Anaerobic degradation of Polychlorinated Biphenyls (PCBs) and Polychlorinated Biphenyls Ethers (PBDEs), and microbial community dynamics of electronic waste-contaminated soil

Science of the Total Environment 502 (2015) 426–433

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Anaerobic degradation of Polychlorinated Biphenyls (PCBs) and Polychlorinated Biphenyls Ethers (PBDEs), and microbial community dynamics of electronic waste-contaminated soil Mengke Song a,e, Chunling Luo a,⁎, Fangbai Li b, Longfei Jiang a,c, Yan Wang a, Dayi Zhang d, Gan Zhang a a

Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou 510640, China Guangdong Institute of Eco-environmental and Soil Sciences, Guangzhou 510650, China c College of Life Sciences, Nanjing Agricultural University, Nanjing 210095, China d Lancaster Environment Centre, Lancaster University, Lancaster LA1 4YQ, UK e Graduate University of Chinese Academy of Sciences, Beijing 100039, China b

H I G H L I G H T S • The biodegradation PCBs and PBDEs in e-waste contaminated soils was studied. • DIRB and arylhalorespiring bacteria were responsive to dehalogenation respiration. • Soil bacteria and Fe ion cycling play synergistic roles in dehalogenation.

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Article history: Received 9 June 2014 Received in revised form 12 September 2014 Accepted 15 September 2014 Available online xxxx Editor: Eddy Y. Zeng Keywords: E-waste PCBs PBDEs Biodegradation Microbial community

a b s t r a c t Environmental contamination caused by electronic waste (e-waste) recycling is attracting increasing attention worldwide because of the threats posed to ecosystems and human safety. In the present study, we investigated the feasibility of in situ bioremediation of e-waste-contaminated soils. We found that, in the presence of lactate as an electron donor, higher halogenated congeners were converted to lower congeners via anaerobic halorespiration using ferrous ions in contaminated soil. The 16S rRNA gene sequences of terminal restriction fragments indicated that the three dominant strains were closely related to known dissimilatory iron-reducing bacteria (DIRB) and those able to perform dehalogenation upon respiration. The functional species performed the activities of ferrous oxidation to ferric ions and further ferrous reduction for dehalogenation. The present study links iron cycling to degradation of halogenated materials in natural e-waste-contaminated soil, and highlights the synergistic roles of soil bacteria and ferrous/ferric ion cycling in the dehalogenation of polychlorinated biphenyls (PCBs) and polybrominated biphenyl ethers (PBDEs). © 2014 Published by Elsevier B.V.

1. Introduction In the past few decades, uncontrolled electronic waste (e-waste) recycling has become common in China. The primitive technologies used, including acid leaching and burning of wire in the open, have released many organic pollutants into the environment, as polychlorinated biphenyls (PCBs) and polybrominated biphenyl ethers (PBDEs). Extremely high levels of PCBs and PBDEs are found in soil, water, air, sediment, and vegetation on/around e-waste recycling sites (Wong et al., 2007). PCBs and PBDEs are persistent, bioaccumulative and toxic, disrupting endocrine systems and posing serious threats to local ecosystems and human health (Egloff et al., 2011; Frye et al., ⁎ Corresponding author. Tel.: +86 20 85290290; fax: +86 20 85290706. E-mail address: [email protected] (C. Luo).

http://dx.doi.org/10.1016/j.scitotenv.2014.09.045 0048-9697/© 2014 Published by Elsevier B.V.

2012; Hamlin and Guillette, 2011; Purser, 2001; She et al., 2013; Wang et al., 2010; Zhou et al., 2001). Bioremediation is an efficient method to remove organic pollutants from soils. Aerobic and anaerobic biodegradation of such compounds have been investigated extensively (Bedard, 2003; Chang et al., 2013; Chen et al., 2010; Cutter et al., 2001; Dabrowska and Rosinska, 2012; Deng et al., 2011; Fennell et al., 2004; Li et al., 2005; Payne et al., 2011, 2013; Rayne et al., 2003; Shih et al., 2012). Aerobic microbial degradation is often possible only when halogenated compounds have a maximum of four-to-five halogen atoms, although an aerobic bacterium responsible for deca-BDE debromination was isolated from sediment (Deng et al., 2011). Generally, the biodegradability of PCBs and PBDEs congeners decreased as the number of halogen atoms rose (Furukawa, 2000). More evidence was identified in PCB-contaminated soils where lower chlorinated PCB congeners were the prior substrates for

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indigenous microorganisms (Bedard et al., 1986), and the amount of PCBs degrading microbes decreased progressively with the increasing number of chlorine atoms (Abramowicz, 1990; Bedard et al., 1986). Anoxic dehalogenation by microorganisms plays an important role in elimination of environmental PCBs and PBDEs. Dehalogenation of these compounds has been observed in soil, sewage sludge, and estuarine and marine sediment, under different redox conditions. Higher halogenated organic compounds are usually first reductively dehalogenated to less halogenated molecules under anaerobic conditions in the presence of organic compounds. During dehalogenation process, the common electron donors include lactate, acetate and glucose (Rhee et al., 2003; Boyle et al., 1999). The metabolites are further degraded to non-toxic compounds by aerobic microorganisms (Bedard, 2003; Bedard et al., 2005; Boyle et al., 1999; Bzdusek et al., 2006a,b; Li et al., 2005, 2012; Magar et al., 2005a,b; Pakdeesusuk et al., 2005; Payne et al., 2013; Sanford et al., 1996; Yan et al., 2006). Dehalorespiration is critical for effective dehalogenation, and the effect of dissimilatory iron reduction on dehalogenation has been discussed (Li et al., 2008; Wu et al., 2010). During dehalorespiration, halogenated compounds act as electron acceptors, resulting in accumulation of lower halogenated congeners and halogen-free compounds (Fetzner, 1998; Holliger et al., 1998; Wohlfarth and Diekert, 1997). A sulfidogenic 2-bromophenol-degrading consortium enriched from estuarine sediment could use 2-bromophenol as the sole electron acceptor when lactate and acetate were present as energy sources (Rhee et al., 2003). Another anaerobic bacterium isolated from estuarine sediment used lactate as electron donor and 2,4,6-tribromophenol as electron acceptor, indicating that 2,4,6-tribromophenol was converted to phenol via dehalorespiration (Boyle et al., 1999). Under anaerobic conditions, the role of dissimilatory iron reducing bacteria (DIRB) on dehalogenation has been well recognized and understood (Li et al., 2008; Wu et al., 2010). In terms of dissimilatory iron reduction, halogenated compounds are bioreduced by DIRB via an electron-shuttling mechanism in the presence of Fe(III) oxides (Feng et al., 2013; McCormick et al., 2002; Tobler et al., 2007; Wu et al., 2010). Here, Fe(III) species serve as electron shuttles, accepting electrons created during biotic oxidation of organic matter by DIRB and donating these to target contaminants (Feng et al., 2013). Carbon tetrachloride reduction is associated with the biogenic iron species produced by Geobacter sulfurreducens (DIRB) and the electron shuttle is active under iron-reducing conditions (Maithreepala and Doong, 2009). In addition, some DIRB, including Geobacter and Shewanella, utilize halogenated compounds as direct terminal elector acceptors (Feng et al., 2013). Recently, Comamonas koreensis CY01 is reported as DIRB to reduce hydrated Fe(III) oxides and 2,4-dichlorophenoxyacetic acid, extending the diversity of iron-reducing bacteria associated with dechlorination (Wu et al., 2010). To date, though many studies have invested the dehalogenation process in the field, little is known about their anaerobic bioremediation process at e-waste sites, leaving the research gaps of the natural dehalogenation in soils. The objectives of the present study were to investigate the degradation of native PCBs and PBDEs in e-waste contaminated soils, and to explore the dynamics of the microbial community responsible for degradation process under anaerobic conditions with lactate as the exotic elector donor and energy source. The work yielded useful information on mechanisms of PCBs and PBDEs dehalogenation in anaerobic bioreactors or biopilings and further bioremediation enhancement possibility. 2. Materials and methods 2.1. Soil collection and characterization Soils were collected from the e-waste recycling town of Qingyuan, Guangdong province, South China (23.57° N, 113.0° E). Topsoil samples (from 0 to 15 cm depth) from an e-waste burning site were collected,

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placed in polythene zip-bags, and transported immediately to the laboratory. Soils were homogenized, sieved through a screen of 2 mm pore size, and stored at 4 °C prior to study. The soil properties are given in Tables S1–S3.

2.2. Experimental set-up Both sterile controls and non-sterile samples were tested in this study. All samples were prepared in triplicates, and three vials were analyzed at each time point. In detail: (1) 15 g amounts of soil (dry weight) were placed in sterile 100 mL serum bottles; (2) 50 mL of 30 mM piperazine-N-N′-bis-2-ethanesulfonic acid (PIPES) buffer (pH 7.0) with lactate (10 mM) were added to each bottle; and (3) each bottle was sealed with a rubber stopper and an aluminum seal, and incubated at 25 °C in the dark. All experiments were conducted in an anaerobic glove box under 99.999% N2 at a flow rate of 80 L min−1. All stock solutions were filtered through 0.2-μm pore-size filters and stored in dark-brown containers. Stock solutions were deoxygenated under 99.999% nitrogen for 2 h prior to use in the anaerobic box. To prepare sterile controls, soils were γ-irradiated (50 kGy) for 2 h before use. Of the three carbon sources (acetate, lactate and glucose), lactate had the best performance for PCBs/PBDEs removal and selected as the electron donor in all the other treatments (see Fig. S1).

2.3. Sampling Samples were taken on 0, 24, 40, 60, and 90 d, subjected to chemical analysis (chloride ions, bromide ions, HCl-extractable Fe(II), HClextractable Fe(III), PCBs, and PBDEs) and DNA extraction. At each time point, the soil samples were taken from the incubator and kept in the anaerobic box. Three biological replicates were conducted for both sterile control and PCBs/PBDEs treatments. After vigorous mixing, 5 mL of soil suspension was directly taken for HCl-extractable Fe(II) and Fe(III) analysis. The rest of the samples passed through 0.2 μm pore-size membranes prior to the analysis of chloride and bromide ions. The soil residues were at −20 °C for further chemical and biological analyses. 2.4. Chemical analysis 2.4.1. Chloride and bromide ions Chloride and bromide ions were measured via ion chromatography (ICS-90, DIONEX) coupled with an RFICTM Ion Pac®AG14A-7 μm Guard Column (50 mm length and 4 mm i.d), an AMMS III micromembrane suppressor, an IonPac®AS14A-7 μm Analytical Column (250 mm; 4 mm i.d), and a DS5 Detection Stabilizer conductivity detector. The eluent (8.0 mM Na2CO3 and 1.0 mM NaHCO3) was pumped at 1.0 mL min−1. Sulfuric acid (0.05 M) was used to regenerate suppression of eluent conductivity. Chromeleon software was used for data analysis.

2.4.2. HCl-extractable Fe(II) and Fe(III) species All HCl-extractable Fe(II) species were extracted into 0.4 M HCl at 25 ± 1 °C in the dark for 1.5 h, filtered through a 0.2 μm pore-size membrane, and measured using the 1,10-phenanthroline method. To measure HCl-extractable iron levels, 1 mL of solution was extracted into 0.4 M HCl at 25 ± 1 °C in the dark for 1.5 h, filtered through a 0.2 μm pore-size membrane, and placed in a centrifuge tube containing 1 mL hydroxylamine hydrochloride (10%, w/v), 4 mL sodium acetate (10%, w/v), and 4 mL 1,10-phenanthroline (0.1%, w/v); iron levels were next determined using the 1,10-phenanthroline method (Heron and Christensen, 1995). The HCl-extractable Fe(III) content was taken to be the difference between the HCl-extractable total iron and Fe(II) levels.

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2.4.3. PCBs and PBDEs Freeze-dried soil samples were homogenized, pulverized, spiked with surrogate standards (TCmX, PCB30), and extracted into dichloromethane (DCM) in a Soxhlet apparatus, for 48 h, with addition of activated copper to remove sulfur. The extract was concentrated to 0.5 mL after solvent exchange to hexane. The soil extracts were purified on a multilayer silica gel/alumina column filled with anhydrous Na2SO4; 50% (w/w) sulfuric acid silica-gel; neutral silica gel (3% w/w; deactivated); and neutral alumina (3% w/w; deactivated) (from top to bottom), via elution with 16 mL hexane/DCM (1:1, v/v). After concentration to 50 μL under a gentle stream of N2, a known amount of 13C labeled PCB-141 was added prior to analysis. A total of 17 PCB congeners (PCB-8, - 28, - 49, - 60, - 66, - 77, - 82, -101, -156, -166, -170, -179, -180, -183, -189, -198, and -209) was detected via GC–EI-MS using a 50-m capillary column (Varian, CP-Sil 8 CB, 0.25 mm i.d., 0.25 μm film thickness). The initial oven temperature was set to 150 °C for 3 min, raised to 290 °C at 4 °C min−1, and held at that temperature for 10 min. The temperatures of the MSD source and the quadrupole were 230 °C and 150 °C, respectively. Eight PBDE congeners (BDE-28, -47, -99, -100, -153, -154, -183, and - 209) were detected via GC–NCI-MS (an Agilent GC7890 column coupled with a 5975C MSD). A DB5-MS (30 m, 0.25 mm i.d., 0.25 μm film thickness) capillary column was used to analyze the first seven congeners; BDE 209 was analyzed separately on a CP-Sil 13 CB column (15 m, 0.25 mm i.d., 0.2 μm film thickness). The oven was set to 130 °C for 1 min, ramped at 12 °C min−1 to 155 °C, at 4 °C min−1 to 215 °C, and at 3 °C min−1 to 300 °C; this temperature was held for 10 min. Mixed standards (17 PCB and 8 PBDE congeners) were used to quantify PCBs and PBDEs. Instrumental performance was subjected to quality control calibration by the standards, after each set of eight samples had been analyzed. Six PCB and PBDE concentrations were used to derive calibration curves. Concentrations in samples were corrected by reference to surrogate recovery levels.

(5′-GGTTACCTTGTTACGACTT) (Operon Biotechnologies) using the following polymerase chain reaction (PCR) program: 94 °C for 10 min for initial melting; 30 amplification cycles of 94 °C for 30 s, 54 °C for 30 s; and 72° for 1.5 min; and a final extension at 72 °C for 10 min. After amplification, PCR products (150 ng) were purified using an Omega ENZA Cycle-Pure kit following the manufacturer's instructions, and digested with HaeIII for 4–5 h at 37 °C. One nanogram of each labeled PCR product was analyzed on an ABI 3730 Genetic Analyzer running the Peak Scanner software version 1.0; the ROX500 set of internal standards was used. The percentage abundance of each fragment was determined. 2.4.4.2. Sequencing. For 16S rRNA gene sequencing, DNA was amplified as above except that the forward primer was unlabeled (27 F 5′-AGAG TTTGATCMTGGCTCAG). The PCR products were purified using an Omega ENZA Cycle-Pure kit. The recovered fragments were cloned into Escherichia coli JM 109 using a TA cloning kit. E. coli clones were grown on Luria-Bertani medium solidified with 15 g agar L− 1 and 50 μg mL−1 ampicillin for 16 h at 37 °C. The plasmids of positive clones were extracted with EZNA plasmid mini-kit and the insertion was confirmed by 0.8% agarose gel electrophoresis. The right clone was directly sequenced on an ABI 3730 genetic analyzer using M13 universal primers (Luo et al., 2009). Sequence similarity searches and alignments were performed with the aid of BLAST and Molecular Evolutionary Genetics Analysis, version 5. The sequences of all TRFs were deposited in GenBank under accession numbers KF535193–KF535198. 2.5. Statistical analysis Statistical analysis was performed using the SPSS package (version 11.0). Values are the means of data from three independent replicates. Analysis of variance (ANOVA) followed by Duncan's test was performed. A p value b0.05 was considered to indicate statistical significance. The equality and normality of data were tested by Brown– Forsythe and Shapiro–Wilk test respectively, and the null hypothesis was rejected for p value less than 0.05. The content variation of individual PCBs and PBDEs on 90 d followed Eq. (1):

2.4.4. Biotic analysis 2.4.4.1. T-RFLP of 16S rRNA genes. After freeze-drying the slurry soil samples, total genomic DNA of the microcosm was extracted using a MoBio Powersoil DNA isolation kit according to the manufacturer's instruction. DNA concentrations were determined using an ND-1000 UV–vis spectrophotometer. Total DNA was subjected to terminal restriction fragment (TRF) length polymorphism analysis using standard procedures (Liu et al., 1997). DNAs were amplified with 27 F-FAM (5′-AGAGTTTG ATCMTGGCTCAG; 5′ end-labeled with carboxyfluorescein) and 1492R

ð1Þ

90d

09

83

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53

40d

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54 E1 BD

C 90 −C 90s  100% C0

where C90 represents the concentration of individual molecule in samples after 90 d treatment, C90s refers to that in sterile treatment, and C0 is the concentration of individual PCBs/PBDEs molecule on day 0.

24d

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80 60 40 20 0 BDE28

BDE47

BDE100

BDE99

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BDE153

BDE183

BDE209

Fig. 1. The change of PBDEs against time. The subfigure shows the content variation of individual compounds on 90 d in samples because of the microbial activity. The stars show significant changes with time with p b 0.05. Error bars denote the standard deviation of the mean of triplicate samples.

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Concentration (ng g-1)

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60d

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sterile90d

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Fig. 2. The change of PCBs against time. The subfigure shows the content variation of individual compounds on 90 d in samples because of the microbial activity. The stars show significant changes with time with p b 0.05. Error bars denote the standard deviation of the mean of triplicate samples.

3.1. Anaerobic dehalogenation of native PCBs and PBDEs in e-waste-contaminated soil The changes of PBDEs concentration against time were illustrated in Fig. 1. No obvious change was observed in sterile controls (p N 0.05, ANOVA test), and then only the 90 d PBDEs concentrations in sterile control are presented (see Fig. S3). The concentrations of BDE-209 and BDE-47 decreased gradually, and 39.7% and 29.5% was removed respectively by 90 d. The concentrations of BDE-183, BDE-158, BDE-154 and BDE-99 had limited increase over time. The levels of less-brominated congeners, including BDE-28 and BDE-100, were steady throughout the experiment. From previous evidence on the BDE-209 degradation attribution to the production of less-brominated congeners (He et al., 2006; Kim et al., 2012; Lee and He, 2010; Tokarz et al., 2008), the loss of deca-BDE in this study might be associated with the rise in hepta-, hexa-, and penta-chlorobiphenyls under anaerobic conditions. For example, hepta-BDEs and octa-BDEs were produced by debromination of deca-BDE, and BDE-154, BDE-99, and BDE-49 were among the debromination products (He et al., 2006). Increases in the levels of nona-, octa-, hepta- and hexa-PBDEs were found upon debromination of BDE-209 in sediment (Tokarz et al., 2008). In this study, BDE-47 was degraded, whereas no significant degradation was observed for the other measured congeners including BDE-158, BDE-154, BDE-99, BDE-28 and BDE-100. The results were similar to what was found in Nan-Kan River sediment, but different from the Er-Jen River work (Yen et al., 2009). The main products of BDE-47 degradation were identified as BDE-17 or BDE-28, dependent on the functional bacteria (Robrock et al., 2008). The debromination capacity and efficiency depended on both the number of bromide atoms and the congener structure/substitution. Bromides at the meta and para positions were prior to be removed than those at the ortho position via bacterial functions (Alder et al., 1993; He et al., 2006). It was often observed that ortho-bromine substituted congeners such as BDE-183, BDE-154, BDE153 and BDE-99 were generated from the octa-BDE mixture (Robrock et al., 2008) and BDE-209 (He et al., 2006; Tokarz et al., 2008), which was consistent with the increase of BDE-183, BDE-154, BDE-153 and BDE-99 in this study. The biodegradation process of PCBs against time was shown in Fig. 2. On day 90, the levels of highly chlorinated congeners including PCB209, PCB-189, PCB-183, and PCB-179 were reduced, whereas the concentrations of PCBs with four to seven chloride atoms had risen.

No obvious change was observed in the levels of lower chlorinated congeners including PCB-66 (p = 0.991), PCB-60 (p = 0.986), PCB-28 (p = 0.972), and PCB-8 (p = 0.893). The control PCB profiles did not change during the entire experiment, and thus only the PCB concentrations at 90 d were shown. Supported by the previous study that tetra-, penta-, hexa- and hepta-chlorobiphenyls were produced during the dechlorinated process of deca- and hepta-chlorobiphenyls (Alder et al., 1993; Dabrowska and Rosinska, 2012; Payne et al., 2011), the rise of tetra-, penta-, hexa-, and hepta-chlorobiphenyls in this study may attribute to the loss of deca- and hepta-chlorobiphenyls. It was also found that approximately 65% of the meta and para chlorines were removed in 2 months in the sediment cultures (Alder et al., 1993). Dabrowska and Rosinska (2012) found that the levels of higher chlorinated congeners fell, and those of lower congeners rose, under anaerobic conditions. In a study on reductive dechlorination of commercial PCBs upon bioaugmentation with a dehalorespiring bacterium, higher chlorinated PCB levels in mesocosms decreased by about 56% (by mass) after 120 d of incubation (Payne et al., 2011). Both PCBs and PBDEs exhibited the same trend that higher halogenated compounds were reduced in amount; the levels of lower halogenated compounds were stable; and compounds with four to seven chlorides or bromides accumulated over 90 d, with the exception of BDE-47. The results might be explained by the fact that the dehalogenation often occurs among the compounds with high number ferrous ion

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3. Results and discussion

sterile ferrous ion sterile ferric ion

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Time (day) Fig. 3. The concentrations of Fe2+ and Fe3+ at different time point. Error bars denote the standard deviation of the mean of triplicate samples.

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Abandance (%)

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Granulicella sapmiensis strain S6CTX5A(85%); Geothrix fermentans strain H5(84%); Geobacter grbiciae(83%)

Anaeromyxobacter dehalogenans strain 2CP-1(86%); Anaeromyxobacter dehalogenans 2CP-C (86%); Anaeromyxobacter sp. Fw109-5 (86%)

Geothrix fermentans strain H5(100%); Desulfotalea arctica strain LSv514 (80%); Geoalkalibacter ferrihydriticus strain Z0531(80%)

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Acidiferrobacter thiooxydans strain m1(93%); Methylococcus capsulatus strain Texas(89%); Methylocaldum gracile strain VKM 14L(87%)

Denitratisoma oestradiolicum strain AcBE2-1(91%); Thiobacter subterraneus strain C55(91%); Nitrosospira multiformis strain Nl13(90%)

220bp Xylophilus ampelinus strain BPIC 48(100%); Aquincola tertiaricarbonis strain L10(96%); Rubrivivax gelatinosus strain ATH 2.2.1(96%)

Fig. 4. The abundance of main T-RFLP fragments in different time. The names below the x-axis were strains with the strongest BLAST hits with the responsive bacteria measured in this work, the number in the bracket were the similarity.

of substituted halogens (Chang et al., 2013; Payne et al., 2013; Shih et al., 2012). Besides, the dehalogenation pathways are dependent on the functions of various reductive dehalogenase-homologous in different bacterial species (Futagami et al., 2013). It has been suggested that dehalogenation may be microbial population-specific in terms of the substitution positions attacked and the numbers of halogen atoms removed, and that different microbial populations may generate different dehalogenation patterns (Alder et al., 1993; Yen et al., 2009).

significantly removed from parent compounds. Further chloride ion analysis demonstrated that chloride concentration increased from 93.84 ± 1.30 mg L− 1 on 0 d to 104.40 ± 1.11 mg L−1 on 24 d and 112.53 ± 1.57 mg L− 1 on 60 d, confirming the dehalogenation of PCBs. Comparing to the bromine background in the soil, the debromination contributed only trace bromine ions, consequently causing the stable bromine level.

3.2. Generation of chloride and bromide ions

3.3. Changes in ferrous and ferric ion levels

Chloride ions increased significantly during the dechlorination process (p b 0.05, ANOVA test), whereas bromide ions kept stable throughout the experiment (Fig. S4). This behavior was similar to the reductive transformation of pentachlorophenol (Li et al., 2008; Wu et al., 2010). In sterile controls, both chloride and bromine ions did not change (p N 0.05, ANOVA test). The results suggested that chloride ions were

As shown in Fig. 3, ferrous iron showed a cycling behavior, reaching the peak concentration at 40 d and 90 d, whereas the trough was at 60 d. Notably, the changes in ferric iron levels were the opposite as declining at 40 d, rising at 60 d, and declining again at 90 d. In the sterile controls, changes in both ferric and ferrous iron levels were similar to those in test samples at 20 d and 40 d except that the ferrous iron level decreased

220bp (KF535193)

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Aquincola tertiaricarbonis L10(NR043913.1)

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Xylophilus ampelinus BPIC 48(NR036931.1) Denitratisoma oestradiolicum AcBE2-1(NR043249.1)

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Methylococcus capsulatus Texas (NR042183.1) Geobacter grbiciae TACP-5(NR041826.1)

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Desulfotalea arctica LSv514 (NR024949.1) 182bp (KF535198)

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Anaeromyxobacter dehalogenans 2CP-1(NR027547.1)

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Anaeromyxobacter sp. Fw109-5(NR074968.1) 191bp (KF535195) 75bp (KF535197) Geothrix fermentans H5 (NR036779.1)

99

0.02 Fig. 5. Phylogenetic relationships among the unclassified functional bacteria and closely related type strains. GenBank accession numbers are in parentheses. The tree was constructed using the neighbor-joining algorithm. Bar, 2% estimated sequence divergence.

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from 21 d to 40 d; neither level changed at the last sample times after 40 d. Soil originally exposed to air was transferred to reductive anaerobic conditions, facilitating redox reactions. Hence, the ferric iron level initially decreased and that of ferrous iron rose in both sterile and nonsterile samples. In sterile control without biological activity, the slight decrease of ferrous iron from 0 d to 40 d might attribute to the changing redox equilibrium from aerobic (0 d before experiment) to anaerobic condition (40 d) (Yang, 2001; Gotoh and Patrick, 1974; Brennan and Lindsay, 1998). In test samples, changes in iron levels were also affected by biological factors. Ferrous iron acts as an electron donor in dehalogenation and some microorganisms can oxidize ferrous iron, generating ferric iron (McCormick et al., 2002). Also, dehalogenation can also be driven by dissimilatory iron reduction (Feng et al., 2013; McCormick et al., 2002; Wu et al., 2010). On the other hand, DIRB can transfer electrons to ferric iron when oxidizing organic material. Therefore, both ferrous and ferric iron are generated and consumed continuously in test samples if energy is available. 3.4. Dynamics of microbial composition and function Terminal restriction fragment length polymorphism (T-RFLP) analysis (Fig. 4) showed that the relative abundances of the principal TRFs changed over time. Six prominent fragments were observed. The 16S rRNA sequencing suggested that the two organisms yielding TRFs of 75- and 191-bp belonged to the family Geobacteraceae, and the organism yielding the 191-bp TRF was closely related to Geothrix fermentans H5 (similarity, 100%) (Fig. 5). However, they had different relative abundance behavior that the 191-bp TRF was dominant at 0 d and rapidly declined at 24 d, whereas the relative abundance of the 75-bp TRF increased at 24 d and 40 d, subsequently declining at 90 d. The Geobacteraceae are strictly anaerobic dissimilatory Fe (III)-reducing bacteria which, in the presence of organic matter, form ferrous iron species abiotically reducing certain chlorinated hydrocarbons (Feng et al., 2013; McCormick et al., 2002; Wu et al., 2010). DIRB bacteria such as Geobacter and Shewanella directly use chlorinated hydrocarbons as terminal electron acceptors, and both Fe(III) oxides and bacterial action promote contaminant bioreduction via electron shuttling (Lonergan et al., 1996; Tobler et al., 2007; Wu et al., 2010). The 16S rRNA gene sequence of the organisms yielding the 182-bp TRF was closely related to that of the myxobacteria, which are unique arylhalorespiring facultative anaerobic organisms. Its relative abundance of increased throughout the whole experiment. Such chlororespiring myxobacteria can use 2,6-dichlorophenol, 2,5-dichlorophenol and 2-bromophenol as terminal electron acceptors, and lactate as an electron donor (Sanford et al., 2002). The relative abundance of the 201-bp TRF has the cycling behavior, and its peak and trough was at 24 d (90 d) and 40 d, respectively. The 16S rRNA sequence indicated its high similarity to Denitratisoma oestradiolicum AcBE2-1 (denitrifying bacteria) which can grow on fatty acids (C2 to C6) with nitrate as the electron acceptor (Fahrbach et al., 2006). The relative abundance of 198- and 220-bp TRF was similar, dominant at 0 d and dramatically declining at 24 d. The 198-bp TRF was closely related to methanotrophic bacteria, some of which grow under fully aerobic conditions whereas others are facultative aerobes. The organism yielding the 220-bp TRF matched with the aerobic Xylophilus ampelinus BPIC 48 (similarity, 100%) and Aquincola tertiaricarbonis L10, neither of which perform dehalogenation. The microcosm environment greatly influences microbial community composition. When soil lacks oxygen, aerobic (and some facultative) bacteria grow poorly, explaining the sharp falls in the relative abundances of the 220- and 198-bp TRFs in the early experiment. Anaerobic bacterial numbers rose over time, eventually becoming predominant. Thus, the relative abundance of organisms yielding 75- and 182-bp TRFs increased to 90 d during all the sample time points. The

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organisms represented by the 191-bp TRF were acetate-oxidizing Fe(III)-reducing bacteria, and variation in the level of this TRF was probably attributable to changes in the growth environment and acetate level (Sanford et al., 2002). Previous studies showed that aerobic bacteria play an important role in the biodegradation of halogenated compounds with less than five halogen atoms (Chang et al., 2013; Chen et al., 2010). Aerobic bacterium with the ability of debrominating deca-BDE was also isolated from sediment by Deng (Deng et al., 2011); while for the microbial degradation of higher halogenated compounds, it was proved by different researches that the compounds were dehalogenated under anaerobic conditions by various anaerobic bacteria (Bedard, 2003; Bedard et al., 2005; He et al., 2006; Tokarz et al., 2008; Yen et al., 2009). The dehalogenation of PCBs and PBDEs observed in this work was attributed to the biochemical process, which probably involved the actions of dissimilatory Fe(III)reducing and arylhalorespiring bacteria. The ferrous iron concentration increased as the amount of dissimilatory Fe(III)-reducing bacteria rose in test samples. A clear peak in ferrous ion concentration was evident at 40 d. Furthermore, arylhalorespiring bacteria represented by the 182-bp TRF grew rapidly. These growth patterns may have caused decreases in the levels of highly chlorinated and brominated congeners including PCB-209, PCB-198, and BDE-209, from 40 d to 90 d. Work with pure cultures showed that myxobacteria can exert a dehalogenating activity under anaerobic conditions (Sanford et al., 2002). The suggestion that dehalogenation involved the action of ferrous ions is supported by the decrease in ferrous ion concentration, with a concomitant increase in ferric ion concentration, from 40 d to 60 d. Ferric ions may serve as electron shuttles accepting electrons from biotic oxidations of organic matters by DIRB, donating such electrons to halogenated compounds. This is consistent with earlier data (Wu et al., 2010), which reported that the maximum decline in 2,4-dichlorophenoxyacetic acid concentration coincided with the fall in Fe(II) concentration. Next, Fe(III)-reducing bacteria (with 75- and 191-bp TRFs) reduce ferric ions, decreasing ferric ion and increasing ferrous ion concentrations between 60 d and 90 d. 3.5. Possible solutions for bioremediation implication Our results have revealed that PCBs and PBDEs could be biodegraded under anaerobic conditions with lactate as electron donor. Dissimilatory Fe(III)-reducing and arylhalorespiring bacteria are important in this context. Hence, measures could be taken to increase the activities of such species in the bioreactor or biopiling. For example, the restricted carbon source (such as lactate) could be dosed in the contaminated soils to stimulate reductive dechlorination by DIRB, and so could iron oxide. Also, humic substances serve as electron shuttles in DIRB and arylhalorespiring bacteria. These microorganisms transfer electrons to dissolved humic substances, and the metabolites can rapidly reduce Fe(III) oxides or halogenated compounds (Cervantes et al., 2002; Lovley et al., 1996; Nevin and Lovley, 2000). Thus, our present study extends the state of knowledge on dehalogenation processes under anoxic conditions in bioreactor or biopiling system, and offers possible solutions for bioremediation implications of e-waste-contaminated soils. Further efforts, to adjust the growth environments of functional bacteria and gain insight into the complex chemical–biological mechanisms of dehalogenation in e-waste-contaminated soils, will aid in extrapolation of our laboratory results to field conditions. Acknowledgments This study was supported by the Joint Funds of the National Natural Science Foundation of China and the Natural Science Foundation of Guangdong Province, China (no. U1133004), and the National Natural Science Foundation of China (nos. 41173082 & 41322008).

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Appendix A. Supplementary data Detailed information can be found in the supporting document on soil properties, initial PCBs/PBDEs contamination, effects of different electron donor, and PCBs/PBDEs concentration change in sterile control. Supplementary data are available online at http://dx.doi.org/10.1016/j. scitotenv.2014.09.045.

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