Anaerobic degradation of tetrabromobisphenol-A in river sediment

Ecological Engineering 49 (2012) 73–76

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Anaerobic degradation of tetrabromobisphenol-A in river sediment Bea-Ven Chang ∗ , Shaw-Ying Yuan, Yen-Lin Ren Department of Microbiology, Soochow University, Taipei, Taiwan

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Article history: Received 3 February 2012 Received in revised form 12 July 2012 Accepted 10 August 2012 Available online 28 September 2012 Keywords: Anaerobic degradation Tetrabromobisphenol-A River sediment

a b s t r a c t The contamination of the environment with tetrabromobisphenol-A (TBBPA), an endocrine disruptor, is a concern. We examined anaerobic degradation of TBBPA in sediment samples from the Erren River in southern Taiwan. Anaerobic degradation of TBBPA was enhanced with the addition of humic acid (0.5 g L−1 ), sodium chloride (1 mass/vol%), zero-valent iron (1 g L−1 ), vitamin B12 (0.025 mg L−1 ), brij 30 (55 ␮M), brij 35 (91 ␮M), rhamnolipid (130 mg L−1 ), or surfactin (43 mg L−1 ) but was inhibited by the addition of acetate (30 mM), lactate (20 mM), or pyruvate (20 mM). Sulfate-reducing bacteria, methanogen, and eubacteria are involved in the anaerobic degradation of TBBPA; sulfate-reducing bacteria is a major component of the sediment. © 2012 Elsevier B.V. All rights reserved.

1. Introduction Tetrabromobisphenol-A (TBBPA) is a flame retardant used in the production of many plastic polymers and electronic circuit boards (de Wit, 2002). It has been found in environmental samples and in human plasma and may have a toxic effect (Darnerud, 2003). Microbial degradation is the primary mechanism for removal of organic toxic compounds in sediment (Yu et al., 2012). Reductive dehalogenation (e.g., substitution of Br or Cl by a hydrogen atom) is an important mechanism (Fetzner, 1998; Davis et al., 2005). Halogenated compounds are electron acceptors in respiratory or cometabolic processes. Environmental factors such as temperature, pH, salinity, plant species selection and the availability of organic carbon and/or inhibiting substances influence the growth and activity of microbes, and the manipulation of some has been investigated (Faulwetter et al., 2009). The addition of sodium chloride, humic acid, zero-valent iron, surfactants, electron donors or electron acceptors influences the anaerobic degradation of organic toxic chemicals in sediment (Chang et al., 2009). However, little is known about the effects of factors on the anaerobic degradation of TBBPA in river sediment. The climatic characteristics of subtropical regions foster diverse microbial communities (Chang et al., 2009). Several techniques used to study microbial communities in environmental samples include phospholipid fatty acid analysis (e.g., Langer and Rinklebe, 2009, 2011) and molecular-biological methods (Faulwetter et al., 2009). Many studies have used PCR-denaturing gradient gel

∗ Corresponding author. Tel.: +886 228806628; fax: +886 228831193. E-mail address: [email protected] (B.-V. Chang). 0925-8574/$ – see front matter © 2012 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2012.08.038

electrophoresis (PCR-DGGE) to examine the effect of pollutants on microbial communities in sediment (Castle et al., 2006; Chang et al., 2009). Little information is available on the effect of TBBPA anaerobic degradation on the change in bacterial communities in river sediment. We aimed to examine the effect of factors on the anaerobic degradation of TBBPA in the sediment of Erren River, one of the most heavily contaminated rivers in southern Taiwan, and changes in the microbial community in the sediment. 2. Materials and methods 2.1. Chemicals TBBPA (98.0%) analytical standard was from Aldrich Chemical Co. (Milwaukee, WI). Solvents were from Mallinckrodt, Inc. (Paris, KY). The biosurfactants used in this study were surfactin and rhamnolipid as described by Yeh et al. (2005) and Wei et al. (2005), respectively. All other chemicals were from Sigma Chemical Co. (St. Louis, MO). The log Kow for TBBPA, tribromobisphenol-A, dibromobisphenol-A and monobromobisphenol-A was 4.5, 2.1, 2.1, and 3.7, respectively. 2.2. Sampling and medium We collected sediment samples from Erren River in July 2008. The three sampling sites, A (22.55◦ 10.98 N, 120.11◦ 3.51 E), B (22.55◦ 14.32 N, 120.11◦ 12.9 E) and C (22.54◦ 51.13 N, 120.13◦ 27.01 E) are well known from previous studies of aquatic pollutants (Yuan et al., 2011). The sediments (>15 cm) were collected by use of a soil core during low tide. Adaptation involved adding 50 ␮g g−1 TBBPA to 500 g sediment at 14-d intervals under

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static incubation at 30 ◦ C without light for 6 months. Here, sediment refers to TBBPA-adapted sediment. For sites A, B, and C, the TBBPA concentration in sediment was 260, 450, and 38.1 ng g−1 , respectively. In our previous study of the degradation of TCBPA in sediment from the 3 sampling sites, the anaerobic degradation rate of TCBPA was higher in site B than other sites (Yuan et al., 2011). Therefore, we used the sediment sample from site B in the following experiments. The experimental medium consisted of (in g L−1 ) NH4 Cl, 2.7; MgCl2 ·6H2 O, 0.1; CaCl2 ·2H2 O, 0.1; FeCl2 ·4H2 O, 0.02; K2 HPO4 , 0.27; KH2 PO4 , 0.35; yeast extract, 0.2; and resazurin, 0.001. The pH was adjusted to 7.0 after autoclaving; 0.9 mM titanium citrate was added as a reducing reagent. 2.3. Experimental design All experiments involved use of 125-mL serum bottles containing 45 mL medium, 5 g river sediment and 50 ␮g g−1 TBBPA. We measured the effect of the following factors on anaerobic degradation in sediment collected from site B: sodium chloride (1 mass/vol%); humic acid (0.5 g L−1 ); zero-valent iron (1 g L−1 ); vitamin B12 (0.025 mg L−1 ); surfactants, brij 30, brij 35, rhamnolipid, and surfactin at 1 critical micelle concentration (CMC; the CMC values were 55 ␮M, 91 ␮M, 130 mg L−1 , and 43 mg L−1 , respectively); the electron donor sodium acetate (30 mM), sodium lactate (20 mM) or sodium pyruvate (20 mM); sodium hydrogen carbonate (20 mM), sodium sulfate (20 mM), or sodium nitrate (20 mM) for methanogenic, sulfate-reducing or nitrate-reducing conditions, respectively; and microbial inhibitors (50 mM BESA, 50 mM vancomycin, 50 mM sodium molybdate-2-hydrate). The concentrations of these factors were from previous studies (Yuan et al., 2011). Inoculated control samples (without sodium hydrogen carbonate, sodium sulfate, or sodium nitrate), considered nonsterile sediment, were shaken before incubation at 30 ◦ C and pH 7.0 in the dark. Sterile controls were autoclaved at 121 ◦ C for 30 min on 3 d. All experiments were conducted in an anaerobic glove box (Forma Scientific, USA) filled with N2 (85%), H2 (10%), and CO2 (5%). The 125-mL serum bottles were capped with butyl rubber stoppers, wrapped in aluminum foil to prevent photolysis, and incubated without shaking at 30 ◦ C in the dark. Each treatment was applied in triplicate. Samples were collected every 7 d to measure residual TBBPA, then underwent PCR-DGGE. Methane was sampled from the headspace of the serum bottles. 2.4. Analytical methods TBBPA was extracted twice from sediment samples by use of dichloromethane, then again 20 min at 30 ◦ C with use of a Branson 5200 ultrasonic cleaner. Extracts were analyzed by use of gas chromatography (Hewlett Packard 6890) equipped with an electron capture detector and HP-5 capillary column. The initial column temperature was set at 250 ◦ C, increased by 2 ◦ C min−1 to 260 ◦ C, and then increased by 10 ◦ C min−1 to 280 ◦ C. Injector and detector temperatures were set at 300 and 320 ◦ C, respectively. The recovery percentage and detection limit for TBBPA was 91.5% and 0.02 mg L−1 , respectively. The anaerobic degradation products of TBBPA and methane levels were analyzed as we described previously (Chang et al., 2011). 2.5. DNA extraction and PCR-DGGE analysis DNA was extracted from sediment samples using the Mo Bio PowerSoil DNA kit (Carlsbad, CA). The primer sequences for DGGE analysis were for FGC968 (Escherichia coli position 968–983),

Table 1 Effect of various substrates on TBBPA anaerobic degradation rate constant (k1 ) and half-life (t1/2 ) in sediment of Erren River, southern Taiwan. Treatment b

Inoculated control Humic acid (0.5 g L−1 ) Sodium chloride (1%) Zero-valent iron (1 g L−1 ) Vitamin B12 (0.025 mg L−1 ) Rhamnolipid (130 mg L−1 ) Surfactin (43 mg L−1 ) Brij 30 (55 ␮M) Brij 35 (91 ␮M) Sodium acetate (30 mM) Sodium lactate (20 mM) Sodium pyruvate (20 mM) Sodium hydrogen carbonate (20 mM) Sodium sulfate (20 mM) Sodium nitrate (20 mM)

k 1 (d−1 )

t1/2 (d)

ra

0.0417 0.0491 0.0502 0.0541 0.0529 0.0686 0.0582 0.0510 0.0554 0.0297 0.0276 0.0340 0.0582 0.0679 0.0525

16.6 14.1 13.8 12.8 13.1 10.1 11.9 13.6 12.5 23.3 25.1 20.4 11.9 10.2 13.2

0.98 0.94 0.95 0.92 0.89 0.94 0.92 0.94 0.95 0.95 0.96 0.97 0.89 0.95 0.96

Each treatment was significantly different from the inoculated control at p < 0.05. a r = correlation coefficient. b Inoculated control: 30 ◦ C, pH 7.0, TBBPA 50 ␮g g−1 .

5 -GCCCGGGGCGCGCCCGGGCGGGGCGGGGGCACGGGGG GAACGCGAAGAACCTTAC-3 , and R1401 (E. coli position 1401–1385), 5 -CGGTG TGTACAAGACCC-3 (Chang et al., 2009). DGGE involved use of a D-gene and D-code system (Bio-Rad Laboratories, CA, USA). Electrophoresis involved 1× TAE buffer at voltage 60 V and temperature 60 ◦ C for 16 h. The bands of interest were excised and soaked in elution buffer overnight at 37 ◦ C. The DNA was re-amplified with the primers for FGC968 and R1401. The re-amplified products were again purified and sequenced with use of an ABI-Prism automatic sequencer. 2.6. Data analysis The TBBPA biodegradation data collected for this study fit well with first-order kinetic equations: S = S0 exp (−k1 t), t1/2 = ln 2/k1 , where t is time, S0 is the initial substrate concentration, S is the substrate concentration at time t, and k1 is the degradation rate constant. Principal component analysis (PCA) was used to examine the DGGE community structure. Statistical analysis involved use of SPSS v10.0 (SPSS Inc., Chicago, IL, USA). 3. Results and discussion 3.1. Effects of various factors on the anaerobic degradation of TBBPA in the sediment The TBBPA concentrations in the sterile controls were first examined at the end of the 35-d incubation. The proportion of TBBPA ranged from 92.1% to 95.3%. Therefore, the TBBPA degradation in the following experiments was due to microbial action. The degradation rate and half-life of TBBPA were 0.0417 d−1 and 16.6 d, respectively (inoculated control) (Table 1). As compared with the inoculated control, the addition of humic acid, sodium chloride, zero-valent iron, and vitamin B12 enhanced the degradation rate of TBBPA by 17.7%, 20.4%, 29.7%, and 26.9%, respectively. Humic acid showed increased showed a higher reducing capacity in deeper layers, probably because of reduction by humic-acidreducing microorganisms (Kappler et al., 2004). TBBPA degradation was enhanced by the addition of sodium chloride (1%). The types of bacteria colonized in sediment and their biodegradation potential are affected by salinity (Tam et al., 2002). The high salinity of sample sediment may significantly inhibit the degradation rate (Yu et al., 2012). The addition of zero-valent iron enhanced the degradation of TBBPA. Zero-valent iron can be an electron donor and can

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Fig. 1. Proposed anaerobic biotransformation pathways of TBBPA in the sediment.

efficiently remove PAH (Chang et al., 2007). The TCE dechlorination and cellular growth rates doubled when vitamin B12 concentration was increased 25-fold to 0.025 mg L−1 (He et al., 2007). The non-ionic surfactants brij 30 and brij 35 enhanced TBBPA degradation rates by 32.9% and 22.3%, respectively, whereas rhamnolipid and surfactin elevated degradation rates by 64.5% and 39.6%, respectively. The use of surfactants can increase the degradation of hydrophobic organic compounds in contaminated environments by increasing the total aqueous solubility of these compounds (Yeh et al., 2005; Wei et al., 2005). The addition of rhamnolipid yielded a higher TBBPA degradation rate than did the other additives. The rhamnolipid used in this study, a commonly isolated glycolipid biosurfactant, was produced by Pseudomonas aeruginosa J4, whereas while the surfactin, a lipoprotein-type biosurfactant, was produced by Bacillus subtilis ATCC 21332. This finding is consistent with that of Yuan et al. (2011), who reported on the anaerobic degradation of TCBPA in sediment. We also noted an inhibition of TBBPA degradation with the addition of acetate, lactate, or pyruvate. Acetate, lactate, and pyruvate may not function as electron donors under our conditions. This result is similar to with our previous studies of anaerobic degradation of nonylphenol in sediment (Chang et al., 2009). 3.2. Biotransformation of TBBPA in the sediment We monitored the intermediate products of the degradation of TBBPA in sediment at 0, 20, and 35 d. We observed 2 intermediate products, dibromobisphenol-A and BPA, at d 20 and d 35, respectively. The concentration of TBBPA decreased from 0, 20, and 35 d, and the 2 products of TBBPA degradation in sediment were obtained. Dehalogenation of TBBPA is probably a stepwise process of sequential removal of bromine atoms. From this result, we propose the following biochemical pathway: TBBPA → → dibromobisphenol-A → → BPA (Fig. 1). Ronen and Abeliovich (2000) observed the reductive degradation of TBBPA to BPA in anaerobic sediment from a wet ephemeral desert stream bed. Similar results were found with transformation of TBBPA in sediment (Voordeckers et al., 2002; Davis et al., 2005). BPA did not anaerobically degrade after 140 d in the sediment (Chang et al., 2011). BPA has an OH substituent on both aromatic rings, which is a possible site for degradation, but the 2 rings are joined by a quaternary carbon, which may inhibit its degradation by anaerobic microbes. 3.3. Comparison among various electron acceptors and microbial inhibitors of the anaerobic degradation of TBBPA Adding electron acceptors sodium hydrogen carbonate, sodium sulfate, or sodium nitrate increased the degradation rate of TBBPA by 39.6%, 62.8%, and 25.9%, respectively (Table 1). Compared with the inoculated control, methanogenic, sulfate-reducing and nitrate-reducing conditions enhanced TBBPA degradation, the degradation rate being sulfate-reducing > methanogenic > nitratereducing conditions. Adding molybdate (a selective inhibitor of sulfate-reducing bacteria), BESA (a selective inhibitor of methanogens), or vancomycin (a selective inhibitor of eubacteria)

(Arbeli et al., 2006) decreased the degradation rate of TBBPA by 76.3%, 64.7%, and 57.1%, respectively (Table 2). The degradation rates for TBBPA decreased greatly when the activities of sulfatereducing bacteria were inhibited. Sulfate-reducing bacteria belong to eubacteria (domain Bacteria), whereas the methanogens are classified within the domain Archaea. The methanogens can only utilize simple substrates such as H2 /CO2 and acetate, but the sulfate-reducing bacteria can use a wide spectrum of organic compounds for growth in the presence or absence of sulfate (Yuan et al., 2011). In addition, methane production was not detected with TBBPA degradation under the three reducing conditions after 35 d of incubation. Our results indicate that sulfate-reducing bacteria constitute a major microbial component in TBBPA degradation, with methanogen and eubacteria microbial populations also involved. 3.4. Microbial community analysis The microbial community changes in DGGE band profiles and PCA with various treatments after 35 d of incubation are shown in Fig. 2. The DGGE profile consisted of at least 6 bands, and the number of bands was changed with various treatments. The first principal component (PC1 = 47.8%) and second principal component (PC2 = 41.6%) explained 89.4% of the variation and discriminated between samples with various TBBPA concentrations. Microbial communities differed significantly in non-TBBPA-adapted and TBBPA-adapted sediment. Microbial communities changed significantly when various substrates were added to the sediment. Highly similar microbial communities were also found in samples with sodium chloride, humic acid, zero-valent iron, brij 30, and brij 35, which can enhance TBBPA degradation. This observation is consistent with our previous finding that treatment with different substrates affects microbial communities in the sediment (Chang et al., 2009). In summary, anaerobic degradation of TBBPA is a major process that results in decontamination of river sediments. The addition of humic acid, sodium chloride, zero-valent iron, vitamin B12 , brij 30, brij 35, rhamnolipid, surfactin, acetate, lactate, or pyruvate can influence the degradation of TBBPA. The optimal condition involved the addition of rhamnolipid into the sediment. DGGE analyses revealed sediment with a specific bacterial community. Treatment with different substrates changed the microbial communities in sediment samples. The results support the feasibility of removing TBBPA in sediment by anaerobic degradation. We will use Table 2 TBBPA anaerobic degradation rate constant (k1 ) and half-life (t1/2 ) in sediment with the addition of 3 microbial inhibitors. Treatment b

Inoculated control Vancomycin (50 mM) BESA (50 mM) Molybdate (50 mM)

k1 (d−1 )

t1/2 (d)

ra

0.0417 0.0318 0.0270 0.0238

16.6 21.8 25.7 29.1

0.98 0.95 0.96 0.97

Each treatment was significantly different from the inoculated control at p < 0.05. a r = correlation coefficient. b Inoculated control: 30 ◦ C, pH 7.0, TBBPA 50 ␮g g−1 .

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Fig. 2. Changes in denaturing gradient gel electrophoresis band profiles (A) and principal component analysis (PCA) (B) after 35 d of incubation with various additives. Symbols: 1–10 represent non-TBBPA-adapted sediment, TBBPA-adapted sediment, sodium chloride, humic acid, vitamin B12 , zero-valent iron, brij 30, brij 35, surfactin, and rhamnolipid, respectively.

similar conditions to further define the operating parameters for river sediment bioremediation and to monitor the microbial dynamics involved in the degradation of TBBPA in river sediment. Acknowledgment This research was supported by the National Science Council, Republic of China, Grant Number 98-2627-B-031-001. References Arbeli, Z., Ronen, Z., Diaz-Baez, M.C., 2006. Reductive dehalogenation of tetrabromobisphenol-A by sediment from a contaminated ephemeral streambed and an enrichment culture. Chemosphere 64, 1472–1478. Castle, D.M., Montgomery, M.T., Kirchman, D.L., 2006. Effects of naphthalene on microbial community composition in the Delaware estuary. FEMS Microbiol. Ecol. 56, 55–63. Chang, M.C., Shu, H.Y., Hsieh, W.P., Wang, M.C., 2007. Remediation of soil contaminated with pyrene using ground nanoscale zero-valent iron. J. Air Waste Manage. Assoc. 57, 221–227. Chang, B.V., Lu, Z.J., Yuan, S.Y., 2009. Anaerobic degradation of nonylphenol in subtropical mangrove sediments. J. Hazard. Mater. 165, 162–167. Chang, B.V., Yuan, S.Y., Chiou, C.C., 2011. Biodegradation of bisphenol-A in river sediment. J. Environ. Sci. Health A 46, 931–937. Darnerud, P.O., 2003. Toxic effects of brominated flame retardants in man and in wildlife. Environ. Inter. 29, 841–853. Davis, J.W., Gonsior, S., Marty, G., Ariano, J., 2005. The transformation of hexabromocyclododecane in aerobic and anaerobic soils and aquatic sediments. Water Res. 39, 1075–1084.

de Wit, C.A., 2002. An overview of brominated flame retardants in the environment. Chemosphere 46, 583–624. Faulwetter, J.L., Gagnon, V., Sundberg, C., Chazarenc, F., Burr, M.D., Brisson, J., Camper, A.K., Stein, O.R., 2009. Microbial processes influencing performance of treatment wetlands: a review. Ecol. Eng. 35, 987–1004. Fetzner, S., 1998. Bacterial dehalogenation. Appl. Microbiol. Biotechnol. 50, 633–657. He, J., Holmes, V.F., Lee, P.K.H., Alvarez-Cohen, L., 2007. Influence of vitamin B12 and cocultures on the Growth of Dehalococcoides isolates in defined medium. Appl. Environ. Microbiol. 73, 2847–2853. Kappler, A., Benz, M., Schink, B., Brune, A., 2004. Electron shuttling via humic acids in microbial iron (III) reduction in a freshwater sediment. FEMS Microbiol. Ecol. 47, 85–92. Langer, U., Rinklebe, J., 2009. Lipid biomarkers for assessment of microbial communities in floodplain soils of the Elbe River (Germany). Wetlands 29, 353–362. Langer, U., Rinklebe, J., 2011. Priming effect after glucose amendment in two different soils evaluated by SIR- and PLFA-technique. Ecol. Eng. 37, 465–473. Ronen, Z., Abeliovich, A., 2000. Anaerobic–aerobic process for microbial degradation of tetrabromobisphenol A. Appl. Environ. Microbiol. 66, 2372–2377. Tam, N.F.Y., Guo, C.L., Yau, W.Y., Wong, Y.S., 2002. Preliminary study on biodegradation of phenanthrene by bacteria isolated from mangrove sediments in Hong Kong. Mar. Pollut. Bull. 45, 316–324. Voordeckers, J.W., Fennell, D.E., Jones, K., Haggblom, M.M., 2002. Anaerobic biotransformation of tetrabromobisphenol-A tetrachlorobisphenol-A, and bisphenol-A in estuarine sediments. Environ. Sci. Technol. 36, 696–701. Wei, Y.H., Chou, C.L., Chang, J.S., 2005. Rhamnolipid production by indigenous Pseudomonas aeruginosa J4 originating from petrochemical wastewater. Biochem. Eng. J. 27, 146–154. Yeh, M.S., Wei, Y.H., Chang, J.S., 2005. Enhanced production of surfactin from Bacillus subtilis by addition of solid carriers. Biotechnol. Progr. 21, 1329–1334. Yu, J., Tao, R., Yu, K., 2012. Anaerobic biodegradation of benzene in salt marsh sediment of the Louisiana Gulf coast. Ecol. Eng. 40, 6–10. Yuan, S.Y., Chen, S.J., Chang, B.V., 2011. Anaerobic degradation of tetrachlorobisphenol-A in river sediment. Int. Biodeter. Biodegr. 65, 185–190.