As(III)] reducing bacteria, bright yellow precipitate forms (Lett et al. 2001). The transforming abilities of the isolates were confirmed by microplate screening assay for the detection of arsenite oxidizing and arsenate reducing bacteria (Simeonova et al. 2004). 20 µL of resting cell suspension were added to the wells of a microtiter plate containing 80 µL of 0.2 M Tris–HCl buffer with [As(III)] or [As(V)] to reach a final concentration 20 mM and 30 mM respectively. 100 µL of 0.1 M AgNO3 was added to the wells and the inoculated microtiter plates were incubated for 4–7 days. The reaction between AgNO3 and [As(III)] and [As(V)] in Tris–HCl resulted in the formation of coloured precipitates, ranging from light brownish–red for [As(V)] to light yellow for [As(III)]. The quantitative estimation was performed by the molybdenum blue method (Rao et al. 1993). Strains that showed positive results on biochemical assay were grown to mid-exponential phase in LB without As and then inoculated into two vials each containing 20 mL of LB either with 10 mM [As(V)] or 5 mM [As(III)] to obtain an OD600 nm of 0.05. Control flasks were set up without inoculums to check abiotic transformation of As. At each sampling time, 1 mL of suspension was removed to measure cell growth at OD600 nm. Determination of the presence of [As(V)] and [As(III)] was performed by spectrophotometric method (Cummings et al. 1999). [As(V)] concentration was determined by acidifying 100 µL sample in 100 µL of HCl. A 100 µL of the acidified sample was added to 900 µL of the reaction mixture having the following composition: ammonium molybdate (6 gL−1), ascorbic acid (10.8 gL−1), potassium antimonyl tartrate (0.136 gL−1) and concentrated H2SO4 (67.3 mL). Samples were heated in a water bath at 78°C for 10 min and placed on ice for 5 min. The optical density at 865 nm was compared to acidified [As(V)] standards. [As(III)] concentration was determined by oxidizing a second sample in 100µL of KIO3 (5 mmol L−1) and HCl (48 mmol L−1) for 10 min and then taking the reading at OD865 nm. Standard curves were prepared for both [As(V)] and [As(III)]. The difference of optical density between oxidized and unoxidized samples represented the concentration of [As(III)]. For determining reduction of [As(V)] to [As(III)], the isolates were aerobically grown in LB medium without As and then transferred into vials with butyl rubber stoppers containing Tris Minimal Medium (TMM) with sodium lactate (10 mmol L−1) and [As(V)] (10 mmol L−1) under an atmosphere of N2 ⁄CO2 ⁄H2 (85:5:10). The vials were incubated in static conditions at 30°C for 7 days. After incubation, the samples were analyzed for growth at OD600 nm and As concentration were determined as described above.

Two EPS producing strains (showed mucoid phenotypes and secreted whitish precipitates when grown in LB broth containing 5% dextrose) were analyzed for dead biomass mediated As biosorption. Dead bacterial biomass was prepared by growing 1 L culture for 48 h in LB (amended with 5% dextrose) in presence of [As(III)]. It was centrifuged at 8000×g for 15 min, washed twice, re-suspended in sterile deionized water, and heated at 120°C for 45 min. The mass concentration of dead biomass was determined spectrophotometrically at 600 nm. The dead cell suspension was incubated at room temperature for 24 h under shaking in the presence of 5 mM [As(III)] or 10 mM [As(V)] and were then centrifuged at 8000×g for 15 min. The pellet was washed three times with sterile deionized water and dried at 70°C. Aqua regia and hydrofluoric acid digested biomass was sent for As detection by AAS.

In a microcosm study, four 50 mL column chromatography columns were used and they were filled (up to 40 mL) with 60 g pond ash slurry. In the first column, water saturated autoclaved ash slurry (autoclaved twice) was used. This served as the control column where no indigenous microbes were present. To confirm the absence of microbes, it was incubated for one week and then the suspension was placed on Luria agar plates. Absence of colony after 72 h of incubation confirmed the sterile nature of the slurry. In the second column, water saturated unautoclaved FA slurry containing its own indigenous flora was used. For the 3rd column, unautoclaved ash slurry was amended with microbial cell suspension (consortia of 10 the isolated microbes). In the 4th column, autoclaved slurry was amended with 30 mg mL−1 of crude EPS prepared from Bacillus subtilis HMR5. Absence of microbes in the 4th column would help to know the As binding ability of the EPS alone. Next, 10 mL solution containing 5 mM [As(III)] was added on the top of the columns and flow rates were kept to a minimum and constant. The leachate collected over 2 weeks period was subjected for analysis of As content.

Results and Discussion

Salinity, pH, electrical conductivity and As concentration of dry FA were 170 ppm, 7.3, 1.1 mS cm−1 and 2.8 mg kg−1 respectively. For pond ash, the above parameters were salinity 205 ppm, pH 7.8, conductivity 1.2 mS cm−1 and As concentration 1.5 mg kg−1. Figure 2 shows heterotrophic and As resistant microbial count in the two types of FA samples. Due to very low moisture content, number of heterotrophs was expectedly low in dry ash in comparison to pond ash. However, in comparison to pond ash, most dry ash heterotrophs (> 90%) were resistant to both forms of As.

Fig. 2
figure 2

Counts of heterotrophs in dry ash and pond ash. [As(V)]-tolerant (5 mM) and [As(III)]-tolerant (2 mM) bacteria expressed as CFU g−1 FA. Data are expressed as mean ± standard deviation of three replicates

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All the selected ten strains isolated from dry ash and pond ash were gram-positive. All the Bacillus species, including Brevibacillus were rod-shaped while all the Staphylococcal species and Kytococcus were coccal. Among them, Kytococcus sedentarius and Micrococcus species formed tetrads. Two strains (B. subtilis and Micrococcus) showed mucoid appearance in presence of 5% dextrose in LB. Accession numbers of the strains given by NCBI GenBank authority and identity of their closest neighbours according to 16srDNA sequence data are present in Table 1. Figure 3 shows the phylogenetic tree.

Table 1 Table shows the accession numbers, identity with nearest bacteria, MIC values of the ten strains towards [As(III)] and [As(V)]
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Fig. 3
figure 3

Phylogenetic tree of the isolates with some related strains. The number on each branch indicates a bootstrap probability (1000 replicates). Scale bar indicates the amount of nucleotide sequence divergence in substitutions per site

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The optimum temperature required for growth and As transformation was found to be 30°C for Bacillus and Brevibacillus and 37°C for Staphylococcus, Kytococcus and Micrococcus. Optimum pH was 7.0–8.0 for all. All the isolates, however, showed growth at all the pH values ranging from pH 4 to pH 9 and at all temperatures except 4°C. Salt tolerance ability differed among the strains and 3%–5% NaCl concentration gave optimum growth for all. Salt tolerance study showed that all the strains were halotolerant. Few strains like HMR1, HMR2, HMR11, HMR9 grow even at 12% NaCl concentration. Two bacteria S. haemolyticus strain HMR1 and Kytococcus sedentarious strain HMR6 were studied for understanding the effect of As and other HMs on their growth profiles as they displayed highest MIC values. Figure 4 shows the growth profile of the two isolates in the presence and absence of As. It was observed that during As stress the lag phase of both of the isolates extended as much as 6 h as compared to the control. The exponential phases were similar in both the strains, but Kytococcus reached a higher OD (1.75) than S. haemolyticus (1.5). The prolongation of lag phases may be due to the time taken for adjustment of the cells to combat against the toxic effect of As. In the case of Kytococcus, it was observed that once the cells were adapted to As stress by adapting some detoxifying strategy they showed even a better OD than LB grown culture.

Fig. 4
figure 4

Effect of arsenic on growth profiles of two selected strains S.haemolyticus HMR1 and Micrococcus sp. strain HMR9

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K. sedentarius strain HMR6 exhibited the highest resistance to both [As(III)] and [As(V)] resisting up to 30 mM of [As(III)] and up to 72 mM of [As(V)]. S. haemolyticus strain HMR1 showed the same resistance as the previous one for [As(V)] but was little lower for [As(III)].The MIC values of the isolates towards [As(III)] and [As(V)] have been shown in Table 1.

The As resistant bacterial strains were found to be resistant to other HMs. Figure 5 shows the MIC values of HMs of the 10 isolates.

Fig. 5
figure 5

Minimum inhibitory concentration of the isolates towards multiple HMs

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Silver nitrate based screening test and microplate screening assay confirmed that among the 10 strains, 6 strains were able to transform As from one speciation state to another. Among the 6 strains, HMR 6 and HMR9 produced yellow precipitate when their plates were flooded with AgNO3 while HMR16, HMR11, HMR2, HMR1 showed brownish precipitates. Same were confirmed during microtitre screening assay.

On quantitative estimation on the biotransformation of [As(III)] to [As(V)], it was found that K. sedentarius strain HMR6 and Micrococcus strain HMR9 both were completely reducing 10 mM of [As(V)] to 10 mM of [As(III)] within 6 days (Fig. 6a) and 4 days (Fig. 6b) respectively. The results also revealed that Brevibacillus borstelensis strain HMR16, S. haemolyticus strain HMR1, S. pasteuri strain HMR2 and S. hominis strain HMR11 oxidized 5 mM [As(III)] to [As(V)] completely within 3.5 days (Fig. 6c), 5 days (Fig. 6d), 3 days (Fig. 6e) and 4 days (Fig. 6f) respectively.

Fig. 6
figure 6

Bacterial growth and As transformation by the 6 isolates. Blue line indicates growth curve; red line indicates [As(V)]; green line indicates [As(III)]. Two isolates HMR6 (a) and HMR9 (b) were showing reduction (i.e. [As(V)] level falling with [As(III)] level increasing) and four isolates HMR16 (c), HMR1 (d), HMR2 (e), HMR11 (f) exhibited oxidation (i.e. [As(III)] level decreasing and [As(V)] level increasing). Data are represented as means ± standard error, n = 3

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Recently scientists are exploring many bacterial cells and specially halophiles that protect themselves from the infiltration of toxic HM ions by covering their cell surfaces with a thick layer of EPS. These exopolysaccharides may either be homopolymeric or heteropolymeric in composition and of diverse high molecular weights (10–1000 kDa). Structural and compositional makeup of EPS favours the sequestration of metal ions and hence obstructs them from penetrating the cell surface (Gupta and Diwan 2017). This study also got two such strains (Micrococcus sp. strain HMR9 and B. subtilis strain HMR5) that produced EPS. In most cases EPS producing strains show dead biomass mediated metal absorption (Gupta and Diwan 2017). So dead biomass mediated As sequestration was studied in these two strains to understand the role of cell bound EPS in bioabsorption of As. Atomic absorption spectroscopy of the digested cell pellet showed presence of 0.5 g kg−1 and 0.56 g kg−1 of As respectively in HMR5 and HMR9.

Leaching experiment set up to demonstrate bioremediation ability of the native microbes showed that the leachate from the control or first column contained 5.66 g kg−1 of As. Arsenic content in the leachate from 2nd and 3rd columns were 3.78 g kg−1 and 2.5 g kg−1 respectively. This showed that retention of As within FA was more in columns containing microbes. Exopolysaccharide impregnated FA column showed 4.5 g kg−1 of As in the leachate. FA is known to be an excellent absorbent of HMs (Li et al. 2009). So it is expected that most of the As added would be retained by the ash particles themselves. But the difference of As concentration in the leachate of control and test microcosms demonstrated the role of native microbes in As sorption along with FA particles. Highest sequestration of As was found in the 3rd column where number of microbes was also highest. This was followed by the 2nd column and 4th column where the As retention were contributed by the microbes and crude EPS of B.subtilis strain HMR5 in addition to FA particles.

Although a vast amount of literature is present on the existence of varied types of As resistant microbes in soil and water, scarce work has been done on their existence in FA samples. To the best of our knowledge, this is the first report where it has been shown that adverse environmental condition of FA dump sites has selected a good number of As tolerant bacteria with varied mechanisms of As resistance. All of them were also tolerant to multiple other metals suggesting their potential applicability for bioremediation of multi-metal contaminated areas. As dry FA contains less microbe due to use of high temperature and less nutrients availability, the enriched bacterial community in the pond ash may be partly contributed from the water that was used for preparing FA slurry (here Damodar river water was used for slurry preparation). Proof of the aquatic origin of some bacteria was reflected by our finding of one facultative halophilic K. sedentarius, known for its marine dwelling. HM analysis data showed pond ash contained less As than dry FA which may be due to microbial detoxification or dilution caused by mixing of water with dry FA for slurry preparation.

To understand natural attenuation of ash pond with these As tolerant native microbes, total heterotrophs versus arsenic resistant heterotrophs counting was performed. 62% of the heterotrophs isolated from pond ash was found to be resistant to 2 mM [As(III)] suggesting an active microbe mediated biogeochemical cycle of As going in the site. Biogeochemical cycle of As is mediated mainly by two types of active communities of microbes. One group of bacteria (mainly anaerobic in nature) mobilizes or increases bioavailable fraction of As in different compartments of environment (helpful for bioleaching, biomining and phytoaccumulation). This group (As mobilizing bacteria) possess respiratory arsenate reductase that reduces less toxic and less bioavailable [As(V)] to more toxic and more bioavailable [As(III)] and in doing so increase the environmental load of As (Oremland and Stolz 2003). Another large group of aerobic bacteria possessing arsenite oxidase gene decreases or detoxifies environmental As concentration by biotransformation of more mobile and toxic [As(III)] to less toxic and immobile [As(V)] (helpful for bioremediation) (Gadd 2004; Silver and Phung 2005). Fortunately, among our 10 isolates, none were As reducing except Kytococcus and Micrococcus. Four strains (HMR1, HMR2, HMR11 and HMR16) biotransformed 5 mM arsenite to arsenate completely within 3–5 days. Two strains (HMR9 and HMR5) showed exopolysaccharide mediated As sequestration offering great bioremediation potential. This third mode of As detoxification not explored very much is biomass or EPS mediated sorption and bioaccumulation. Although biomass mediated metal absorption has been studied in depth on other metals (toxic HMs like Cd2+, Hg2+, Zn2+, and Pb2+, precious metals like Au3+, Pd2+, and Ag+, base HMs like Co2+, Ni2+, and Cu2+ and radionuclides like U6+ and Th4+) but very few reports are there on EPS or biomass mediated anionic As sequestration (Muthu et al. 2017). In only one study EPS mediated sorption of As was noticed in one biofilm forming strain Herminiimonas arsenicoxydans that was able to detoxify arsenic-contaminated environments by oxidizing [As(III)] to [As(V)] and by scavenging arsenic ions in their EPS layer (Marchal et al. 2010).

Natural attenuation is viewed as a low cost remediation technique for contaminated soil and groundwater (Mulligan and Yong 2004). Microcosm experiments have long been recommended as a method for documenting natural attenuation of pollutants in soil or ground water or landfills and constitute a valuable supplement to field measurements. Microcosms mimic natural microbial habitats under conditions simulating those of the sample’s original location (Colombo et al. 2011). In the present study, the 15 days old column based leaching study was designed to understand the in situ As bioremediation ability of the native microbes of FA sample. Columns containing indigenous microbes (2nd column) and indigenous microbes augmented with added consortia (3rd column) documented less As leaching than untreated one indicating role of microbes in immobilization or bioaccumulation of As within FA. Thus the microcosm experiment gave idea that natural attenuation of As was occurring within FA sample and microbe mediated immobilization may be thought of as the principle process behind this. Less As leaching from the fourth column containing only the crude EPS than control column proved successful sequestration of As by the added EPS. However, more experiments are needed with purified EPS to fully understand and explore the potential of EPS on HM sequestration. In summary, this study showed that bioaugmentation of FA sites with various As resistant microbes including the isolated ones can enhance the rate of natural attenuation which can form a valuable tool for restricting leaching of As from FA dumps to underground drinking water reservoir.