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Ecological Engineering 49 (2012) 73– 76

Contents lists available at SciVerse ScienceDirect

Ecological Engineering

j ourna l ho me page: www.elsev ier .com/ locate /eco leng

hort communication

naerobic degradation of tetrabromobisphenol-A in river sediment

ea-Ven Chang ∗, Shaw-Ying Yuan, Yen-Lin Renepartment of Microbiology, Soochow University, Taipei, Taiwan

r t i c l e i n f o

rticle history:eceived 3 February 2012eceived in revised form 12 July 2012

a b s t r a c t

The contamination of the environment with tetrabromobisphenol-A (TBBPA), an endocrine disruptor, is aconcern. We examined anaerobic degradation of TBBPA in sediment samples from the Erren River in south-ern Taiwan. Anaerobic degradation of TBBPA was enhanced with the addition of humic acid (0.5 g L−1),

ccepted 10 August 2012vailable online 28 September 2012

eywords:naerobic degradationetrabromobisphenol-A

sodium chloride (1 mass/vol%), zero-valent iron (1 g L−1), vitamin B12 (0.025 mg L−1), brij 30 (55 �M), brij35 (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 eubacteriaare involved in the anaerobic degradation of TBBPA; sulfate-reducing bacteria is a major component ofthe sediment.

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. Introduction

Tetrabromobisphenol-A (TBBPA) is a flame retardant used in theroduction of many plastic polymers and electronic circuit boardsde Wit, 2002). It has been found in environmental samples andn human plasma and may have a toxic effect (Darnerud, 2003).

icrobial degradation is the primary mechanism for removal ofrganic toxic compounds in sediment (Yu et al., 2012). Reductiveehalogenation (e.g., substitution of Br or Cl by a hydrogen atom)

s an important mechanism (Fetzner, 1998; Davis et al., 2005).alogenated compounds are electron acceptors in respiratory orometabolic processes. Environmental factors such as tempera-ure, pH, salinity, plant species selection and the availability ofrganic carbon and/or inhibiting substances influence the growthnd activity of microbes, and the manipulation of some has beennvestigated (Faulwetter et al., 2009). The addition of sodium chlo-ide, humic acid, zero-valent iron, surfactants, electron donors orlectron acceptors influences the anaerobic degradation of organicoxic chemicals in sediment (Chang et al., 2009). However, little isnown about the effects of factors on the anaerobic degradation ofBBPA in river sediment.

The climatic characteristics of subtropical regions foster diverseicrobial communities (Chang et al., 2009). Several techniques

sed to study microbial communities in environmental samples

nclude phospholipid fatty acid analysis (e.g., Langer and Rinklebe,009, 2011) and molecular-biological methods (Faulwetter et al.,009). Many studies have used PCR-denaturing gradient gel

∗ Corresponding author. Tel.: +886 228806628; fax: +886 228831193.E-mail address: bvchang@mail.scu.edu.tw (B.-V. Chang).

TB1pca

925-8574/$ – see front matter © 2012 Elsevier B.V. All rights reserved.ttp://dx.doi.org/10.1016/j.ecoleng.2012.08.038

© 2012 Elsevier B.V. All rights reserved.

lectrophoresis (PCR-DGGE) to examine the effect of pollutants onicrobial communities in sediment (Castle et al., 2006; Chang et al.,

009). Little information is available on the effect of TBBPA anaer-bic degradation on the change in bacterial communities in riverediment. We aimed to examine the effect of factors on the anaero-ic degradation of TBBPA in the sediment of Erren River, one of theost heavily contaminated rivers in southern Taiwan, and changes

n the microbial community in the sediment.

. Materials and methods

.1. Chemicals

TBBPA (98.0%) analytical standard was from Aldrich Chemi-al Co. (Milwaukee, WI). Solvents were from Mallinckrodt, Inc.Paris, KY). The biosurfactants used in this study were surfactinnd rhamnolipid as described by Yeh et al. (2005) and Wei et al.2005), respectively. All other chemicals were from Sigma Chemi-al Co. (St. Louis, MO). The log Kow for TBBPA, tribromobisphenol-A,ibromobisphenol-A and monobromobisphenol-A was 4.5, 2.1, 2.1,nd 3.7, respectively.

.2. Sampling and medium

We collected sediment samples from Erren River in July 2008.he three sampling sites, A (22.55◦10.98′N, 120.11◦3.51′E),

(22.55◦14.32′N, 120.11◦12.9′E) and C (22.54◦51.13′N,

20.13◦27.01′E) are well known from previous studies of aquaticollutants (Yuan et al., 2011). The sediments (>15 cm) wereollected by use of a soil core during low tide. Adaptation involveddding 50 �g g−1 TBBPA to 500 g sediment at 14-d intervals under

7 ical Engineering 49 (2012) 73– 76

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

Treatment k 1 (d−1) t1/2 (d) ra

Inoculated controlb 0.0417 16.6 0.98Humic acid (0.5 g L−1) 0.0491 14.1 0.94Sodium chloride (1%) 0.0502 13.8 0.95Zero-valent iron (1 g L−1) 0.0541 12.8 0.92Vitamin B12 (0.025 mg L−1) 0.0529 13.1 0.89Rhamnolipid (130 mg L−1) 0.0686 10.1 0.94Surfactin (43 mg L−1) 0.0582 11.9 0.92Brij 30 (55 �M) 0.0510 13.6 0.94Brij 35 (91 �M) 0.0554 12.5 0.95Sodium acetate (30 mM) 0.0297 23.3 0.95Sodium lactate (20 mM) 0.0276 25.1 0.96Sodium pyruvate (20 mM) 0.0340 20.4 0.97Sodium hydrogen carbonate (20 mM) 0.0582 11.9 0.89Sodium sulfate (20 mM) 0.0679 10.2 0.95Sodium nitrate (20 mM) 0.0525 13.2 0.96

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4 B.-V. Chang et al. / Ecolog

tatic incubation at 30 ◦C without light for 6 months. Here, sedi-ent refers to TBBPA-adapted sediment. For sites A, B, and C, the

BBPA concentration in sediment was 260, 450, and 38.1 ng g−1,espectively. In our previous study of the degradation of TCBPAn sediment from the 3 sampling sites, the anaerobic degradationate of TCBPA was higher in site B than other sites (Yuan et al.,011). Therefore, we used the sediment sample from site B in theollowing experiments.

The experimental medium consisted of (in g L−1) NH4Cl, 2.7;gCl2·6H2O, 0.1; CaCl2·2H2O, 0.1; FeCl2·4H2O, 0.02; K2HPO4, 0.27;

H2PO4, 0.35; yeast extract, 0.2; and resazurin, 0.001. The pHas adjusted to 7.0 after autoclaving; 0.9 mM titanium citrate was

dded as a reducing reagent.

.3. Experimental design

All experiments involved use of 125-mL serum bottles con-aining 45 mL medium, 5 g river sediment and 50 �g g−1 TBBPA.

e measured the effect of the following factors on anaerobicegradation in sediment collected from site B: sodium chloride1 mass/vol%); humic acid (0.5 g L−1); zero-valent iron (1 g L−1);itamin B12 (0.025 mg L−1); surfactants, brij 30, brij 35, rham-olipid, and surfactin at 1 critical micelle concentration (CMC;he CMC values were 55 �M, 91 �M, 130 mg L−1, and 43 mg L−1,espectively); the electron donor sodium acetate (30 mM), sodiumactate (20 mM) or sodium pyruvate (20 mM); sodium hydrogenarbonate (20 mM), sodium sulfate (20 mM), or sodium nitrate20 mM) for methanogenic, sulfate-reducing or nitrate-reducingonditions, respectively; and microbial inhibitors (50 mM BESA,0 mM vancomycin, 50 mM sodium molybdate-2-hydrate). Theoncentrations of these factors were from previous studies (Yuant al., 2011). Inoculated control samples (without sodium hydrogenarbonate, sodium sulfate, or sodium nitrate), considered non-terile sediment, were shaken before incubation at 30 ◦C and pH 7.0n the dark. Sterile controls were autoclaved at 121 ◦C for 30 minn 3 d.

All experiments were conducted in an anaerobic glove boxForma Scientific, USA) filled with N2 (85%), H2 (10%), and CO2 (5%).he 125-mL serum bottles were capped with butyl rubber stoppers,rapped in aluminum foil to prevent photolysis, and incubatedithout shaking at 30 ◦C in the dark. Each treatment was applied

n triplicate. Samples were collected every 7 d to measure residualBBPA, then underwent PCR-DGGE. Methane was sampled fromhe headspace of the serum bottles.

.4. Analytical methods

TBBPA was extracted twice from sediment samples by use ofichloromethane, then again 20 min at 30 ◦C with use of a Bran-on 5200 ultrasonic cleaner. Extracts were analyzed by use of gashromatography (Hewlett Packard 6890) equipped with an elec-ron capture detector and HP-5 capillary column. The initial columnemperature was set at 250 ◦C, increased by 2 ◦C min−1 to 260 ◦C,nd then increased by 10 ◦C min−1 to 280 ◦C. Injector and detec-or temperatures were set at 300 and 320 ◦C, respectively. Theecovery percentage and detection limit for TBBPA was 91.5% and.02 mg L−1, respectively. The anaerobic degradation products ofBBPA and methane levels were analyzed as we described previ-usly (Chang et al., 2011).

.5. DNA extraction and PCR-DGGE analysis

DNA was extracted from sediment samples using the Mo BioowerSoil DNA kit (Carlsbad, CA). The primer sequences for DGGEnalysis were for FGC968 (Escherichia coli position 968–983),

tsed

ach 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.

′-GCCCGGGGCGCGCCCGGGCGGGGCGGGGGCACGGGGGAACGCGAAGAACCTTAC-3′, and R1401 (E. coli position401–1385), 5′-CGGTG TGTACAAGACCC-3′ (Chang et al., 2009).GGE involved use of a D-gene and D-code system (Bio-Radaboratories, CA, USA). Electrophoresis involved 1× TAE buffer atoltage 60 V and temperature 60 ◦C for 16 h. The bands of interestere excised and soaked in elution buffer overnight at 37 ◦C. TheNA was re-amplified with the primers for FGC968 and R1401.he re-amplified products were again purified and sequenced withse of an ABI-Prism automatic sequencer.

.6. Data analysis

The TBBPA biodegradation data collected for this study fit wellith first-order kinetic equations: S = S0 exp (−k1t), t1/2 = ln 2/k1,here t is time, S0 is the initial substrate concentration, S is the

ubstrate concentration at time t, and k1 is the degradation rateonstant. Principal component analysis (PCA) was used to examinehe DGGE community structure. Statistical analysis involved use ofPSS v10.0 (SPSS Inc., Chicago, IL, USA).

. Results and discussion

.1. Effects of various factors on the anaerobic degradation ofBBPA in the sediment

The TBBPA concentrations in the sterile controls were firstxamined at the end of the 35-d incubation. The proportion ofBBPA ranged from 92.1% to 95.3%. Therefore, the TBBPA degra-ation in the following experiments was due to microbial action.he degradation rate and half-life of TBBPA were 0.0417 d−1 and6.6 d, respectively (inoculated control) (Table 1). As comparedith the inoculated control, the addition of humic acid, sodium

hloride, zero-valent iron, and vitamin B12 enhanced the degrada-ion rate of TBBPA by 17.7%, 20.4%, 29.7%, and 26.9%, respectively.umic acid showed increased showed a higher reducing capacity

n deeper layers, probably because of reduction by humic-acid-educing microorganisms (Kappler et al., 2004). TBBPA degradationas enhanced by the addition of sodium chloride (1%). The types

f bacteria colonized in sediment and their biodegradation poten-

ial are affected by salinity (Tam et al., 2002). The high salinity ofample sediment may significantly inhibit the degradation rate (Yut al., 2012). The addition of zero-valent iron enhanced the degra-ation of TBBPA. Zero-valent iron can be an electron donor and can

B.-V. Chang et al. / Ecological Engineering 49 (2012) 73– 76 75

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to3cianalyses revealed sediment with a specific bacterial community.Treatment with different substrates changed the microbial com-munities in sediment samples. The results support the feasibility ofremoving TBBPA in sediment by anaerobic degradation. We will use

Table 2TBBPA anaerobic degradation rate constant (k1) and half-life (t1/2) in sediment withthe addition of 3 microbial inhibitors.

Treatment k1 (d−1) t1/2 (d) ra

Inoculated controlb 0.0417 16.6 0.98Vancomycin (50 mM) 0.0318 21.8 0.95BESA (50 mM) 0.0270 25.7 0.96

Fig. 1. Proposed anaerobic biotransfor

fficiently remove PAH (Chang et al., 2007). The TCE dechlorinationnd cellular growth rates doubled when vitamin B12 concentrationas increased 25-fold to 0.025 mg L−1 (He et al., 2007).

The non-ionic surfactants brij 30 and brij 35 enhanced TBBPAegradation rates by 32.9% and 22.3%, respectively, whereashamnolipid and surfactin elevated degradation rates by 64.5%nd 39.6%, respectively. The use of surfactants can increase theegradation of hydrophobic organic compounds in contaminatednvironments by increasing the total aqueous solubility of theseompounds (Yeh et al., 2005; Wei et al., 2005). The addition ofhamnolipid yielded a higher TBBPA degradation rate than did thether additives. The rhamnolipid used in this study, a commonlysolated glycolipid biosurfactant, was produced by Pseudomonaseruginosa J4, whereas while the surfactin, a lipoprotein-type bio-urfactant, was produced by Bacillus subtilis ATCC 21332. Thisnding is consistent with that of Yuan et al. (2011), who reported onhe anaerobic degradation of TCBPA in sediment. We also noted annhibition of TBBPA degradation with the addition of acetate, lac-ate, or pyruvate. Acetate, lactate, and pyruvate may not functions electron donors under our conditions. This result is similar toith our previous studies of anaerobic degradation of nonylphenol

n sediment (Chang et al., 2009).

.2. Biotransformation of TBBPA in the sediment

We monitored the intermediate products of the degrada-ion of TBBPA in sediment at 0, 20, and 35 d. We observed

intermediate products, dibromobisphenol-A and BPA, at d 20nd d 35, respectively. The concentration of TBBPA decreasedrom 0, 20, and 35 d, and the 2 products of TBBPA degradationn sediment were obtained. Dehalogenation of TBBPA is proba-ly a stepwise process of sequential removal of bromine atoms.rom this result, we propose the following biochemical path-ay: TBBPA → → dibromobisphenol-A → → BPA (Fig. 1). Ronen

nd Abeliovich (2000) observed the reductive degradation of TBBPAo BPA in anaerobic sediment from a wet ephemeral desert streamed. Similar results were found with transformation of TBBPA inediment (Voordeckers et al., 2002; Davis et al., 2005). BPA did notnaerobically degrade after 140 d in the sediment (Chang et al.,011). BPA has an OH substituent on both aromatic rings, which

s a possible site for degradation, but the 2 rings are joined by auaternary carbon, which may inhibit its degradation by anaerobicicrobes.

.3. Comparison among various electron acceptors and microbialnhibitors of the anaerobic degradation of TBBPA

Adding electron acceptors sodium hydrogen carbonate, sodiumulfate, or sodium nitrate increased the degradation rate of TBBPAy 39.6%, 62.8%, and 25.9%, respectively (Table 1). Comparedith the inoculated control, methanogenic, sulfate-reducing anditrate-reducing conditions enhanced TBBPA degradation, the

egradation rate being sulfate-reducing > methanogenic > nitrate-educing conditions. Adding molybdate (a selective inhibitorf sulfate-reducing bacteria), BESA (a selective inhibitor ofethanogens), or vancomycin (a selective inhibitor of eubacteria)

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n pathways of TBBPA in the sediment.

Arbeli et al., 2006) decreased the degradation rate of TBBPA by6.3%, 64.7%, and 57.1%, respectively (Table 2). The degradationates for TBBPA decreased greatly when the activities of sulfate-educing bacteria were inhibited. Sulfate-reducing bacteria belongo eubacteria (domain Bacteria), whereas the methanogens arelassified within the domain Archaea. The methanogens can onlytilize simple substrates such as H2/CO2 and acetate, but theulfate-reducing bacteria can use a wide spectrum of organic com-ounds for growth in the presence or absence of sulfate (Yuant al., 2011). In addition, methane production was not detectedith TBBPA degradation under the three reducing conditions

fter 35 d of incubation. Our results indicate that sulfate-reducingacteria constitute a major microbial component in TBBPA degra-ation, with methanogen and eubacteria microbial populationslso involved.

.4. Microbial community analysis

The microbial community changes in DGGE band profilesnd PCA with various treatments after 35 d of incubation arehown in Fig. 2. The DGGE profile consisted of at least 6 bands,nd the number of bands was changed with various treat-ents. The first principal component (PC1 = 47.8%) and second

rincipal component (PC2 = 41.6%) explained 89.4% of the vari-tion and discriminated between samples with various TBBPAoncentrations. Microbial communities differed significantly inon-TBBPA-adapted and TBBPA-adapted sediment. Microbial com-unities changed significantly when various substrates were

dded to the sediment. Highly similar microbial communitiesere also found in samples with sodium chloride, humic acid,

ero-valent iron, brij 30, and brij 35, which can enhance TBBPAegradation. This observation is consistent with our previousnding that treatment with different substrates affects microbialommunities in the sediment (Chang et al., 2009).

In summary, anaerobic degradation of TBBPA is a major processhat results in decontamination of river sediments. The additionf humic acid, sodium chloride, zero-valent iron, vitamin B12, brij0, brij 35, rhamnolipid, surfactin, acetate, lactate, or pyruvatean influence the degradation of TBBPA. The optimal conditionnvolved the addition of rhamnolipid into the sediment. DGGE

Molybdate (50 mM) 0.0238 29.1 0.97

ach 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.

76 B.-V. Chang et al. / Ecological Engineering 49 (2012) 73– 76

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imilar conditions to further define the operating parametersor river sediment bioremediation and to monitor the microbialynamics involved in the degradation of TBBPA in river sediment.

cknowledgment

This research was supported by the National Science Council,epublic of China, Grant Number 98-2627-B-031-001.

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