Bioresource Technology 102 (2011) 10401–10406
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Acid tolerance of an acid mine drainage bioremediation system based on biological sulfate reduction Jian Lu a,b, Tianhu Chen a, Jun Wu a,⇑, P. Chris Wilson b, Xiangyang Hao c, Jiazhong Qian a a
School of Resources and Environmental Engineering, Hefei University of Technology, Hefei 230009, Anhui Province, PR China Indian River Research and Education Center, University of Florida, Fort Pierce, FL 34945-3138, USA c School of Material and Chemical Engineering, China University of Geosciences, Beijing 100083, PR China b
a r t i c l e
i n f o
Article history: Received 9 August 2011 Received in revised form 11 September 2011 Accepted 12 September 2011 Available online 17 September 2011 Keywords: Tolerance Bioremediation Sulfate Acid shock Acidic mine drainage
a b s t r a c t The acid tolerance response of an AMD bioremediation system based on sulfate reduction was investigated. Efﬁcient sulfate reduction was observed with a maximum sulfate reduction rate of 12.3 ± 0.8 mg L1 d1 and easily available organic carbon was released during high acid treatment with an initial pH of 2.0. The rapid increase in sulfate reduction was observed when the extreme acid treatment with an initial pH of 1.0 was stopped. Column experiment on acid shock showed that efﬁcient sulfate reduction was maintained while precipitation of Cu or Zn still occurred during extreme or high acid shock. More than 98% of Cu and 85% of Zn were removed in the high acid column experiment with inﬂuent pH of 2.0. The majority bacteria in the remediation system used for high acid drainage belonged to genera Clostridiaceae, Eubacterium, Pseudobutyrivibrio, and Clostridium. These ﬁndings showed high acid tolerance of the straw remediation system. Ó 2011 Elsevier Ltd. All rights reserved.
1. Introduction Mining and milling of sulﬁde ore generate large quantities of waste rocks and ﬁnely crushed mill tailings (Jurjovec et al., 2002; Egiebor and Oni, 2007) in which the exposure of the sulﬁde minerals to atmospheric oxygen may ultimately lead to the formation of acid mine drainage (AMD) (Egiebor and Oni, 2007). This process usually leads to the release of heavy metals to the surrounding environments (Younger et al., 2002; Egiebor and Oni, 2007). Lime neutralization/precipitation is often used for the treatment of AMD (Akcil and Koldas, 2006). During lime treatment, AMD is usually discharged into a rapid mix chamber where hydrated lime is added in dry form or as slurry. Then the AMD with high ferrous ion concentrations and have pH 8–10, are passed through an aeration tank, where the ferrous hydroxide precipitate is converted to ferric hydroxide, before being discharged into a settling chamber, where heavy metals are precipitated from the solution. AMD with low ferrous ion concentration (<50 mg L1) was treated to a pH of 6.5–8.0 and then diverted directly to a settling chamber (Akcil and Koldas, 2006). Although lime treatment is not expensive and produces a lower volume of sludge, it is not widely used largely because it is difﬁcult to raise the pH above 6 due to the buffer effects of carbon dioxide on the reaction (Akcil and Koldas, ⇑ Corresponding author. Tel.: +86 551 2901739; fax: +86 551 2901649. E-mail addresses: [email protected]
(J. Lu), [email protected]
(J. Wu). 0960-8524/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2011.09.046
2006). One attractive approach for decontamination of AMD is metal precipitation using biogenic sulﬁde (Egiebor and Oni, 2007; Zhuang, 2009). A lot of researches have been performed on AMD remediation systems based on biological sulfate reduction process, which consumes protons and produces sulﬁde, thus increasing the pH and decreasing heavy metal concentrations through formation of metal-sulﬁde precipitates (Hulshof et al., 2003; Ziemkiewicz et al., 2003; Egiebor and Oni, 2007; Jin et al., 2008; Zhuang, 2009). In such bioremediation system, the most important bioprocess is sulfate reduction facilitated by sulfate reducing bacteria (SRB), whose growth optimum pH range is between 6.8 and 7.2 (Rees et al., 1995; Sievert and Kuever, 2000; Rozanova et al., 2001). Acid shock in bioremediation reactors based on biological sulfate reduction might frequently occur due to the unsuitably low pH in AMD (Akcil and Koldas, 2006; Egiebor and Oni, 2007). In some cases, the pH of AMD below 2.0 has been reported (Morin et al., 1988). However, few studies have been performed on the acid tolerance response of AMD remediation system based on biological sulfate reduction. Efﬁcient biological sulfate reduction in such AMD remediation systems relies on the sufﬁcient supply of organic carbon to provide carbon and energy sources for SRB (Hulshof et al., 2003; Tsukamoto et al., 2004; Jin et al., 2008; Wu et al., 2010). For this reason, additional carbon source should be added to establish an AMD bioremediation system based on sulfate reduction (Hulshof et al., 2003; Jin et al., 2008). One method to promote sulfate reduction is adding carbon source directly to the tailings and promote in situ sulfate
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reduction to treat the AMD at source (Hulshof et al., 2003; Jin et al., 2008; Lindsay et al., 2009; Zhuang, 2009). The direct addition of an organic carbon source to the tailings has the potential to promote sulfate reduction and the subsequent removal of heavy metals from pore water in the tailings (Hulshof et al., 2003). However, as far as the in situ sulfate-reducing bioremediation is concerned, the acid shock problem becomes much more severe due to the severe pH variation of pore water in tailings impoundments and the frequently extreme acid shock during bioremediation. Acidic waters were generated in the unsaturated zone of a tailings impoundment where the water was usually extremely acidic (Jurjovec et al., 2002). On the basis of ﬁeld data, Morin et al. (1988) provided a diagram of pH changes with depth through a cross-section for the Elliot Lake uranium mill tailings. According to the conceptual model, the pH of the pore water near the acidic water generation zone is about 1 and increases with depth due to buffering by dissolution of a series of minerals in the tailing impoundment. Johnson et al. (2000) reported that the low pH plume of contaminated water in the tailings impoundment could move at one tenth of the average linear groundwater velocity. Low cost carbon sources such as woodchips and straw have been successfully used for promoting bacterial sulfate reduction in mine drainage remediation processes (Hulshof et al., 2003; Wu et al., 2010). According to our previous study, straw is more efﬁcient in promoting sulfate reduction than other low cost carbon sources such as woodchips for the bioremediation of low pH AMD (Wu et al., 2010). In this study, the acid tolerance response of an AMD bioremediation system using straw as carbon source was investigated. The objective of this study was to investigate the acid tolerance response of this remediation system based on biological sulfate reduction through batch and column experiments. The ﬁnal goal was to obtain initial information on the acid tolerance of the AMD bioremediation system based on biological sulfate reduction. 2. Methods 2.1. Materials and media Rice straw (Oryza sativa L.) was obtained as previously described (Wu et al., 2010). Taq polymerase, primers, and UNIQ-10 DNA puriﬁcation kits, were obtained from Sangon (Shanghai, China). All other reagents used were of reagent grade. The synthetic 1 AMD water contained 500 mg L1 SO2 Fe2+. Cu2+ 4 and 500 mg L 1 2+ 1 (1.5 mg L ) and Zn (1.5 mg L ) were added into the synthetic AMD water as heavy metals to simulate the real AMD environment. Cu and Zn were chosen as the typical heavy metals in the synthetic AMD since they are the most common heavy metals in AMD (Huisman et al., 2006). Samples were obtained by dissolving measured amounts of reagent grade chemicals. The pH was adjusted using 1 M HCl. Dissolved oxygen was removed by purging the medium with high purity N2 for at least 15 min. Anaerobic enrichment cultures obtained from the anaerobic serum bioreactor of previous study (Wu et al., 2010) were used as inoculum. 2.2. Batch experiments Bioremediation tests were carried out in 150 mL serum bottles. Synthetic AMD (90 mL) and anaerobic enrichment culture (10 mL) were added to each bottle. Serum bottles were capped with rubber stoppers and crimped with aluminum seals. The headspace of the bottles was high purity N2. The incubation was performed at 25 °C in darkness. Rice straw (1 g) was added as the carbon source. All treatments were performed in triplicate. Strict anaerobic microbial techniques were used throughout the experimental
manipulations. At each sampling point the cultures were vigorously shaken and sampled with sterile syringes ﬂushed with high purity N2. To study the acid tolerance response of biological sulfate reduction under different pH environments, the initial pH value of the media was adjusted to 1.0 (extreme acid), 2.0 (high acid), 3.0 (moderate acid), and 7.0 (control) at the beginning of the experiment. Hydrogen chloride (2 M), sodium hydroxide (2 M), and sodium carbonate was used for pH adjustment. To study the evolution of the SRB activity submitted to an extreme acid environment that evolved to a high or moderate acid environment, the initial pH was 1.0 (extreme acid). The pH was adjusted to 2.0 (high acid) or 3.0 (moderate acid) at day 7 for the short-term acid treatment test and at day 28 for the long-term acid treatment test. To study the effect of acid treatment on organic carbon release, the initial pH of synthetic AMD was 2.0. All incubations were performed at 25 °C in darkness. 2.3. Column experiments Column experiments were performed using the same column as previously described (Wu et al., 2010) with slight modiﬁcation. Brieﬂy, layers of clean river sand placed at the top and bottom of the column were replaced by clean quartz. Straw residue (10 g) from the previous column study (Wu et al., 2010) was mixed with new straw (100 g) and packed between the quartz layers to form an organic layer. The packed column was ﬂushed with high purity N2 before the experiment started. Oxygen-free AMD was pumped into an infusion bag to store the synthetic AMD under anaerobic conditions (Wu et al., 2008). To simulate the frequent acid shock during AMD remediation, the pH of the inﬂuent was adjusted from 3.0 to 1.0 at day 27. Then the pH value was adjusted back to 3.0 at day 38, followed by another pH adjustment from 3.0 to 2.0 at 60 d. In the case of bioremediation of high AMD, the pH of the column inﬂuent was maintained at 2.0 during the entire experimental period. Bioremediation of extreme AMD with a pH of 1.0 was not performed since sulfate reduction rate was low. 2.4. Analytical methods All aqueous samples analyzed were ﬁrst ﬁltered using a polytetraﬂuoroethylene (PTFE) ﬁlter unit (0.22 lm, Millipore). The pH was measured by using an Ultrameter II™ 6P (Myron L Company, USA). Samples were subject to inductively coupled plasma atomic emission spectroscopy (Iris Advantage 1000, Thermo Jarrell Ash Corporation, USA) for analyzing Cu and Zn. Sulfate was measured by a MIC ion chromatograph (Metohm, Switzerland). The dissolved organic carbon (DOC) was measured using a Shimadzu TOC-5050A total carbon analyzer. Volatile fatty acids (VFAs) were assayed using a GC-2010AF GC (Agilent Technologies, USA) equipped with a ﬂame ionization detector. Enumeration of viable SRB was performed as previously described (Wu et al., 2010) with slight modiﬁcation. Due to the strong acidity of the water samples from the extreme acid treatments, 2 M NaOH solution and sodium carbonate was used to neutralize water samples before incubation to ensure that the indicator (ferrous sulﬁde precipitate) could be easily formed. Next, a 1 mL water sample aliquot was added to each of ﬁve serum bottles. Inoculated samples were sequentially diluted and incubated under anaerobic conditions for 30 days. 2.5. DNA extraction, 16 S rDNA ampliﬁcation and denaturing gradient gel electrophoresis (DGGE) analysis At the end of the experiment, the straw samples were collected from the column used for the bioremediation of extreme acid mine drainage and the genomic DNAs of microorganisms in the straw
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were extracted using the method previously described by Wu et al. (2010). Brieﬂy, the straw sample was homogenized by vortexing them in 20 mL of phosphate-buffered saline. The supernatant was then centrifuged to remove the straw residual. Bacterial cells in the combined supernatant were harvested by centrifugation at 12,000g for 10 min and lysed by bead beating. The lysed cells were puriﬁed by adding 110 lL of sodium dodecyl sulfate (10%) and 150 lL of chloroform–isopropanol (25:1, vol/vol), followed by centrifugation at 15,000g for 10 min. The supernatant was mixed with 1/10 volume of 3 M sodium acetate and one volume of phenol, followed by centrifugation at 15,000g for 10 min. The supernatant was then extracted twice with chloroform–isopropanol (24:1, vol/vol). Nucleic acids in the supernatant were precipitated with cold ethanol and resuspended in double-distilled water. Each DNA fragment encoding 16S rRNA (corresponding to the positions 50–341 to 927–30 in the Escherichia coli sequence) was ampliﬁed using the eubacterial primer GM5F and the universal primer 907R (Muyzer et al., 1995). A 40-base GC clamp was attached to the end-50 of the GM5F primer for DGGE analysis. The primer sequences were 50 -CCTACGGGAGGCAGCAG-30 for GM5F, 50 -CCGTCAATTCCTTTRAGTTT-30 for 907R, and 50 -CGCCCGCCGCGCCCCGCGCCCGTCCCGCCGCCCCCGCCCGCCTACGGGAGGCAGCAG-30 for GC-G M5F. PCR ampliﬁcations were performed with a Mastercycler gradient PCR system (Eppendorf China Ltd.) as described by Nakagawa et al. (2002). The PCR solution consisted of 76 lL of sterile water, 10 lL of 10 Taq buffer with MgCl2, 25 pmol each of the primers, 10 lL of deoxynucleotide triphosphates mixture, and 1 lL of template DNA solution. To minimize nonspeciﬁc annealing of the primers to nontarget DNA, 2.5 U of Taq polymerase (Sangon, Shanghai, China) was added to the reaction mixture at 80 °C after an initial denaturing step of 94 °C for 5 min. The temperature was subsequently cooled to 65 °C for 1 min. This temperature was decreased by 1 °C every second cycle until a touchdown of 55 °C, the temperature at which 10 cycles were carried out. Denaturation and primer extension were carried out at 94 °C for 1 min and at 72 °C for 3 min, respectively. Cycling was completed by a ﬁnal extension at 72 °C for 10 min. PCR products were veriﬁed by electrophoresis in a 0.8% agarose gel, and then analyzed by DGGE. DGGE analysis was performed by loading PCR-ampliﬁed DNA product onto a 6% (wt/vol) acrylamide gel containing a denaturant gradient of 20–60% parallel to the electrophoresis detection using the D-code system (Bio-Rad, Hercules, CA). Electrophoresis was performed at 60 °C with a voltage of 200 V for 4 h. After electrophoresis, the gels were incubated for 10 min in ethidium bromide, rinsed in distilled water, and then analyzed with UV transillumination. For further identiﬁcation of predominant DGGE bands in individual samples, DGGE fragments were cut and eluted in 50 mL of TE buffer overnight. Recovered DNA was used as DNA template in a following PCR ampliﬁcation with the same DGGE primer set as used previously. The PCR products were analyzed in a separate DGGE for purity, puriﬁed using UNIQ-10 DNA puriﬁcation kit (Sangon, Shanghai, China), and then sequenced commercially (Sangon, Shanghai, China) with an ABI 3730 DNA analyzer (Applied Biosystems) using the GM5F primer. Multiple sequence alignment of partial sequences was carried out with ClustalW (Thompson et al., 1994) and was further reﬁned manually. Sequence similarities were checked using the NCBI Blastn program (Altschul et al., 1997). Phylogenetic trees were constructed using the maximum likelihood distance method.
3. Results and discussion 3.1. Acid tolerance response of sulfate biological reduction under different acidic environments Treatment of AMD using sulfate-reducing reactor depends on biological sulfate reduction (Hulshof et al., 2003; Huisman et al., 2006; Jin et al., 2008; Wu et al., 2010), which is the critical bioprocess for AMD bioremediation. To know the acid tolerance potential of biological sulfate reduction, acid tolerance response of this key bioprocess under different acidic environments was investigated. In the extreme acid treatment with a pH value of 2.0, efﬁcient sulfate reduction was observed (Fig. 1). The maximum sulfate bioreduction rate decreased by 42.5% from 21.4 ± 2.3 to 12.3 ± 0.8 mg L1 d1 when the initial pH changed from 7.0 to high acid treatment with an initial pH of 2.0. It decreased by 31.3% when the initial pH changed to moderate acid treatment with an initial pH of 3.0. The maximum sulfate rate of the high acid treatment was slightly lower than that of the moderate acid treatment whose maximum sulfate bioreduction rate was 14.7 ± 1.2 mg L1 d1. Even in the extreme acid treatment with initial pH of 1.0, the sulfate reduction had a maximum reduction rate of 1.2 ± 0.3 mg L1 d1, showing that the biological sulfate reduction process did not totally cease. Although the biological sulfate reduction sharply decreased with the decrease in initial pH, efﬁcient or notable sulfate reduction was still observed in the high or extreme acid treatment, indicating good acid tolerance of this bioprocess. Efﬁcient sulfate reduction was observed in previous research with fairly low pH (pH = 2.5) (Tsukamoto et al., 2004). To our knowledge, this is the ﬁrst time to show the good tolerance potential of biological sulfate reduction at pH 6 2.0. Since the pH of AMD near or below 2.0 had been frequently observed (Morin et al., 1988; Johnson et al., 2000; Jurjovec et al., 2002), the demonstrated high acid tolerance of sulfate reduction has signiﬁcant environmental implications for the practical application of the sulfate reducing bioremediation system in mining industry. Enumeration of viable SRB was performed at the beginning and end of each experiment to measure the survival of SRB in extremely acidic mine drainage. Successful survival or efﬁcient reproduction was observed in the extreme acid shock treatments. The numbers of SRB at the beginning were 2.8 106 MPN mL1, 2.7 106 MPN mL1, 2.8 106 MPN mL1, and 2.6 106 MPN mL1 for the extreme acid treatment, high acid treatment, moderate treatment, and control, respectively. SRB in the high acid treatment with a pH value of 2.0 were enumerated at 8.7 107 MPN mL1 at the end while those in the extreme acid treatment with pH value of 1.0 were enumerated at 2.0 106 MPN mL1. Efﬁcient reproduction of SRB
2.6. Data calculation The maximum bioreduction rate of sulfate was determined from the time course of sulfate disappearance, using points in the linear portion of the graphs depicting the released sulfate over time following the previously described method (Wu et al., 2010).
Fig. 1. Acid tolerance response of biological sulfate reduction in different acid treatments.
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Fig. 2. Changes in SRB activity submitted to an extreme acid environment that evolved to a high or moderate acid environment. The initial pH was 1.0. The pH was adjusted to 2.0 or 3.0 at day 7 for the short-term acid treatment test, and at day 28 for the long-term acid treatment test.
was observed in the high acid treatment with pH of 2.0 while more than 70% of SRB still survived in the extreme acid treatment with pH of 1.0 after 56 d of acid treatment. The wide occurrence of acidophilic or acid-tolerant SRB has been previously reported (Koschorreck, 2008; Alazard et al., 2010). The survival or reproduction of SRB under extremely or highly acidic conditions conﬁrmed the potential significant and symbiotic role of these functional bacteria in the bioremediation process under extremely or highly acidic conditions. 3.2. Evolution of the SRB activity submitted to an extreme acid environment that evolved to a high or moderate acid environment To know the acid tolerance of the sulfate-reducing bioremediation system better, evolution of the SRB activity submitted to extreme acid environment (pH = 1) that evolved to a high (pH = 2) or moderate (pH = 3) acid environment was investigated. Rapid increase in biological sulfate reduction after an acclimation stage was observed when the extreme acid environment was changed to high acid environment (Fig. 2). Rapid increase in biological sulfate reduction after a short acclimation stage also occurred when the extreme acid environment was changed to moderate acid environment in case of a relatively long-term extreme acid treatment was applied. However, rapid increase in biological sulfate reduction without acclimation stage was observed when the extreme acid environment was changed to moderate acid environment after a short period. The maximum sulfate reduction rate of all treatments was greater than 9.0 mg L1 d1. Biotreatment of AMD using sulfate-reducing bioreactor was dependent on the efﬁcient reduction of sulfate (Jin et al., 2008; Wu et al., 2010). The rapid increase of sulfate reduction after extreme acid treatment suggested that the remediation system could recover rapidly after the frequently extreme acid shocks during bioremediation, and adapt well to the high pH variation of water during AMD remediation. According to a previous study (Tsukamoto et al., 2004), sulfate reduction could recover after moderate acid shock during AMD treatment. 3.3. Recovery of AMD bioremediation from extreme or high acid shock in column experiment
Fig. 3. Changes in sulfate concentration, pH value (A), and metal concentrations (B) in the efﬂuent of the column bioremediation system subjected to different acid shocks. The initial pH was 3.0. To simulate the acid shock during remediation, the pH of the inﬂuent was adjusted to 1.0 from day 27. Initial concentration of both Cu and Zn were 1.5 mg L1. The pH value was adjusted to 3.0 again at day 28, followed by another pH adjustment to pH 2.0 from day 60 to day 90.
(Fig. 3A) illustrated that the bioremediation system was tolerant of acid shock. After the extreme acid shock with pH of 1.0, the sulfate removal efﬁciency recovered rapidly and the sulfate concentration in the efﬂuent decreased to a relatively low level again. Efﬁcient sulfate reduction was achieved even when the inﬂuent pH was adjusted to 2.0 and the sulfate concentration in the efﬂuent maintained at a low level. Biotreatment of AMD was dependent on the reduction of sulfate to hydrogen sulﬁde, which could form precipitates with the metals. The efﬂuent pH remained above 5.0 during most of operation period, expect for the period of extreme acid shock with inﬂuent pH of 1.0. This should provide a suitable pH condition for heavy metal precipitation (Huisman et al., 2006) during most of the operation time. The results showed that both Cu and Zn could be removed efﬁciently when the system was suffering from extreme acid shock in which the inﬂuent pH was 2.0 (Fig. 3B). Only Cu was efﬁciently removed when a much more severe acid shock (pH = 1.0) occurred. Although results showed that the bioremediation system could only selectively prevent heavy metals from releasing in the case of more extreme acid shock, the good acid tolerance of sulfate reduction and efﬁcient removal of typical metals during the extreme acid shock period has significant environmental implications in terms of bioremediation of extreme acid mine drainages. 3.4. Effect of acid treatment on easily available organic carbon release
A column study was performed to investigate the acid tolerance and recovery response of the bioremediation system to acid shock during AMD bioremediation. Changes in the sulfate concentration
Although the sulfate reducing process is the key reaction of the bioremediation system based on sulfate reduction, this process
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relies on the release of easily available carbon (such as DOC) to provide sufﬁcient carbon and energy sources (Wu et al., 2010). Concentrations of soluble carbon under extreme acid shock conditions were monitored to investigate the effect of extreme acid shock on the organic carbon release process. High acid treatment with an initial pH of 2.0 was applied since sulfate reduction with a relatively high rate was observed under these conditions. A rapid increase in the DOC concentration in water was observed within the ﬁrst week (Fig. 4). The rapid increase in easily available carbon concentration did not immediately lead to rapid sulfate reduction, indicating that the biological sulfate reduction might be more susceptible to acid treatment than the straw breakdown process. Rapid sulfate reduction was observed after 1 week when the concentration of DOC began to decrease. To understand this bioprocess better, VFAs, which were known as the easily available carbon for biological sulfate reduction (Rittmann and McCarty, 2001), were also measured during remediation. Changes in the concentrations of VFAs showed that acetic acid was the main component of VFAs in straw remediation systems under extremely acidic conditions and the results were similar with those under slightly acidic conditions (Wu et al., 2010). VFAs concentrations increased rapidly within the ﬁrst week and then began to decrease slowly. The release of easily available carbon from straw subsequently led to the increase in sulfate reduction. These results also provide evidence that biological sulfate reduction was much more susceptible to acid shock than the straw breakdown process. Concentrations of VFAs increased again from day 21 and peaked at day 35, followed by another slow decrease. The relatively high concentrations of VFAs during the entire bioremediation period suggested that the straw break down process was tolerant of the extreme acid shock, and that the supply of easily available carbon for SRB was sufﬁcient.
Fig. 4. Changes of the easily available organic carbon and sulfate concentrations under high acid environment with initial pH of 2.0.
3.5. AMD bioremediation during the high acid treatment period A column study was carried out to investigate the bioremediation performance of extreme acid mine drainage by taking advantage of the strong acid tolerance of SRB and straw degraders. Efﬁcient sulfate reduction and sufﬁcient DOC release was observed (Fig. 5), suggesting the feasibility of using the sulfate bioreactor for remediation of extreme acid mine drainage. Concentrations of sulfate in the efﬂuent ranged from148.8 to 12.3 mg L1 while that of DOC ranged from 58.6 to 31.2 mg L1 during the entire remediation period. Both Cu and Zn were maintained at low concentration in the efﬂuent. More than 98% of Cu and 85% of Zn were removed from the AMD, showing the efﬁcient bioremediation for extreme acid mine drainage. The efﬁcient AMD bioremediation during the high acid treatment period conﬁrmed the high acid tolerance of this AMD bioremediation system.
Fig. 5. Changes of sulfate, DOC and heavy metal concentrations, as well as pH in column efﬂuent under high acid conditions.
3.6. Microbial community of the remediation system used for high AMD treatment Study of composition and structure of microbial communities from bioreactor was essential to ensure the proper functioning of treatment system (Kaushik et al., 2010). For this reason, the microbial community of the remediation system used for high AMD treatment was investigated. DGGE analysis was applied to know the microbial community of the AMD remediation system since DGGE technique has been widely used for microbial community analysis of bioreactors (Kaushik et al., 2010). The six majority sequenced DGGE bands from the remediation system belonged to the genera Clostridiaceae, Eubacterium, Pseudobutyrivibrio and Clostridium (Fig. 6). These bacteria have been associated with plant residue degradation. Clostridiaceae is associated with the degradation of rice plant residue (Akasaka et al., 2003), while Eubacterium
Fig. 6. Phylogenetic tree of dominant bacteria in straw bioremediation system under high acid conditions. Acid 1–6 refer to the clones from the remediation system. The numbers in the tree refer to percentage bootstrap values based on 1000 replications. The bar represents 5% sequence divergence.
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cellulosolvens (Taguchi et al., 2008) and Pseudobutyrivibrio ruminis strain Ce1 (Van de Vossenberg and Joblin, 2003) are common cellulose degraders. Clostridium strains Clostridium sp. (AB286217) (Ishii et al., 2008), Clostridium XB 90, and Clostridium sp. clone AC036 (Burrell et al., 2004) are also common cellulose degraders. SRB were also detected. Sequence extreme acid 1 belongs to SRB since it had high similarity with Desulfotomaculum sp. CYP9, Clostridium sp. U42, and Clostridium celerecrescens. These bacteria survived in mine water in the presence of heavy metals and were known as sulfur compound reducing bacteria (Selenska-Pobell, 2002; Takahashi, 2006; Yanez et al., 2006). The results showed that each type of bacteria (SRB and cellulolytic bacteria) had a signiﬁcant and symbiotic role in the bioremediation process even under extreme acid conditions. The anaerobic cellulolytic microorganisms broke down complex organic materials into easily available carbon to provide SRB with carbon and energy sources, suggesting that the bioremediation process under extreme acid conditions was similar with reports on bioremediation of slight or moderate AMD (Jin et al., 2008; Wu et al., 2010). 4. Conclusions High acid tolerance of an AMD bioremediation system based on sulfate reduction was ﬁrstly observed. Efﬁcient sulfate reduction was maintained during high acid treatment with plentiful release of DOC as the easily available carbon source for SRB. SRB survived the extreme or high acid treatment and sulfate reduction increased rapidly when the acid treatment was ceased. Selective precipitation of Cu and Zn could occur during extreme acid treatment period. Both SRB and straw degraders have symbiotic roles in the AMD bioremediation process under high acid conditions. These ﬁndings have signiﬁcant environmental implications in terms of AMD bioremediation for mine industry. Acknowledgements This work is ﬁnancially supported by the National Foundation of Science of China (Grant No. 40801217), National Key Basic Research Program (Grant No. 2007CB815603), and the Science Foundation of Hefei University of Technology (Grant No. 2009HGCX0233). The authors would like to thank the reviewers for their valuable suggestions and comments on the manuscript. References Akasaka, H., Izawa, T., Ueki, A., Ueki, K., 2003. Phylogeny of numerically abundant culturable anaerobic bacteria associated with degradation of rice plant residue in Japanese paddy ﬁeld soil. FEMS Microbiol. Ecol. 43, 149–161. Akcil, A., Koldas, S., 2006. Acid mine drainage (AMD): causes, treatment and case studies. J. Clean. Prod. 14, 1139–1145. Alazard, D., Joseph, M., Battaglia-Brunet, F., Cayol, J.L., Ollivier, B., 2010. Desulfosporosinus acidiphilus sp. nov.: a moderately acidophilic sulfatereducing bacterium isolated from acid mining drainage sediments. Extremophiles 14, 305–312. Altschul, S.F., Madden, T.L., Schaffer, A.A., Zhang, J., Zhang, Z., Miller, W., Lipman, D.J., 1997. Gapped BLAST and PSIBLAST: a new generation of protein database search programs. Nucleic Acids Res. 25, 3389–3402. Burrell, P.C., O’Sullivan, C., Song, H., Clarke, W.P., Blackall, L.L., 2004. Identiﬁcation, detection, and spacial resolution of Clostridium populations responsible for cellulose degradation in a methanogenic landﬁll leachate bioreactor. Appl. Environ. Microbiol. 70, 2414–2419. Egiebor, N.O., Oni, B., 2007. Acid rock drainage formation and treatment: a review. Asia-Pac. J. Chem. Eng. 2, 47–62. Huisman, J.L., Schouten, G., Schultz, C., 2006. Biologically produced sulphide for puriﬁcation of process streams, efﬂuent treatment and recovery of metals in the metal and mining industry. Hydrometallurgy 83, 106–113.
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