Arsenate reduction and mobilization in the presence of indigenousaerobic bacteria obtained from high arsenic aquifers of the Hetaobasin, Inner Mongolia
Huaming Guo a, b, *, Zeyun Liu a, b, c, Susu Ding b, Chunbo Hao b, Wei Xiu a, b, Weiguo Hou a
a State Key Laboratory of Biogeology and Environmental Geology, China University of Geosciences, Beijing 100083, PR Chinab School of Water Resources and Environment, China University of Geosciences (Beijing), Beijing 100083, PR Chinac Shanxi Conservancy Technical Institute, Yuncheng 044004, PR China
a r t i c l e i n f o
Article history:Received 18 January 2015Received in revised form19 March 2015Accepted 23 March 2015Available online 8 April 2015
Intact aquifer sediments were collected to obtain As-resistant bacteria from the Hetao basin. Two strainsof aerobic As-resistant bacteria (Pseudomonas sp. M17-1 and Bacillus sp. M17-15) were isolated from theaquifer sediments. Those strains exhibited high resistances to both As(III) and As(V). Results showed thatboth strains had arr and ars genes, and led to reduction of dissolved As(V), goethite-adsorbed As(V),scorodite As(V) and sediment As(V), in the presence of organic carbon as the carbon source. Afterreduction of solid As(V), As release was observed from the solids to solutions. Strain M17-15 had a higherability than strain M17-1 in reducing As(V) and promoting the release of As. These results suggested thatthe strains would mediate As(V) reduction to As(III), and thereafter release As(III), due to the highermobility of As(III) in most aquifer systems. The processes would play an important role in genesis of highAs groundwater.
Arsenic is one of the most common and harmful carcinogens inthe environment. The problem of As contamination in groundwaterhas become a worldwide environmental issue (Ravenscroft et al.,2009). It directly endangers human health, and influences thesustainable development of society. Therefore, the source and fateof groundwater As is one of the hottest topics in the field of envi-ronmental science. Several biogeochemical processes control therelease and transport of As in natural waters, including adsorption,oxidationereduction, and microbe-mediated electron transfer(Diesel et al., 2009). More and more studies have shown that mi-crobes play an important role in the release, migration and trans-formation of As in aqueous systems (Oremland and Stolz, 2005;Anderson and Cook, 2004; Chang et al., 2008; Pepi et al., 2007;Lievremont et al., 2009; Chang et al., 2012; Mirza et al., 2014).
Although As is generally toxic to life, a lot of microorganisms can
es and Environment, ChinaR China.
get the energy for growth through metabolizing As (Oremland andStolz, 2003). These microorganisms have evolved the necessarygenetic components which confer resistance mechanisms,including arsenite-oxidation, arsenate-reduction and As(V) resis-tance minimizing the amount of As that enters the cells (Cervanteset al., 1994; Ji and Silver, 1992a,b). Additionally, the microorganismscan use As compounds as electron donors or electron acceptors,and possess As detoxification mechanisms (Ahmann et al., 1994;Johnson et al., 2003). These microorganisms are taxonomicallydiverse and metabolically versatile, mainly including the dissimi-latory arsenate-respiring prokaryotes (DARPs), chemoautotrophicarsenite-oxidizing bacteria (CAOs), heterotrophic arsenite-oxidizing bacteria (HAOs), and arsenate-resistant microbes(Oremland and Stolz, 2005; Anderson and Cook, 2004; Chang et al.,2008; Pepi et al., 2007). Since As(III)-oxidizing bacteria have firstlybeen found by Green (1918), a series of As(III)-oxidizing bacteriahave been isolated (Anderson et al., 1992; Santini et al., 2000;Kashyap et al., 2006; Muller et al., 2007; Duquesne et al., 2008).
Arsenic(V)-reducing bacteria are one of the dominant bacterialgroups involved in the cycle of As (Lievremont et al., 2009). A seriesof As(V)-reducing bacteria have been isolated in previous studies(Santini et al., 2004; Handley et al., 2009; Chang et al., 2012;
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Anderson and Cook, 2004; Liao et al., 2011). Microbial As(V)reduction occurs via two common mechanisms, respiration anddetoxification. Respiration occurs only under anaerobic conditions.During the respiration, the arrA gene, which encodes for a reduc-tase, catalyzes respiratory As(V) reduction (Saltikov and Newman,2003; Afkar et al., 2003; Muphy and Saltikov, 2009). Detoxifica-tion is an efficient As(V)-reducing mechanism that occurs underboth aerobic and anaerobic conditions (Silver and Phung, 2005;Murphy and Saltikov, 2009). During the detoxification process,the arsC gene, which is responsible for the biotransformation ofAs(V) to As(III) (Krumova et al., 2008; Musingarimi et al., 2010), andthe arsB, which extrudes As(III) from the cytoplasm (Rosen, 2002;Silver and Phung, 1996), would be involved. Furthermore, organo-arsenicals are presumed to be formed via microbial activities(Cullen and Reimer, 1989). Bacteria have a function of As methyl-ation, i.e., having arsenite methylation transferases that transfer As(III) to As methyl products, such as MMA, DMA (Oremland andStolz, 2003; Hong, 2006; Cai and Wang, 2009).
The Hetao basin, as one of the important commodity grain basesin China with more than 10,000 km2 of farmland and 1 millionresidents, has been severely affected by high-As groundwater (Guoet al., 2008a, 2014a). Groundwater has been used as a source fordrinking water and irrigation water (Guo et al., 2012, 2013a). Thishas led to many cases of chronic As poisoning in this area (Jin et al.,2003). Previous studies have investigated characteristics of sedi-ment, high As groundwater distribution, chemical and isotopiccharacteristics, and migration and transformation of groundwaterAs (Guo et al., 2008b, 2010, 2011a, 2011b, 2012, 2013a, 2013b).However, few studies have been conducted to investigate in-fluences of microbiological processes on As speciation and mobilityin the aquifers of the Hetao basin (Li et al., 2014). Li et al. (2014)observed that Pseudomonas, Dietzia and Rhodococcus widelyoccurred in aquifer systems and anaerobic NO�
3 -reducing bacteriaPseudomonas sp. was the largest group, followed by Fe(III)-reducing, SO2�
4 -reducing and As(V)-reducing bacteria, althoughbacterial diversity was dependent on groundwater As concentra-tion. However, the roles of indigenous aerobic bacteria in As cyclingare unclear and need more investigation. Therefore, it is importantto understand As resistance of indigenous aerobic bacteria andtheir roles in As transformation and mobility. The objectives of thisstudy are to 1) characterize effects of As on indigenous aerobicbacteria from aquifer sediments hosting high As groundwater, 2)evaluate As(V) reduction in As(V)-containingmedia in the presenceof the indigenous aerobic bacteria, and 3) assess As mobilizationduring bacteria-solid-water interactions.
2. Materials and methods
2.1. Sample collection
Sediment samplesM14 andM17were collected at depths of 12.2and 25.0 m, respectively, from the borehole (40�58.0200N,107�00.5210E) in the Hetao basin, Inner Mongolia, in 2009. Imme-diately after removal from the borehole, the sediment sampleswere put into sterilized plastic bags and sealed under pure N2 gas(N2 > 99.999%). Theywere transported at 4 �C to the laboratory, andstored in a refrigerator at �20 �C for experiments. M17 is gray sand,and used for isolation of aerobic As-resistant bacteria, since graysand universally occurs in aquifers hosting high As groundwater(Guo et al., 2008a). M14 is brown clay with 100% As as As(V) spe-cies, which allowed us to investigate the reduction of sedimentAs(V). Therefore, it was used as the solid phase in microcosmexperiments.
In comparisonwithM14,more SiO2, but lower Al2O3, Fe2O3, As, Fand Mn were observed in sediment M17 (Table 1). Arsenic content
was 88.2 and 7.9 mg/kg in M14 and M17, respectively. At the deptharound 25 m, groundwater had high As concentrations, with As(III)of 528 mg/L and As(V) of 54.0 mg/L.
2.2. Isolation of aerobic As-resistant bacteria
Aerobic As-resistant bacteria were cultivated in CDM medium(Text 1 in Supplementary Materials). Sediment sample M17 (10 g)was mixed with 3 mL solution with 1.5 mg/L As(III) or As(V), whichwas filtered by 0.22 mm filter membrane to removemicroorganismsbeforehand, in 100 mL glass flasks sealed with sterilized air-permeable polypropylene membranes (Thermo ScientificABgene). The mixtures were incubated for 10 d in a shaking waterbath at 150 rpm at 28 �C. In order to keep the humidity of samples,sterilized water was added each 1e2 days (Zhao, 2009). After that,50 mL sterilized water was added to the mixtures. The suspensionwas incubated for 1 h at 150 rpm at 28 �C. The supernatant wasdiluted to 10-2e10�5. For each dilution, 100 mL supernatant wasadded to CDM plates. The plates were incubated for 2 d at 28 �C.Then, single strains were selected and plate streaking method wasused to obtain pure bacteria (Liao et al., 2011). The pure strainswerestored in 4 �C.
2.3. Identification of aerobic As-resistant bacteria
The aerobic As-resistant strains were identified, using the 16SrRNA and As marker gene sequence analysis method. Sequences of16S rRNA gene and As marker genes are shown in Table 2. After theDNA extraction, bacterial 16S rRNA gene was amplified using bac-terial universal primers. Degenerate primers used to amply the Asmarker genes were designed especially for arrA, arsB, and arsC. Theamplification procedure was provided in details in Text 2 of Sup-plementary Materials.
The clone sequencing was detected by the Beijing ZhongKe XilinBiotechnology Company Limited. The DNA sequencing results wereanalyzed for similarities and aligned in the BLAST program pack-ages (http://blast.ncbi.nlm.nih.gov). The NCBI (USA) was used asthe reference database to identify 16S rRNA gene and amplified Asmarker gene sequences.
2.4. Experimental procedure
The pure strains were cultured in the CDM medium. The strainsuspensions were used in microcosm experiments. To examineeffects of bacteria on As(V) reduction and mobilization, four As(V)sources were prepared. One is dissolved As(V) with As concentra-tions of 0.5 and 7.5 mg/L (Treatment I); one is As(V)-adsorbinggoethite (a-FeOOH) with As content of 5.6 mg/g (Text 3 in Sup-plementary Materials) (Treatment II); one is crystalline scorodite(FeAsO4$2H2O) (Treatment III), and the other sediment M14 withAs content of 88.2 mg/kg (Treatment IV). In each treatment, blank
Table 2The primer sequences used for the PCR of arrA gene and ars genes.
Targeted gene Primer Primer sequences (from 50 to 30) Reference
16S 27F AGA GTT GAT CCT GGC TCA G Lane (1991)1492R TAC GGT TAC CTT TTA CGA CTT
arrA arrA-F AAGGTGTATGGAATAAAGCGTTT Malasarn et al. (2004)arrA- R CCTGTGATTTCAGGTGCC
arsC amlt-42-F TCGCGTAATACGCTGGAGAT Sun et al. (2004)amlt-376-R ACTTTCTCGCCGTCTTCCTT
arsB arsB-f ATGGCAACCGAAAGGTTTAG Anderson and Cook (2004)arsB-r GTTGGCATGTTGTTCATAAT
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batches (free of the strains) were also carried out to investigateeffect of the strains, which were sterilized at 121 �C for 20 min. Thesolideliquid ratio in Treatments II, III was 0.1 g: 50 mL, and Treat-ment IV 1 g: 50 mL. The ratio of bacterial suspensions to mediumwas 1: 50 (w/w), which ensured the similar initial microbial pop-ulation. All batches were incubated at 28 �C at 150 rpm. Eachexperiment was carried out in duplicate.
Dissolved As solutions were prepared from sodium arsenite forAs(III) and sodium arsenate for As(V). Arsenic(V)-adsorbinggoethite was synthesized as described in Text 3 of SupplementaryMaterials. Crystalline scorodite was obtained from a mineral com-pany in Guizhou province, PR China, which contained 100% scor-odite (Fig.S1 in Supplementary Materials). In the Sediment M14,100% of As was As(V) species as shown by XANES spectra (Fig.S4 inSupplementary Materials).
2.5. Analytical techniques
Dissolved As and Fe were analyzed by ICP-MS (7500C, Agilent).Arsenic species (including inorganic As(III), inorganic As(V), MMAand DMA) were determined using an HPLC-ICP-MS. For total As, thedetection limit was 0.1 mg/L, and the relative standard deviation(RDS) was less than ±2%. The detection limit for As(III) and As(V)was 0.2 mg/L, and the relative standard deviation (RDS) was lessthan ±2%. The pH values were measured using pHmeter (Sartorius,PB-21), dissolved Fe(II) using spectrophotometer (DR 2800, HACH)with 1, 10 phenanthroline method. Analytical details of sedimentsamples were provided in Text 4 of Supplementary Materials.
Bacterial growth was monitored by measuring the optical den-sity (OD) of the cultures at 600 nm with spectrophotometer(Shanghai Precision Scientific UV752N). Concentrations of ATPwere determined by ATP fluorescence detector (AF-100, DKK-TOA).
Arsenic K-edge X-ray absorption near-edge structure (XANES)spectra were recorded at room temperature at beamline BL15U1 ofthe Shanghai Synchrotron Radiation Facility (SSRF), China. Detailswere provided in Text 4 of Supplementary Materials. Since nodissolved Fe(II) was detected in all incubation batches, Fe K-edgeXANES spectra were not recorded for solid samples.
3. Results and discussion
3.1. Characteristics of aerobic As-resistant bacteria
Two strains of aerobic As-resistant bacteria, M17e15 andM17-1,
Table 3The arrA and ars genes of strains M17e15 and M17-1 by PCR amplification using thedesigned primers.
Isolate
Genes Strain M17-15 Strain M17-1
arrA þ þarsB e e
arsC þ þ
were isolated and identified. The 16S rRNA gene sequences of thetwo strains showed the highest similarity to Bacillus flexus strainJMC-UBL 24 (HM451429.1) with 99% similarity and Pseudomonasstutzeri strain Cpa_a4 (Sas, 2009) with 99% similarity, respectively.The high similarity suggested that the bacterial strains M17e15 andM17-1 were Bacillus sp. and Pseudomonas sp., respectively.
The arsC gene was successfully amplified (Table 3). The length ofputative arsC gene fragments was 326 bp. Phylogenetic clusteranalysis showed that the arsC of strain M17-15 was phylogeneti-cally located close to the arsC of strain M17-1, and their arsC genesequence was placed together in the cluster of arsC family. Theclosest arsC sequence of strainsM17e15 andM17-1 was unculturedbacterium clone ZJ-arsC-40 arsenate reductase (arsC) gene, withthe similarity of 93%. The arsB gene was not observed in these twostrains.
The arrA gene was detected in these two isolates (Table 3). A148 bp putative arrA gene fragment of the two strains were suc-cessfully amplified. The sequences of strains M17e15 and M17-1had the high similarity to uncultured bacterium arrA gene forAs(V) respiratory reductase.
3.2. Effect of As on bacterial growth
Fig. 1 shows the growth of strains M17e15 and M17-1 in treat-ments with different As concentrations. The logarithmic growth ofstrain M17-15 began at 24 h for all treatments (Fig. 1a). Themaximum biomass was reached at around 48 h in controls withoutAs amendment, and at around 72 h in treatments with 0.5 mg/LAs(V)/As(III). However, high As treatments with 7.5 mg/L As(V)/As(III)) obtained the maximum biomass at longer incubation time(around 120 h). Although the incubation times obtaining themaximum biomass were different, the maximum biomasses wereidentical for all treatments. Therefore, addition of As(III) and As(V)had no obvious inhibitory effect on the growth of strain M17-15,although the growth rate was retarded by the presence of As.
Strain M17-1 experienced the logarithmic growth at the incu-bation time between 12 and 60 h (Fig. 1b). The biomass reached themaximum at around 60 h, although the maximum biomass wasdifferent for different treatments. Controls and treatments with7.5 mg/L As(III) had the largest biomass, followed by treatmentswith 7.5 mg/L As(V) and treatments with 0.5 mg/L As(V)/As(III).Although As(III) shows 60 times more toxicity than As(V) (Korteand Fernando, 1991), As(III) did not exhibit evident inhibitory ef-fect on the growth of strains in comparison with As(V). Those re-sults indicated that both strain M17-15 and strain M17-1 had aresistance to dissolved As(III) and As(V).
Other study showed that the P. stutzeri strain As-1 isolated in thelaboratory grew in a culture mediumwith 50 mM As(V) or 0.2 mMAs(III) (Joshi et al., 2008). Joshi et al. (2008) found that, comparedwith the growth in the medium with the absence of As, the straingrowth in the media with 50 mM As(V) or 0.2 mM As(III) decreasedby 43% and 56%. Liao et al. (2011) found that, under aerobic con-ditions on MSM agar plates, the lowest As(III) concentration that
Fig. 1. Growth of strains M17-15 (a) and M17-1 (b) in the treatments with different As concentrations (Treatments: , free of As; D 0.5 mg/L As(III); B 0.5 mg/L As(V);: 7.5 mg/LAs(III); C 7.5 mg/L As(V)).
Fig. 2. Biological reduction of As(V) to As(III) with different initial As(V) concentrations: (a) strain M17-15 with 0.5 mg/L As(V); (b) strain M17-15 with 7.5 mg/L As(V); (c) strainM17-1 with 0.5 mg/L As(V); (d) strain M17-1 with 7.5 mg/L As(V).
H. Guo et al. / Environmental Pollution 203 (2015) 50e59 53
completely inhibited the growth of the Bacillus sp. (AR-9) andPseudomonas sp. (AR-2) is 5 mM and 2 mM, respectively, and thelowest As(V) concentration is 200 mM and 100 mM, respectively.Since As concentrations in our experiments were far less than thelowest inhibition concentration, either As(V) or As(III) showed nosignificantly inhibitory effect on the growth of strains.
3.3. Biotransformation of dissolved As
Results showed that As(III) concentration did not change withtime in the presence of strains M17e15 andM17-1 in culture media
with As(III) (data not shown). No As(V) was detected, indicatingthat strains M17e15 and M17-1 did not oxidize dissolved As(III).
However, As(V) was reduced to As(III) in the presence of strainM17e15 or M17-1 in growth media. Fig. 2 shows variations in Asspecies during experiments with strain M17e15 or M17-1 and0.5 mg/L or 7.5 mg/L As(V). During the experiments, no methyl Asproducts (such as MMA, DMA) were detected. Strain M17-15reduced As(V) quickly at the beginning of logarithmic growth.The strain M17-15 completely transformed 0.5 mg/L As(V) to As(III)within 6 h (the incubation time between 24 and 30 h) (Fig. 2a).Initial As(V) concentrations did not have a significant influence on
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reduction duration (p < 0.05). When initial As(V) concentrationincreased to 7.5 mg/L, the time for transformation of all As(V) toAs(III) was still 6 h (between 24 and 30 h) (Fig. 2b).
Arsenic(V) was reduced in the batches with the presence of theM17-1 more slowly than in the batches with the presence of theM17-15. All dissolved As(V) was reduced to As(III) by strain M17-1within 12 h (the incubation time between 12 and 24 h) at initialconcentration of 0.5 mg/L (Fig. 2c), and within 24 h at initial con-centration of 7.5 mg/L As(V) (the incubation time between 12 and36 h) (Fig. 2d). Concentrations of ATP showed that the microbialactivity of the strain M17-1 was systematically higher than that ofthe strain M17-15 in all batches (Fig.S2 in Supplementary Mate-rials). Therefore, in comparison with strain M17-15, strain M17-1had a lower reduction rate for dissolved As(V).
As shown above, both strains M17e1 and M17-15 had the arsCgene. The arsC gene encodes the enzyme for As(V) reductase, whichis responsible for the biotransformation of As(V) to As(III)(Krumova et al., 2008; Musingarimi et al., 2010). After As(V)reduction, the arsB gene acts as a specific efflux pump, which ex-trudes As(III) from the cytoplasm (Rosen, 2002; Silver and Phung,2005). The very likely absence of arsB, as observed in our isolates,probably indicated gene variations that decreased the homology ofthe primer sets. In this pathway, As(V) reduction would be adetoxification process, whereby the bacteria reduced As(V) toAs(III), which would be more easily pumped out of the cell (Maityet al., 2011). Some studies have shown that Pseudomonas sp. andBacillus sp. reduced As(V) under anaerobic conditions. They allbelong to the facultative anaerobic bacteria (Wu et al., 2013).
The arrA gene was also observed in strains M17e15 and M17-1,
Fig. 3. Variation of dissolved total As (a), Fe (b), As species (c), pH (
which encodes for a reductase that catalyzes respiratory As(V)reduction (Saltikov and Newman, 2003; Afkar et al., 2003; Muphyand Saltikov, 2009). It may express in our study. Although twostrains reduced 100% of As(V) at different initial As concentrations(0.5 and 7.5 mg/L), the reduction rates of As(V) by two strains weredifferent. Some researchers showed that the reduction rate wasdependent on initial As(V) concentrations, and incubation time(Maity et al., 2011). With the similar initial As(V) concentrations,and incubation time, less microbial activity and shorter reductiontime indicated the relatively higher reduction rate of strainM17-15.
3.4. Bioreduction of solid phase As
3.4.1. Reduction of As(V) in scoroditeResults showed that a small amount of As (0.5 mg/L) was dis-
solved into the solution from the scorodite in controls where themediumwas free of strains (Fig. 3a). Both strains M17e15 andM17-1 promoted the release of As from scorodite. In the presence ofstrains, dissolved As increased significantly. At the incubation timeof 10 d, dissolved As was around 14.5 mg/L with strain M17-1,which is lower than that with strain M17-15 (18.1 mg/L) (Fig. 3a).
In addition, As species were controlled by the presence ofstrains. In controls, only As(V) was detected in solutions, withAs(III) < 0.2 mg/L. Dissolved As(V) was expected to originate fromscorodite dissolution. Harvey et al. (2006) showed that concen-tration of dissolved As was around 450 mg/L at pH 6.0 at 22 �C,although incongruent dissolution occurred above pH 3. In experi-ments, As(III) concentration increased and reached the maximumat 4 and 2 d for strains M17e15 and M17-1 batches, respectively
d) with incubation time in the system of culture and scorodite.
Fig. 4. Variation of dissolved As species with incubation time in the system of cultureand As(V)-adsorbing goethite.
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(Fig. 3b). It indicated that As(V) was reduced to As(III) in solutionsand/or solids in the presence of the strains. The reduction of As(V)to As(III) would be the main reason for the enhanced release of Asfrom scorodite. The reduction rate of As(V) by strain M17-15 isgreater than strain M17-1 (Fig. 3b). This is similar to the reductionof dissolved As(V). The pH value increased from8.4 to 9.5 during 6 dincubation (Fig. 3d). After 6 d, the growth of strains was inhibited,and their activity decreased due to high pH (around 9.5). Therefore,As(III) concentration gradually decreased to below detection limitfrom 6 to 10 d, during which As(V) concentration increased. Thevariations in As(III) and As(V) concentrations would be As(III)oxidation by dissolved oxygen in solutions (DO around 6.2 mg/L).
In the presence of strains, Fe was also released from the solid,although no Fe(II) was detected in solutions (<0.01 mg/L). Dis-solved Fe was 0.17 mg/L with strain M17-1, and 0.18 mg/L withstrain M17-15 (Fig. 3c). However, dissolved total Fe is far less thandissolved total As. The reason is that dissolved Fe was precipitatedas Fe hydroxide colloids at high pH (around 9.5), which was evi-denced by high total Fe (including dissolved and colloidal forms) insolutions (4 mg/L). The low Fe(III)/As(V) and high pHwould explainthe observation that As(V) concentration increased after 6 d eventhough strain activity was inhibited by high pH. Krause (1989)showed that the solubility of As in scorodite is related to the Fe/As and pH, with the higher solubility at the higher pH and lower Fe/As.
3.4.2. Reduction of goethite-adsorbed As(V)The synthesized goethite had As content of 5.6 mg/g Fig. 4
shows change of dissolved As species in the batches. The adsor-bed As(V) was desorbed in control batches without strains. Thedesorbed As reached around 2.1 mg/L at 2 d and exhibited a rela-tively stable concentration after 2 d (Fig. 4). The desorbed Asaccounted for 19.6% of total As adsorbed on goethite. O'Reilly et al.(2001) also found that after the initial rapid desorption, only a smallamount of additional desorption occurred from goethite at longertime. Besides, a control with inactive cells showed that the deadcells did not evidently affect As(V) desorption (data not shown).
In the presence of strains, dissolved As concentration increasedin the system of culture and As(V)-adsorbing goethite. After 20 dincubation, total dissolved As concentrations were 4.0 and 3.75mg/L in batches with strains M17e15 and M17-1, respectively. If wesubtracted the desorbed As obtained in controls, the released As by
strain M17-15 accounted for 16.1% of total As on goethite, and bystrain M17-1 13.8%. Therefore, strain M17-15 induced more Asreleased from As(V)-adsorbing goethite than strain M17-1.
During 0e4 d incubation, only As(III) was detected with unde-tectable As(V) in solutions of batches with strains M17e15 andM17-1. It indicated that As(V) was reduced to As(III) by strainsM17e15 and M17-1. After 4 d, both As(III) and As(V) concentrationsincreased gradually. The increase in As(III) concentration was theresult of As(V) reduction by strains. However, the increase in As(V)concentration would be caused by the increase in solution pH dueto desorption of As(V) from the surface of goethite at high pH. Inthis study, pH value increased from 8.4 to 9.5 during 6 d incubation(date not shown). In addition, As(III) re-oxidation by dissolvedoxygen in solutions would lead to the increase in As(V) concen-trations (DO around 6.0 mg/L) although the decrease in As(III)concentration was not detected. Grafe et al. (2001) showed thatadsorption of As(V) on goethite was dependent on solution pH,with the lower adsorption at the higher pH. Both As(III) and As(V)concentrations kept relatively stable after 16 d, due to the limitedstrain activity. During the experiments, Fe(II) concentrations insolutions were below the detection limit in both batches (<0.1 mg/L), which indicated that goethitewas not reduced in the presence ofstrains M17e15 or M17-1. Therefore, results of Fe species alsosupported that desorption of As(V) and/or reductive desorption ofAs(III) contributed to the release of As in the presence of the strains.
Arsenic species in goethite were determined by XANES, whichare shown in Fig. 5. At 0 d incubation, no As(III) was detected, andAs(V) was the only As species in goethite. In the presence of strainM17e15, As(III) accounted for 8% of total As at 4 d incubation, whichincreased to 21% after 16 d incubation (Fig. 5a). In comparison withstrain M17-15, less As(III) was observed in goethite from the systemin the presence of strain M17-1 (Fig. 5b). At 4 d incubation, As(III)accounted for 1% of total As, after which relative proportion ofAs(III) generally kept constant (around 6%). Considering dissolvedAs species and As species in goethite during incubation, the pres-ence of strain M17-15 led to a higher As(III) content in goethite, andfurther released more As from goethite, relative to the presence ofstrain M17-1. Accordingly, it suggested that reduction of adsorbedAs(V) should be the major cause of As release from goethite, inaddition to desorption of As(V).
At 0e4 d, only As(III) was detected in solutions, concentration ofwhich was generally equal to As desorbed from goethite. Therefore,As(V) reduction was expected to occur in solution at this period.After 4 d incubation, As(III) was detected on the surface of goethite,and As(III) content in goethite generally increased with the increasein incubation time (Fig. 5). Arsenic(III) content of goethite increasedat 4e16 d in the presence of strain M17-15, while between 4 and12 d in the presence of strain M17-1. Increases in As(III) concen-trations in both solution and goethite indicated that As(V) reduc-tionmay directly occur on the surface of goethite. After reduction ofadsorbed As(V) to As(III), As(III) was partly released into solutionand partly retained on goethite. After that, both dissolved As(III)and As(V) kept relative constant. Due to the limited activity ofstrains, relative proportion of As(V) and As(III) in goethite keptstable after 16 d in the presence of strain M17-15, and after 12 d inthe presence of strain M17-1.
3.4.3. Release and transformation of sedimentary AsFig. 6 shows change of dissolved As concentration in batches
with sediment M14. In controls without strains, As release wasobserved from the sediment at the first 2 d. After that, concentra-tion of dissolved As showed relatively constant, which kept ataround 125 mg/L (Fig. 6). However, strains M17e15 and M17-1promoted the release of As from the sediment into solutions.With the strains, higher As concentration was observed at the first
Fig. 5. Arsenic K-edge XANES spectra of goethite in batches with the presence of strains M17-15 (a) and M17-1 (b).
Fig. 6. Variation in dissolved As concentration with the incubation time in the systemof culture and sediment (M14).
H. Guo et al. / Environmental Pollution 203 (2015) 50e5956
2 d than that in controls. After 2 d incubation, dissolved As con-centration gently increased with the incubation time in batcheswith strains (Fig. 6). Dissolved As concentrations were 200 and 175mg/L with strains M17-15 and M17-1, respectively, at 6 d. It indi-cated that As was released from the sediment with the presence ofstrains M17e15 and M17-1. Since solution pH kept relatively con-stant during experiments (7.0 ± 0.2), As desorption due to pHvariation would not be the cause for the increase in As concentra-tion in the incubation batches. Furthermore, dissolved As concen-tration is higher in the batches with strain M17-15 than in thebatches with strain M17-1. It demonstrated that strain M17-15promoted As release from the sediment more apparently thanstrain M17-1, which is consistent with As release from scoroditeand As(V)-loading goethite. Since clay minerals (64%) are the majorphases in M14, followed by quartz (20%), calcite (12%), feldspar (2%)and dolomite (1%), As would be mainly fixed to clay minerals andcalcite. Therefore, As was expected to be released from clay
minerals and/or calcite in M14 in the presence of the strains.In controls, it was found that As(V) was the only As species in
solutions. It may indicate that As(V) was desorbed from the sedi-ment after sterilization. However, only As(III) was detected in so-lutions of batches with strains (Fig.S3 in Supplementary Materials),although DO content was around 5.0 mg/L in suspensions. XANESspectra showed that As predominantly occurred as As(V) in thepristine sediment (M14) (Fig.S4 in Supplementary Materials).Therefore, both strainsM17e15 andM17-1 led to As(V) reduction toAs(III). Two possible pathways were expected for the release of Asfrom the sediment in the presence of the strains. One is that As(V)in the sediment was firstly released into solution and then reducedto As(III) with the presence of the strains. The other is that thestrain firstly reduced sediment As(V) to As(III) in-situ, and then theproduced As(III) on the surface of sediment was released into so-lution due to low affinity of As(III) species to mineral surfaces (Guoet al., 2007; Manning et al., 1998). Due to the absence of dissolvedAs(V), the plausible pathway included reduction of sediment As(V)and As(III) mobilization, which is also supported by the data fromexperiments with scorodite and As(V)-loading goethite. Mirza et al.(2014) also observed As(V) reduction and mobilization from sedi-ments by dissimilatory arsenate-respiring bacteria. During the in-cubation, the As released in the presence of strains M17e15 andM17-1 accounted for 17.6% and 12% of total As in the sediment,respectively. This As fraction would be the exchangeable and/oradsorbed As in the sediment. At the end of incubations, dissolvedAs concentration generally kept constant, demonstrating that therelease of As was not obvious due to the limited bacterial activity(2.0e3.5 nmol/L ATP). It showed that the activity of bacteria had asignificant effect on the release of As from the sediment (Xie et al.,2011).
3.5. Environmental implication
Arsenic(V)-reducing bacteria are of significance in controlling Ascycling in groundwater environment. Our previous study showedthat anaerobic bacteria, including NO�
4 -reducing and As(V)-reducingbacteria, were widely present in the aquifer sediments from theHetao basin (Li et al., 2014). In addition to anaerobic As(V) reducingbacteria, which led to As mobilization due to the higher mobility ofAs(III) than As(V) (Li et al., 2014), anaerobic Fe-respiring bacteria
H. Guo et al. / Environmental Pollution 203 (2015) 50e59 57
(IRB) would stimulate the reductive dissolution of Fe(III) oxides/oxyhydroxides, and therefore the release of As (Guo et al., 2008b;Kar et al., 2010; Selim et al., 2010a,b; Rowland et al., 2009). Thisstudy showed that aerobic As(V)-reducing bacteria,Pseudomonas stutzeri strain and Bacillus flexus strain, also led toAs(V) reduction and subsequently As release from As(V)-adsorbingminerals, As(V)-containing crystalline, and natural sediment in oxicconditions. Bachate et al. (2009) showed that As(V)-reducing bac-teria potentiallymediated As transformations in soils, and their rolein As cycling may become relevant with changing environmentalconditions.
Two microbial pathways were related to As(V) reduction toAs(III). One is As(V) respiration reduction, which is associated witharr genes. Arsenic(V)-respiring genes have been found in Geobacter(Reyes et al., 2008), Shewanella (Murphy and Saltikov, 2009), andClostridium (Mirza et al., 2014). Mumford et al. (2012) concludedthat the solubility and transport of As in shallow groundwater wasinfluenced by microbial respiratory reduction of As(V) to As(III).The other is As(V) detoxification, which is related to ars genes. It hasobserved that Escherichia coli (Wu and Rosen, 1993), Pseudomonas(Cai et al., 1998), Acidithiobacillus (Butcher et al., 2000), Coryne-bacterium (Ordonez et al., 2005), Staphylococcus (Lin et al., 2007),and Halobacterium sp. Strain NRC-1 (Wang et al., 2004) had arsgenes. As shown in strains M17e15 and M17-1 in this study, She-wanella sp. strain ANA-3 also had both detoxification ars genes andAs(V)-respiring arr genes. Although several environmental condi-tions affected the expression of respiration arr and detoxificationars As(V) reductase genes (Saltikov et al., 2005), As(V)-reducingmicrobes would regulate As transformation and migration ingroundwater systems.
No Fe(III) reduction was observed in this study, although theinvestigated As(V)-reducing bacteria caused As mobilization fromthe sediment via As(V) reduction. Accordingly, they led to the in-crease in dissolved As concentration, but did not affect dissolved Feconcentration. It may be used to explain the fact that dissolved Aswas decoupled with dissolved Fe in most high As groundwaters(Selim et al., 2010a,b; Lu et al., 2011; Xie et al., 2012; Guo et al.,2011a, 2014b).
After reducing As(V) in the solids, those strains would result inthe release of As(III) from the solids in oxic conditions. Therefore,As(III) may be observed in oxic aquifers in the presence of As(V)-reducing bacteria. Some studies have reported a relative As(III)abundance to total As in groundwaters in oxidizing conditions(Abdullah et al., 1995; O'Reilly et al., 2010). The ratio of As(III) toAs(V) can vary greatly as a result of variations in the abundance ofredox-active solids, especially the activity of microorganisms(Smedley and Kinniburgh, 2002). Therefore, aerobic As(V)-reducingbacteria are the plausible causes for the fact that As(III) occurred infacultative oxic aquifers.
4. Conclusions
Two strains, isolated from the aquifer sediment with high Asgroundwater from the Hetao basin, included Bacillus sp. M17-15and Pseudomonas sp. M17-1. Strains M17e15 and M17-1 grew inculture media with high As concentrations under oxic conditions.They had a good resistance to both As(III) and As(V). The strains hadAs(V)-respiring arr and As(V) detoxification ars genes, and led toreduction of dissolved As(V), adsorbed As(V), crystalline As(V) andsediment As(V) to As(III). The reduction rate of solid As(V) waslower than dissolved As(V) in the presence of the strains. Thereduction promoted the release of solid As into solution. In com-parison with strain M17-1, strain M17-15 had the higher As(V)reduction rate and therefore the more apparent release of As fromsolids into solutions. It was suggested that the indigenous
microorganisms would affect As transformation in groundwatersystems, and play an important role in the formation of high Asgroundwater.
Acknowledgment
The study has been financially supported by the National Nat-ural Science Foundation of China (Nos. 41222020 and 41172224),the National Key Basic Research Development Program (973 Pro-gram, No. 2010CB428804), the Geological Survey Program of ChinaGeological Survey (No. 12120113103700), the FundamentalResearch Funds for the Central Universities (No. 2652013028 and2652012028), and the Fok Ying-Tung Education Foundation, China(Grant No. 131017). The authors would like to thank the ShanghaiSynchrotron Radiation Facility (Beamlines BL14W1 and BL15U1)and its staff (Z. Jiang, S. Zhang, X. Yu, and A. Li) for allowing us toperform the EXAFS and XANES analyses. Constructive commentsfrom anonymous reviewers and editor (Dr. Bernd Nowack) aremuch appreciated.
Appendix A. Supplementary data
Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.envpol.2015.03.034.
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