Application of a novel gene encoding bromophenol dehalogenase from Ochrobactrum sp. T in TBBPA degradation

Accepted Manuscript Application of a novel gene encoding bromophenol dehalogenase from Ochrobactrum sp. T in TBBPA degradation Zhishu Liang, Guiying Li, Bixian Mai, Huimin Ma, Taicheng An PII:

S0045-6535(18)32101-5

DOI:

https://doi.org/10.1016/j.chemosphere.2018.11.004

Reference:

CHEM 22488

To appear in:

ECSN

Received Date: 14 April 2018 Revised Date:

27 October 2018

Accepted Date: 1 November 2018

Please cite this article as: Liang, Z., Li, G., Mai, B., Ma, H., An, T., Application of a novel gene encoding bromophenol dehalogenase from Ochrobactrum sp. T in TBBPA degradation, Chemosphere (2018), doi: https://doi.org/10.1016/j.chemosphere.2018.11.004. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

ACCEPTED MANUSCRIPT

AC C

EP

TE D

M AN U

SC

RI PT

Graphic Abstract

ACCEPTED MANUSCRIPT 1

Application of a novel gene encoding bromophenol dehalogenase

2

from Ochrobactrum sp. T in TBBPA degradation

3

Zhishu Liang,a, c Guiying Li,b,* Bixian Mai,a Huimin Ma,a Taicheng Anb

RI PT

4 5

Chinese Academy of Sciences, Guangzhou 510640, China;

7 8

2

Guangzhou Key Laboratory of Environmental Catalysis and Pollution Control, School of Environmental Science and Engineering, Institute of Environmental Health and Pollution

9

Control, Guangdong University of Technology, Guangzhou 510006, China;

10 11

State Key Laboratory of Organic Geochemistry, Guangzhou Institute of Geochemistry,

SC

1

M AN U

6

3

University of Chinese Academy of Sciences, Beijing 100049, China;

13

TE D

12

Running Title: Gene encoding the dehalogenase for TBBPA degradation

16 17 18

* Corresponding author:

AC C

15

EP

14

Prof. Guiying Li

Tel: 86-20-23883536 E-mail address: [email protected].

19

1

ACCEPTED MANUSCRIPT ABSTRACT

21

Tetrabromobisphenol-A (TBBPA), a typical brominated flame retardant, leaked from

22

commercial products into the environments has attracted people’s attention around the world.

23

Ochrobactrum sp. T capable of degradation and mineralization of TBBPA was isolated in our

24

early work. In this study, the identification of TBBPA-degrading gene from the strain was

25

further carried out by combining whole-genome sequencing with gene cloning and expression

26

procedures. In total, 3877 open reading frames were found within 3.9 Mb genome and seven

27

of them were identified as dehalogenating-relating genes. One gene with a significant ability

28

to degrade TBBPA was designated as tbbpaA. Sequence alignments analysis showed that it

29

shared 100% identity with haloacid dehalogenases. Furthermore, tbbpaA gene was cloned and

30

expressed into E. coli to achieve a constructed strain. Like the original strain, the constructed

31

strain could degrade TBBPA (6 mg L-1) with 78% of debromination efficiency and 37.8%

32

mineralization efficiency within 96 h. Gene expression study revealed that tbbpaA was

33

up-regulated in the presence of TBBPA. Overall, we report the identification of a functional

34

TBBPA-degrading gene in an aerobe, which can deepen the knowledge of enhancing TBBPA

35

removal by Strain T at the genetic level and facilitate in situ TBBPA bioremediation.

36

Keywords: TBBPA; Ochrobactrum sp. T; Dehalogenase; Gene expression; Constructed strain

37

1. Introduction

AC C

EP

TE D

M AN U

SC

RI PT

20

38

Brominated flame retardants (BFRs) are widely used as additives in commercial

39

products such as furniture, building materials, electronic equipment, carpets and other thermal

40

insulation materials to prevent ignition (Yang et al., 2015). Given this, the leaching of BFRs

41

into the environments during the manufacturing, transferring, recycling, and disposal of these 2

ACCEPTED MANUSCRIPT products is inevitable. The seriousness of this problem is compounded by constantly

43

increasing annual outputs of BFRs. It is reported that the annual global consumption of BFRs

44

was increased from 0.13 million tons in 2002 to over 0.2 million tons in 2013 (Xiong et al.,

45

2015). As one of the most frequently used BFRs, tetrabromobisphenol-A (TBBPA) has been

46

detected in various environmental matrices including water, soils, sediments, aquatic animals,

47

and even human tissues, maternal and cord serum and breast milk (Sun et al., 2014; Li et al.,

48

2015b; Wang et al., 2015b). Due to its high toxicity, bioaccumulation and lipophilicity,

49

significant health risk on humans and ecosystems including neurotoxicity, immunotoxicity,

50

endocrine disruption, and hepatocytes destruction will be posed by TBBPA (Peng et al., 2017).

51

As such, it is necessary to remove them completely and cost-effectively.

M AN U

SC

RI PT

42

Biodegradation with reductive dehalogenation bacteria has proved as a promising way to

53

clean up the sites contaminated with halogenated pollutants. This is because of the fact that

54

the bromine/chlorine of these compounds can be replaced with hydrogen stemmed from

55

molecular hydrogen or other oxidizable compounds such as succinate, propionate, pyruvate,

56

lactate, acetate and formate by bacteria to supply energy for their growth or serve as election

57

acceptors (Ding and He, 2012; Richardson, 2016). As a result, with the breaking of C-halide

58

bond through reduction and the production of corresponding halide anion and the

59

nonhalogenated product, the toxicity of these compounds is significantly reduced or

60

completely removed (Hug et al., 2013). Recently, a growing number of bacteria were

61

successfully isolated with the ability to degrade different contaminates and these bacterial

62

strains include facultative anaerobe strains and obligate aerobes (Comamonas sp. (Peng et al.,

63

2013), Desulfitobacterium (Atashgahi et al., 2016), Ochromobacterium sp. (Zu et al., 2014),

AC C

EP

TE D

52

3

ACCEPTED MANUSCRIPT Bacillus sp. (Zu et al., 2012)) and obligate anaerobe bacteria (Dehalococcoides,

65

Dehalogenimonas, Dehalobium, Dehalobacter sp.) (Ding and He, 2012; Richardson, 2013).

66

Among them, Dehalococcoides is the most well studied and documented strain because it

67

exhibits versatile dechlorination activity to hazardous chlorine-containing chemicals such as

68

polychlorinated biphenyls and trichloroethene (Tas et al., 2010).

RI PT

64

Furthermore, with the availability of modern molecular techniques such as polymerase

70

chain reaction (PCR)-based fingerprinting methods, high throughput techniques, microarray

71

and next-generation sequencing, the enzymes designated as reductive dehalogenases (RDases)

72

from these strains and the corresponding encoding genes were also studied widely, to deeply

73

understand the molecules mechanism of these strains to degrade these compounds (Fricker et

74

al., 2014; Liang et al., 2015). For example, previous genomic annotation has led to the

75

discovery of 1,469,720 and 1,395,502 bp genome of D. ethenogenes strain 195 and strain

76

CBDB1, which possess 17 and 32 rdh genes potentially encoding RDases for chlorinated

77

compounds-degrading, respectively (Kube et al., 2005; Seshadri et al., 2005). In addition, the

78

subsequent analysis of the draft genome (1,462,509 bp) of Dehalococcoides mccartyi JNA

79

revealed the presence of 29 putative RDase genes (Wang et al., 2015a). However, as for

80

TBBPA-degrading strain, Ochrobactrum sp. T (Strain T), only few recent works focused on

81

its biodegradation kinetics and mechanisms of TBBPA (An et al., 2011; Zu et al., 2014;

82

Xiong et al., 2015; Li et al., 2016) without further investigating its related functional

83

dehalogenase genes.

AC C

EP

TE D

M AN U

SC

69

84

Therefore, the primary objective of this study is to explore potential brominated

85

compounds removal encoding genes using a series of molecular biological techniques. By 4

ACCEPTED MANUSCRIPT genome sequencing of Strain T, the open reading frames (ORFs) related to the dehalogenase

87

could be achieved. Then these genes were systematically cloned and expressed in E. coli to

88

examine the potential of their expression enzymes for TBBPA degradation. Furthermore, the

89

biodegradation efficiency and cell growth were extensively compared between the

90

constructed strain with wild-type bacterium. In addition, both debromination and

91

mineralization of TBBPA were also detected by the constructed strain cultivated with TBBPA.

92

Lastly, the transcript expression level of the target gene during the TBBPA biodegradation

93

process by these two strains was also analyzed using quantitative real-time PCR (qPCR). This

94

information will help us to deeply understand the TBBPA biodegradation mechanism at the

95

genetic level.

96

2. Experimental section

97

2.1. Chemicals, microorganisms, media and cultivation conditions

TE D

M AN U

SC

RI PT

86

TBBPA (97%) and 2,4,6-tribromophenol (TBP, 99%) were purchased from

99

Sigma-Aldrich (St. Louis, Mo, USA). Bisphenol A (BPA, 97%) was from Acros Organics

100

(New Jersey, USA). All other chemicals were of analytical grade and obtained from

101

Guangzhou Chemical Reagent Co., Inc., China.

AC C

EP

98

102

Strain T (HM543185) was previously isolated and identified by our group (An et al.,

103

2011) from a sludge sample collected from an electronic waste recycling site. The detail

104

recipe of growth medium (GM) and the mineral medium (MM) used for enrichment and

105

biodegradation of TBBPA by this strain were provided in the Supporting Information (SI). E.

106

coli DH5a and E. coli BL21 (DE3) (Tiangen) were used as hosts for cloning and gene

5

ACCEPTED MANUSCRIPT 107

expression, respectively. The plasmids pMDTM18-T (Takara) and pET30a (+) (Novagen) were

108

used for cloning and expression vectors, respectively.

109

2.2. Genome sequencing and nucleotide sequence accession numbers The whole genome sequencing of Strain T was conducted and compared with the

111

sequences in the NCBI database to provide organism-specific genomic template for the

112

molecular mechanism analysis of TBBPA degradation (details in SI). The sequence was

113

trimmed and assembled based on a previous reference (Liang et al., 2016). This draft genome

114

project has been deposited at DDBJ/EMBL/GenBank under the accession number of

115

LXEK00000000.

116

2.3. PCR amplification, cloning and sequencing of putatively tbbpa dehalogenase gene

M AN U

SC

RI PT

110

In total, 24 ORFs, putatively related to encode the dehalogenases capable of degrading

118

halogenated compounds were amplified by PCR using genomic DNA as template and the

119

specific primers. Emphasis would be given on designing specific primers targeting

120

dehalogenases genes using Primer 3 Plus software based on selected ORFs. A total of 24

121

primer sets listed in Table S1 were synthesized by integrated DNA technologies and tested for

122

the specificity using primer-BLAST on NCBI web site. BamHI (in the forward primer) and

123

EcoRI or XhoI (in the reverse primer) restriction sites were introduced to ensure the correct

124

direction of gene insertion. Experimental details of PCR amplification were provided in SI.

AC C

EP

TE D

117

125

The purified PCR products of individual genes amplified with specify primers were

126

ligated into the pMDTM18-T Vector (Takara, China) and transformed into E. coli TOP10

127

chemically competent cells according to the manufacturer’s protocol. The blue-white screen

128

was used for rapid detection of constructed bacteria, and white colonies on 6

ACCEPTED MANUSCRIPT ampicillin-resistant solid plates were enriched to extract the plasmid based on the alkaline

130

lysis method using TaKaRa MiniBEST Plasmid Purification Kit. In all experiments, E. coli

131

cells carrying empty pMDTM18-T were used as a negative control. The correct insert was

132

identified using PCR amplification and verified by DNA sequencing on an ABI 3730 xl DNA

133

Analyzer using the M13 forward and reverse primers, which bind to sites flanking the

134

insertion site on the plasmid.

135

2.4. Expression and identification of TBBPA dehalogenase

SC

RI PT

129

Target gene fragments were obtained from the positive clone plasmids digested with

137

BamHI and EcoRI or XhoI simultaneously. After purification, these genes were inserted into

138

T7 expression vector pET30a (+) (Novagen), which had been predigested with the same

139

restriction enzymes. Ligated plasmids were transformed into E. coli BL21 (DE3) competent

140

cells and grew on LB plate supplemented with 30 µg mL-1 kanamycin to select positive

141

constructed bacteria. Then, 100 µL of the above enriched constructed strain was inoculated

142

into solid MM plate with 6 mg L-1 TBBPA for 24-48 h at 37 oC. Strain successfully survived

143

on this plate was considered as the functional bacterium with ability to use TBBPA as carbon

144

and energy source, which was selected as the target constructed strain. Meanwhile, one clone

145

carrying an empty pET30a (+) was cultured under the identical conditions as the negative

146

control. Then, PCR and sequencing of the plasmid from the constructed strain was used to

147

verify whether the corresponding genes fragment has been successfully inserted into the

148

expression vector. Finally, the encoding gene expression procedure was carried out according

149

to the standard method (Michael. and Joseph., 2012). Detailed steps were provided in SI.

150

2.5. Growth pattern and TBBPA-degrading activity of the constructed strain

AC C

EP

TE D

M AN U

136

7

ACCEPTED MANUSCRIPT 151

As a newly constructed strain, for further application in TBBPA degradation, its growth

152

pattern, the debromination and mineralization activity were also determined (details in SI).

153

2.6. RNA extraction, genomic DNA removal and reverse transcription Total RNA was extracted from liquid samples at each sampling time (0, 24, 48, 72 and

155

96 h) using Trizol reagent (Invitrogen) according to the manufacturer’s instruction. Its

156

concentration was determined at OD260 using a NanoDrop 2000 (Thermo), and the purity and

157

quality were analyzed by measuring OD260/OD280 ratios. First-strand cDNA synthesis was

158

performed with 1 µg of total RNA using PrimeScript™ RT reagent Kit with gDNA Eraser

159

(Takara, China) in a total volume of 20 µL as described by the manufacturer. Prior to

160

qRT-PCR, purified RNA and synthesized cDNA were stored at -80 oC.

161

2.7. Quantitative real-time polymerase chain reaction

M AN U

SC

RI PT

154

The qPCR analysis was performed on the CFX96™ real-time system instrument

163

(Bio-Rad) using the SYBR® Premix Ex Taq™ II (Tli RNaseH Plus) (Takara, China), in a final

164

volume of 25 µL, containing 2 µL of the above cDNA sample (100 ng), 12.5 µL SYBR®

165

Premix, 8.5 µL of nuclease-free water and 1 µL of each primer (0.4 µM). The specific primers

166

were designed using Primer 3 Plus program (Table 1). Triplicate reactions were performed for

167

each sample using the thermal program provided in SI. Each PCR run included a no-template

168

control (with water instead of cDNA) and a no-RT negative control (with total RNA instead

169

of cDNA). The 16s rRNA was selected as an endogenous housekeeping gene because its

170

average mRNA expression levels remain relatively stable (Johnson et al., 2005; Xiu et al.,

171

2010; Li et al., 2015a; Yu et al., 2015). Therefore, to correct the variations in the amount of

172

RNA applied to real-time PCR, the mRNA expression levels of 16s rRNA were used to

AC C

EP

TE D

162

8

ACCEPTED MANUSCRIPT normalize that of TBBPA-degrading genes in the samples. Relative mRNA expression was

174

calculated with the 2−∆∆Ct method (Livak and Schmittgen, 2001). Moreover, for a valid

175

reaction, the amplification efficiencies (E) and linear standard curves (R2) of the target gene

176

as well as the reference (16S rRNA) gene expression levels must be within the range of 100%

177

± 10% and ≥0.99, respectively (Bustin et al., 2009; Padilla-Crespo et al., 2014). Student’s t

178

test was used to determine whether up-regulation or down-regulation of gene expression from

179

different treatments is statistically significant (Li et al., 2015a). When p<0.05, differences

180

were considered statistically significant relative to baseline or control conditions.

181

2.8. Concurrent degradation of TBBPA

M AN U

SC

RI PT

173

Due to most of the additional electron donor and acceptor involved in electron transfer

183

chain are essential to organic compound metabolism, especially the redox related process (Zu

184

et al., 2014), the effects of electron donors and acceptors on the debromination need to be

185

systematically evaluated. The TBBPA biodegradation by the constructed strain was

186

investigated with addition of approximately 0.1 g carbon sources and electron donors,

187

including ethanol, glucose, NH4NO3, sodium acetate and sodium citrate into MM mixture

188

(100 mL) containing 6 mg L-1 TBBPA. In addition, the influence of two major TBBPA

189

biodegradation intermediates, TBP and BPA (6 mg L-1), on TBBPA debromination was also

190

investigated using a culture only added substrate TBBPA as the reference. All measurements

191

were progressed independently three times.

192

2.9. TBBPA concentration, bromide concentration and TOC analysis

AC C

EP

TE D

182

9

ACCEPTED MANUSCRIPT The TBBPA concentration, bromide concentration and TOC were also systemically

194

analyzed during the TBBPA biodegradation process by the constructed strain. The detailed

195

methods were provided in SI.

196

3. Results and discussion

197

3.1. Draft genome sequencing

RI PT

193

The genome information and responding annotation are important assets to better

199

understand the physiology and biodegrading potential of Strain T, and are valuable for future

200

research to enhance the bioremediation of brominated contaminants. The first step to identify

201

gene encoding for TBBPA metabolism was draft genome sequencing. Information obtained

202

from the assembly of the Strain T genome was described in our previous paper (Liang et al.,

203

2016). Briefly speaking, a 3.9 Mb complete genome sequence of Strain T was achieved.

204

Among them, total 2976 proteins could be assigned to the Clusters of Orthologous Groups

205

(COGs) families (Fig. S1). These proteins were annotated to different categories, such as

206

metabolism, genetic information processing and environmental information processing. The

207

proteins associated with amino acid transport and metabolism were the most abundant group

208

of COGs (411 ORFs, accounted for 13.8% of total ORFs), followed by those associated with

209

inorganic ion transport and metabolism (270 ORFs, 9.1%), and those proteins function

210

unknown (323 ORFs, 10.8%). Therefore, the genes possibly responsible for the TBBPA

211

biodegradation were analyzed in the genome of Strain T. Although no bromophenol or

212

brominated compounds dehalogenase encoding genes was found in the genome, there do exist

213

some genes encoding enzymes that could bioremediate chlorinated compounds and phenol.

AC C

EP

TE D

M AN U

SC

198

10

ACCEPTED MANUSCRIPT As shown from Table S2, in total, seven ORFs (orf 04638, orf 05726, orf 00062, orf 04028,

215

orf 04434, orf 01638 and orf 06062) encoding haloacid dehalogenases and two ORFs (orf

216

01379 and orf 00734) encoding pentachlorophenol 4-monooxygenase and monophenol

217

monooxygenase, were found. Furthermore, four ORFs (orf 05787, orf 05349, orf 05348, orf

218

06285) encoding dichlorophenolindophenol oxidoreductases genes, three ORFs (orf 06401,

219

orf 06400 and orf 05850) encoding phenolic acid decarboxylases genes and other eight ORFs

220

encoding phenol-relevant enzymes such as hydroxylase, oxidase and hydrolase were also

221

obtained.

M AN U

SC

RI PT

214

In summary, these proteins were characterized as dehalogenases, monooxygenases,

223

oxidoreductases, decarboxylases, hydroxylase, oxidase and hydrolase, and were related

224

functionally with the degradation of environmental contaminants, including monobromoacetic

225

acid and monochloroacetic acid (Chiba et al., 2009), phenolic and non-phenolic substrates

226

(Coconi-Linares et al., 2015), phenol and toluene (Heinaru et al., 2016). Due to these

227

xenobiotic biodegradation and metabolism orthology may account for more of the

228

functionally annotated upregulated proteins than any other orthology (Zhou et al., 2015), they

229

were most likely to be involved in TBBPA degradation, and as such were given particular

230

focus in our work. Therefore, these ORFs were selected to identify the putative genes

231

catalyzed the TBBPA biodegradation by the following gene cloning and expression.

232

3.2. Cloning, functional overexpression and confirmation of TBBPA dehalogenases-related

233

genes

AC C

EP

TE D

222

234

Total 24 ORFs putatively encoding TBBPA-degrading were PCR-amplified with 24 sets

235

of specific primers. Results indicated that specific bands of correct sizes had been obtained 11

ACCEPTED MANUSCRIPT (Fig. S2). By ligating target genes into cloning T vector and transforming into E. coli TOP10,

237

many white colonies and some blue colonies were observed on the solid LB plate

238

supplemented with Amp, X-gal and IPTG (transformation efficiency>1×108 colony-forming

239

unit (cfu) µg-1). The white colonies assumed as the positive constructed strains were further

240

validated with colony PCR. The results revealed that all the positive constructed strains

241

contained inserted fragments with correct sizes, indicating that these 24 TBBPA

242

dehalogenase-related genes were successfully cloned. That is, they could be used to perform

243

the following gene expression experiments.

M AN U

SC

RI PT

236

After digesting the plasmids extracted from constructed strains and vector with BamHI

245

and EcoRI or XhoI, some bands with the size corresponding to target genes can be observed

246

in the gel (data not shown), meaning that these plasmids have already been successfully

247

digested. Followed by ligating with the expression vector pET30a (+), the restriction

248

fragments were successfully expressed in E. coli BL21 (DE3) to generate constructed strain E.

249

coli BL21 (DE3) pET30a-a1 to pET30a-a24. Due to the fact that the addition of IPTG can

250

efficiently induce the expression of target genes (Nicolini et al., 2013), the TBBPA

251

degradation function of constructed strain was confirmed through inoculating them into a

252

solid MM plate with addition of IPTG and TBBPA. As Fig. S3 shows, the constructed strain

253

E. coli BL21 (DE3) pET30a-a6 (hereafter called constructed strain) was able to grow on the

254

TBBPA-containing plate, implying that this strain possessed the ability to use TBBPA as

255

carbon and energy source. As such, ORF 04638 was regarded as the target gene and

256

designated as tbbpaA. Whereas those constructed strains unable to survive on the

257

TBBPA-containing plate were regarded as the strain not carrying the TBBPA-degrading

AC C

EP

TE D

244

12

ACCEPTED MANUSCRIPT genes. Sequencing the inserted fragment of the plasmid from constructed strain showed that a

259

total of 1851 bp sequence was obtained, which were almost the same size as ORF 04638.

260

Besides, no mutation had occurred during the cloning and expressing process. This

261

confirmation has allowed us to deposit the sequence of TBBPA bromophenol dehalogenase

262

gene at GenBank under accession number KY483638.

RI PT

258

Moreover, SDS-PAGE analysis of the extracted total protein was carried out to identify

264

whether the enzyme responsible for the conversion of TBBPA was expressed or not. Results

265

showed that approximately 128.7 kDa band was present in the constructed strain, but absent

266

from the control strain E. coli BL21 (DE3) pET30a(+) (hereafter called control strain) (Fig.

267

S4), further suggesting that this protein is encoded by tbbpaA gene. However, this value is

268

almost twice as the size of the TBBPA bromophenol dehalogenase (616 amino acids), which

269

has a calculated molecular mass of 64.4 kDa, indicating that the enzyme is a dimer. Previous

270

study also achieved similar size (117 kDa) dehalogenase (BhbA protein) used to catalyze the

271

reductive dehalogenation reaction (Chen et al., 2013). Coincidently, this protein is also

272

approximately twice as large as previously reported respiratory reductive dehalogenase

273

(Bisaillon et al., 2010) and was able to catalyze the dehalogenation under the aerobic

274

conditions (Chen et al., 2013), further implying that dehalogenation can be conducted by

275

oxygen-tolerant reductive dehalogenase (Nijenhuis and Kuntze, 2016).

AC C

EP

TE D

M AN U

SC

263

276

Sequence alignments in Fig. 1 showed that this dehalogenase shared 100% identity with

277

the haloacid dehalogenases from Ochrobactrum sp. (WP 021588323) and Ochrobactrum

278

intermedium (WP 025091885). Thus, the above results indicated that the enzyme encoding by

279

gene tbbpaA is an oxygen-tolerant dehalogenase responsible for the conversion of TBBPA. 13

ACCEPTED MANUSCRIPT This is contradicted with other studies, which reported a two-step process including anaerobic

281

debromination and aerobic mineralization for biodegradation of TBBPA by sequential

282

anaerobic-aerobic strains (Ronen and Abeliovich, 2000; Liu et al., 2013). Whereas in

283

accordance with our previous finding, which revealed that the novel wild strain T has the

284

ability to debrominate and mineralize TBBPA in a one-step process under aerobic conditions

285

(An et al., 2011). This oxygen-tolerant dehalogenase were identified further verified this one

286

step TBBPA decontamination mechanism, suggesting that it may be functioned as both

287

dehydrogenase and debrominase, thus enabling it to simultaneously reductively debrominate

288

and oxidatively mineralize bromide phenols. For future application, the dehalogenase

289

activities of the enzyme from constructed strain need to be further analyzed.

290

3.3. The growth of the constructed strain, biodegradation and debromination of TBBPA

M AN U

SC

RI PT

280

The growth conditions of the constructed strain are important parameters to obtain high

292

biodegradation efficiency of TBBPA. Therefore, the dependence of debromination activity of

293

constructed strain and Strain T on the cell growth was further compared. As Fig. S5 shows, all

294

the strains have almost the same growth trend with lag and stationary phases of 4 and 16 h,

295

respectively, and the highest OD600 was obtained approximately 1.2 under the optimal

296

conditions of pH 7.0 and 30 oC. Considering the appropriate strain growth could ensure

297

enough cell amount and highest activity for the subsequent substrate degradation experiment

298

(Stasinakis et al., 2010), 14 h of the culture time, which was close to the end stage of the log

299

phase of the strain, was chosen as the enrichment time. Besides, the results of the genetic

300

stability of constructed strain showed that after 10, 20 and 30 passages, no significant

301

difference was found in the biodegradation efficiency for these constructed strains compared

AC C

EP

TE D

291

14

ACCEPTED MANUSCRIPT with the initial passage (Fig. S6). Further, PCR amplification and sequencing results showed

303

no mutation of the tbbpaA gene sequence, indicating that constructed plasmid pET30a-a6 had

304

high heredity stability in E. coli BL21 (DE3). All the above results proved that the obtained

305

constructed strain was a potential stable bacterium with TBBPA removal activity.

RI PT

302

TBBPA biodegradation profiles by these three bacteria were also compared. As shown in

307

Fig. 2a, different trends of the biodegradation activity existed among them, although these

308

strains behaved similarly during the enrichment process. To be specific, the constructed strain

309

exhibited much higher TBBPA biodegradation activity than the original Strain T. Complete

310

removal of TBBPA was seen within 96 h by constructed strain, while relatively slow removal

311

of TBBPA (89.7%) and no removal of TBBPA were observed by the Strain T and control

312

strain, respectively, under the same conditions.

M AN U

SC

306

Furthermore, a positive correlation of TBBPA removal and bromide ion production were

314

also evident. As shown in Fig. 2b, the bromide ion concentration increased in a time

315

dependent manner for different strains. For example, under the aerobic conditions, as reaction

316

proceeded from 24 to 96 h by the strain T, the Br- concentrations increased from 10.3 to 33.2

317

µM corresponding to the debromination efficiencies from 23.3% to 75.2%, with the

318

biodegradation efficiencies of TBBPA increased from 57.0% to 89.8%. While for the

319

constructed strain, after 96 h biodegradation of 6 mg L-1 TBBPA, 34.4 µM Br- was produced,

320

which is correspondent to the debromination efficiency of 78.0%. These results suggested that

321

most bromine atoms on the TBBPA molecule could be converted into Br- by the constructed

322

strain. Comparatively, the constructed strain had an equal but a slightly faster degradation and

323

debromination of TBBPA function than Strain T. Similar results of increase enzyme activity

AC C

EP

TE D

313

15

ACCEPTED MANUSCRIPT also found in the constructed Trichoderma reesei strains (Xiong et al., 2016). One possible

325

explanation is that the specific enzyme (dehalogenase) related to TBBPA bioremediation

326

from the constructed bacteria were extensively upregulated in response to TBBPA exposure

327

as compared with the strain T (Zhou et al., 2015). However, considering the previous papers

328

demonstrating that the microbial debromination was functioned by the dehalogenase

329

originated from anaerobic bacteria (Voordeckers et al., 2002; Arbeli et al., 2006), the results

330

that a constructed strain containing a dehalogenase encoding gene can degrade TBBPA under

331

the oxic condition was contradicted with them. Thus, from other aspect, further verified that

332

the dehalogenation can occur under both aerobic and anaerobic conditions (Arora and Bae,

333

2014; Nijenhuis and Kuntze, 2016).

M AN U

SC

RI PT

324

As revealed by our previous study that the Strain T could simultaneously debrominate

335

and mineralize TBBPA (An et al., 2011), the mineralization of TBBPA was also compared

336

between the constructed strain and Strain T through analyzing TOC removal. As shown in Fig.

337

2c, TBBPA mineralization efficiencies increased slightly during the first 24 h, followed by

338

increased rapidly during the next 48 h, and then leveled off with further prolonging of the

339

reaction time. In total, up to 37.8% and 35.7% of carbon contents of TBBPA (6 mg L-1) were

340

mineralized by the constructed strain and Strain T, respectively. This means that, similar with

341

the Strain T (Zu et al., 2012), the constructed strain could also simultaneously debrominate

342

and mineralize TBBPA. The reason why the degradation efficiency (100%) well exceeded the

343

mineralization efficiency may be attributed to the formation of some recalcitrant intermediates

344

during the biodegradation processes (Horikoshi et al., 2008). In addition, partial cleavage

AC C

EP

TE D

334

16

ACCEPTED MANUSCRIPT 345

rather than complete degradation of the TBBPA molecule during TBBPA biodegradation

346

could also result in lower mineralization and higher removal of TBBPA (Guo et al., 2014).

347

3.4. Cell density and the tbbpaA gene expression after exposure to TBBPA Analysis of the cell density revealed that the debromination of TBBPA was accompanied

349

by the growth of constructed strain. In cultures with initial TBBPA concentration of 6 mg L-1,

350

the cell densities of constructed strain increased from 1.5 × 106 to 3.3 × 106 cfu mL-1 as the

351

incubation time rise from 24 to 72 h (Fig. 3a), and these values were comparable to those of

352

Strain T. However, the cell densities did not increase significantly during the first 48 h

353

incubation time, suggesting that the cell growth was suppressed by the toxicity of TBBPA to

354

some extent (Zhang et al., 2007). Subsequently, a steep increase of the cell density was found

355

for both bacterial strains during 48 to 72 h after a period of steadily adapting to the TBBPA

356

environments. It should be noted that, during the later stage, the densities of the constructed

357

strain were higher than that of Strain T, which is matching well with the degradation curve of

358

TBBPA (Fig. 2a). Eventually, the cell densities decreased as the TBBPA was gradually

359

consumed. This is an additional evidence for the growth of both bacterial strains by using

360

TBBPA as respiratory electron acceptor and carbon source (Yang et al., 2015).

AC C

EP

TE D

M AN U

SC

RI PT

348

361

The transcript levels of the TBBPA-degrading gene in both strains were assayed using

362

qPCR. As shown in Fig. 3b, significant up-regulation of tbbpaA gene was observed following

363

the exposure to TBBPA (6 mg L-1) (p<0.05). After 24 h, the tbbpaA gene expression relative

364

to the housekeeping gene (16s rRNA) increased 2.2-fold and 1.6-fold for the Strain T and

365

constructed strain, respectively, as compared with the time 0 h. This up-regulation was likely

366

due to the presence of TBBPA in the mineral medium inducing the expression of tbbpaA gene. 17

ACCEPTED MANUSCRIPT Coincidentally, the trend of tbbpaA gene expression was similar with that of the tceA and

368

vcrA genes in Dehalococcoides spp, which were also up-regulated by trichloroethylene

369

exposure for 24 h (Xiu et al., 2010). Besides, this result also fitted well with the up-regulation

370

of the bvcA and vcrA gene expression upon exposed to vinyl chloride (Baelum et al., 2013).

RI PT

367

Compare to the tbbpaA gene expression of Strain T (6.6-fold up-regulation), a relative

372

higher degree of up-regulation was observed for constructed strain with maximum 10-fold

373

up-regulation within 72 h. These results were consistent with our abovementioned finding that

374

TBBPA removal efficiency was accelerated by the constructed strain (Fig. 2a). Whereas, the

375

expression of tbbpaA decreased to 5.7-fold and 4.2-fold for constructed strain and Strain T,

376

respectively, with further prolonging the biodegradation time to 96 h. One possible

377

explanation for this phenomenon is that, during this period, both bacterial strain began to

378

enter their anaphase of the stationary period or in prophase of the death period (Annweiler et

379

al., 2000). This result can also be confirmed by the decreased densities of both bacterial

380

strains at 96 h.

TE D

M AN U

SC

371

Overall, the above obtained results revealed that the key step of TBBPA removal by the

382

constructed bacteria is catalyzed by specific enzyme encoded by the tbbpaA gene in this work.

383

The enhanced biodegradation of TBBPA by constructed strain is due to the up-regulated

384

expression of the tbbpaA gene from 24 h to 72 h.

385

3.5. The effects of various additives on the aerobic TBBPA biodegradation

AC C

EP

381

386

Previous study showed that most microorganisms could obtain the redox power needed

387

for the contaminant metabolized through the electron transfer chain (Zu et al., 2014).

388

Therefore, it is worth exploring the effect of various carbon sources and electron donors, such 18

ACCEPTED MANUSCRIPT as ethanol, glucose, NH4NO3, sodium acetate and sodium citrate on the removal activity of

390

the constructed strain. As shown in Fig. 4a, compared with the inoculated control (91.1%),

391

higher aerobic degradation efficiencies of 100%, 93.1% and 91.9% were observed in the

392

system with addition of NH4NO3, ethanol and glucose, respectively. These were consistent

393

with a previous work that the addition of ethanol, pyruvate and glucose can enhance

394

biodegradation of TBBPA in sediment by halorespiring bacteria (Arbeli et al., 2006).

395

Therefore, we concluded that they might also be used directly by the TBBPA debromination

396

bacteria as the carbon and energy sources or electron donors. This result also agreed with the

397

report that the supply of NaNO3 and NH4Cl can effectively improve the degradation

398

efficiencies of TBBPA from 19.1% to 77.6% and 83.9%, respectively (Peng and Jia, 2013).

399

Nevertheless, the increase of degradation efficiencies was not so obvious with the addition of

400

glucose, indicating that no noticeable stimulation of the microbial activity existed in this

401

system. This property of the constructed strain was also compared with the original Strain T

402

(Zu et al., 2014), further confirming that the constructed strain retained some characteristics

403

of the Strain T. In contrast, the use of citric acid and acetic acid will greatly decrease the

404

biodegradation efficiency from 91.1% to 67.0 and 62.8% at 96 h, respectively, probably

405

because these two additives do not function as electron donors by the constructed strain

406

(Chang et al., 2012).

AC C

EP

TE D

M AN U

SC

RI PT

389

407

Based on our previous work, two main debromination intermediates, TBP and BPA were

408

produced during degradation of TBBPA by Strain T (An et al., 2011). Therefore, the

409

influence of these two compounds on the biodegradation efficiency was also investigated

410

when TBBPA was degraded by the constructed strain. As shown in Fig. 4b, when BPA or 19

ACCEPTED MANUSCRIPT TBP alone was subjected to the biodegradation systems of TBBPA by the constructed strain,

412

BPA could not be eliminated, and only 20.0% TBP was removed after 96 h reaction,

413

suggesting that TBP can slowly be used by the constructed strain. These results were different

414

from those of Strain T probably due to that the degradation of these organics by these two

415

strains was under the regulation of different enzymes (Zu et al., 2014). Furthermore, in

416

TBBPA+BPA system, the addition of BPA could obviously inhibit the TBBPA degradation

417

by constructed strain from degrading TBBPA (74.7% efficiency after 96 h). The inhibition

418

effect is due to the estrogen-like effect of BPA, a well-known estrogen disrupter, on the

419

constructed strain (Olsen et al., 2003; Uhnakova et al., 2011). In addition, a gradual decrease

420

trend of biodegradation efficiency in the TBBPA+BPA system is different from the results

421

obtained in original Strain T system, which decreased rapidly within first 36 h, then exhibited

422

an increase trend and finally declined at the stage of 36-60 h (Zu et al., 2014). This may

423

explain that no BPA could be produced during the biodegradation of TBBPA by constructed

424

strain. Different from the BPA+TBBPA system, the evolution curve in the TBBPA+TBP

425

system by the constructed strain can be described in two stages. First stage, within 24 h

426

inoculation, faster degradation efficiency of TBBPA was observed in TBBPA+TBP system

427

than that in TBBPA alone system; Second stage, after that, with the increase of the reaction

428

time from 24 to 96 h, the degradation efficiency is reversed due to the toxicity of the increase

429

concentration of TBP to the TBBPA biodegradation strain (Arbeli et al., 2006). As a result,

430

we conclude from the above different effects of various additives on the aerobic TBBPA

431

biodegradation between constructed strain and Strain T that, except for the enzyme encoding

AC C

EP

TE D

M AN U

SC

RI PT

411

20

ACCEPTED MANUSCRIPT 432

by the tbbpaA gene, there also existed other enzymes in the strain T associated with the fully

433

degradation of TBBPA.

434

4. Conclusions In sum, a novel TBBPA-degrading gene (tbbpaA) was cloned from the genomic DNA of

436

Strain T after draft genomic sequencing. Subsequently, this gene was successfully expressed

437

in the E. coli BL21 (DE3) to generate constructed strain, which exhibited faster

438

debromination and mineralization activity as well as more robust growth than Strain T with

439

TBBPA as the sole carbon and energy source. NH4NO3, ethanol and glucose addition

440

promoted the biodegradation of TBBPA, while the addition of citric acid and acetic acid

441

inhibited the biodegradation by this constructed strain. Much higher transcription expression

442

level of tbbpaA suggested that the up-regulated expression of tbbpaA would enhance the

443

TBBPA biodegradation. These results provide sound basis knowledge for future studies

444

aimed to better understand the molecular principles of TBBPA biodegradation. Meanwhile, it

445

also opens up possibilities for the potential application of bioremediation of TBBPA in situ

446

environments.

447

Acknowledgments

AC C

EP

TE D

M AN U

SC

RI PT

435

448

This work was supported by NSFC (41373103 and 41573086), National Natural Science

449

Funds for Distinguished Young Scholars (41425015), Science and Technology Program of

450

Guangzhou, China (201704020185), and Science and Technology Project of Guangdong

451

Province, China (2017A050506049).

452

Appendix A. Supplementary data 21

ACCEPTED MANUSCRIPT 453

Supplementary data related to this article can be found at

References

455

Ahn, Y.B., Rhee, S.K., Fennell, D.E., Kerkhof, L.J., Hentschel, U., Haggblom, M.M., 2003.

456

Reductive dehalogenation of brominated phenolic compounds by microorganisms

457

associated with the marine sponge Aplysina aerophoba. Appl. Environ. Microbiol. 69(7),

458

4159-4166.

RI PT

454

An, T.C., Zu, L., Li, G.Y., Wan, S.G., Mai, B.X., Wong, P.K., 2011. One-step process for

460

debromination and aerobic mineralization of tetrabromobisphenol-A by a novel

461

Ochrobactrum sp T isolated from an e-waste recycling site. Bioresource Technol,

462

102(19), 9148-9154.

M AN U

SC

459

Annweiler, E., Richnow, H.H., Antranikian, G., Hebenbrock, S., Garms, C., Franke, S.,

464

Francke, W., Michaelis, W., 2000. Naphthalene degradation and incorporation of

465

naphthalene-derived carbon into biomass by the thermophile Bacillus thermoleovorans.

466

Appl. Environ. Microbiol. 66(2), 518-523. Arbeli,

Z.,

EP

467

TE D

463

Ronen,

Z.,

Diaz-Baez,

M.C.,

2006.

Reductive

dehalogenation

of

tetrabromobisphenol-A by sediment from a contaminated ephemeral streambed and an

469

enrichment culture. Chemosphere 64(9), 1472-1478.

AC C

468

470

Arora, P.K. Bae, H. 2014. Role of dehalogenases in aerobic bacterial degradation of

471

chlorinated aromatic compounds. J. Chem. 2014, http://dx.doi.org/10.1155/2014/157974.

472

Atashgahi, S., Lu, Y., Smidt, H., 2016. Overview of Known Organohalide-Respiring

473

Bacteria—Phylogenetic Diversity and Environmental Distribution. Springer Berlin

474

Heidelberg. 22

ACCEPTED MANUSCRIPT Baelum, J., Chambon, J.C., Scheutz, C., Binning, P.J., Laier, T., Bjerg, P.L., Jacobsen, C.S.,

476

2013. A conceptual model linking functional gene expression and reductive

477

dechlorination rates of chlorinated ethenes in clay rich groundwater sediment. Water Res.

478

47(7), 2467-2478.

RI PT

475

479

Bisaillon, A., Beaudet, R., Lepine, F., Deziel, E., Villemur, R. 2010., Identification and

480

characterization of a novel CprA reductive dehalogenase specific to highly chlorinated

481

phenols

482

Microbiol.76(22), 7536-7540.

hafniense

strain

PCP-1.

SC

Desulfitobacterium

Appl.

Environ.

M AN U

from

Bustin, S.A., Benes, V., Garson, J.A., Hellemans, J., Huggett, J., Kubista, M., Mueller, R.,

484

Nolan, T., Pfaffl, M.W., Shipley, G.L., Vandesompele, J., Wittwer, C.T., 2009. The

485

MIQE guidelines: minimum information for publication of quantitative real-time PCR

486

experiments. Clin. Chem. 55(4), 611-622.

487 488

TE D

483

Chang, B.V., Yuan, S.Y., Ren, Y.L., 2012. Anaerobic degradation of tetrabromobisphenol-A in river sediment. Ecol. Eng. 49, 73-76.

Chen, K., Huang, L.L., Xu, C.F., Liu, X.M., He, J., Zinder, S.H., Li, S.P., Jiang, J.D., 2013.

490

Molecular characterization of the enzymes involved in the degradation of a brominated

491

aromatic herbicide. Mol. Microbiol. 89(6), 1121-1139.

AC C

EP

489

492

Chiba, Y., Yoshida, T., Ito, N., Nishimura, H., Imada, C., Yasuda, H., Sako, Y., 2009.

493

Isolation of a bacterium possessing a haloacid dehalogenase from a marine sediment

494

core. Microbes Environ. 24(3), 276-279.

23

ACCEPTED MANUSCRIPT Coconi-Linares, N., Ortiz-Vazquez, E., Fernandez, F., Loske, A.M., Gomez-Lim, M.A., 2015.

496

Recombinant expression of four oxidoreductases in Phanerochaete chrysosporium

497

improves degradation of phenolic and non-phenolic substrates. J. Biotechnol. 209, 76-84.

498

Ding, C. He, J.Z., 2012. Molecular techniques in the biotechnological fight against

499

RI PT

495

halogenated compounds in anoxic environments. Microbial Biotechnol. 5(3), 347-367. Fricker, A.D., LaRoe, S.L., Shea, M.E., Bedard, D.L., 2014. Dehalococcoides mccartyi Strain

501

JNA dechlorinates multiple chlorinated phenols including pentachlorophenol and harbors

502

at least 19 reductive dehalogenase homologous genes. Environ. Sci. Technol. 48(24),

503

14300-14308.

M AN U

SC

500

Guo, Y.G., Zhou, J., Lou, X.Y., Liu, R.L., Xiao, D.X., Fang, C.L., Wang, Z.H., Liu, J.S.,

505

2014. Enhanced degradation of Tetrabromobisphenol A in water by a UV/base/persulfate

506

system: Kinetics and intermediates. Che. Eng. J. 254, 538-544.

TE D

504

Heinaru, E., Naanuri, E., Grunbach, M., Joesaar, M., Heinaru, A., 2016. Functional

508

redundancy in phenol and toluene degradation in Pseudomonas stutzeri strains isolated

509

from the Baltic Sea. Gene 589(1), 90-98.

EP

507

Horikoshi, S., Miura, T., Kajitani, M., Horikoshi, N.. Serpone, N., 2008. Photodegradation of

511

tetrahalobisphenol-A (X = Cl, Br) flame retardants and delineation of factors affecting

512

the process. App. Catal. B-Environ. 84(3-4), 797-802.

AC C

510

513

Hug, L.A., Maphosa, F., Leys, D., Loffler, F.E., Smidt, H., Edwards, E.A., Adrian, L., 2013.

514

Overview of organohalide-respiring bacteria and a proposal for a classification system

515

for reductive dehalogenases. Philos. Trans. R. Soci. B-Biol. Sci. 368(1616), 20120322,

516

doi:10.1098/rstb.2012.0322. 24

ACCEPTED MANUSCRIPT Johnson, D.R., D.R., Lee, P.K.H., Holmes, V.F., Alvarez-Cohen, L., 2005. An internal

518

reference technique for accurately quantifying specific mRNAs by real-time PCR with a

519

application to the tceA reductive dehalogenase gene. Appl. Environ. Microbiol. 71(7),

520

3866-3871.

RI PT

517

Kube, M., Beck, A., Zinder, S.H., Kuhl, H., Reinhardt, R., Adrian, L., 2005. Genome

522

sequence of the chlorinated compound respiring bacterium Dehalococcoides species

523

strain CBDB1. Nat. Biotechnol. 23(10), 1269-1273.

SC

521

Li, F., Zhu, L.Z., Wang, L.W., Zhan, Y., 2015a. Gene expression of an arthrobacter in

525

surfactant-enhanced biodegradation of a hydrophobic organic compound. Environ. Sci.

526

Technol. 49(6), 3698-3704.

M AN U

524

Li, F.J., Jiang, B.Q., Nastold, P., Kolvenbach, B.A., Chen, J.Q., Wang, L.H., Guo, H.Y.,

528

Corvini, P.F.X., Ji, R., 2015b. Enhanced transformation of tetrabromobisphenol A by

529

nitrifiers in nitrifying activated sludge. Environ. Sci. Technol. 49(7), 4283-4292.

TE D

527

Li, G.Y., Xiong, J.K., Wong, P.K., An, T.C., 2016. Enhancing tetrabromobisphenol A

531

biodegradation in river sediment microcosms and understanding the corresponding

532

microbial community. Environ. Pollut. 208, 796-802.

AC C

EP

530

533

Liang, X.M., Molenda, O., Tang, S.Q., Edwards, E.A., 2015. Identity and substrate specificity

534

of reductive dehalogenases expressed in dehalococcoides-containing enrichment cultures

535

maintained on different chlorinated ethenes. Appl. Environ. Microbiol. 81(14),

536

4626-4633.

25

ACCEPTED MANUSCRIPT 537

Liang, Z.S., Li, G.Y., An, T.C., Zhang, G.X., Das, R. 2016. Draft genome sequence of a

538

tetrabromobisphenol a-degrading strain, Ochrobactrum sp. T, isolated from an electronic

539

waste recycling site. Genome Announc. 4(4), DOI: 10.1128/genomeA.00680-16. Liu, J., Wang, Y.F., Jiang, B.Q., Wang, L.H., Chen, J.Q., Guo, H.Y., Ji, R., 2013. Degradation,

542

tetrabromobisphenol a in soil during sequential anoxic-oxic incubation. Environ. Sci.

543

Technol. 47(15), 8348-8354.

bound-residue

formation

and

release

of

Livak, K.J. Schmittgen, T.D., 2001. Analysis of relative gene expression data using real-time

M AN U

545

and

SC

541

544

metabolism,

RI PT

540

quantitative PCR and the 2 (-Delta Delta C(T)) method. Methods 25(4), 402-408. Michael., G., Joseph., S., 2012. Molecular Cloning: A Laboratory Manual Fourth ed.

547

Nicolini, C., Bruzzese, D., Cambria, M.T., Bragazzi, N.L., Pechkova, E., 2013. Recombinant

548

Laccase: I. Enzyme cloning and characterization. J. Cell. Biochem. 114(3), 599-605.

549

Nijenhuis, I. Kuntze, K., 2016. Anaerobic microbial dehalogenation of organohalides - state

550

TE D

546

of the art and remediation strategies. Curr. Opin. Biotechnol. 38, 33-38. Olsen, C.M., Meussen-Elholm, E.T.M., Samuelsen, M., Holme, J.A., Hongslo, J.K., 2003.

552

Effects of the environmental oestrogens bisphenol A, tetrachlorobisphenol A,

553

tetrabromobisphenol A, 4-hydroxybiphenyl and 4,4 '-dihydroxybiphenyl on oestrogen

554

receptor binding, cell proliferation and regulation of oestrogen sensitive proteins in the

555

human breast cancer cell line MCF-7. Pharmacol. Toxicol. 92(4), 180-188.

AC C

EP

551

556

Padilla-Crespo, E., Yan, J., Swift, C., Wagner, D.D., Chourey, K., Hettich, R.L., Ritalahti,

557

K.M., Loffler, F.E., 2014. Identification and environmental distribution of dcpA, which

558

encodes

the

reductive

dehalogenase

catalyzing

the

dichloroelimination

of 26

ACCEPTED MANUSCRIPT 559

1,2-dichloropropane to propene in organohalide-respiring chloroflexi. Appl. Environ.

560

Microbiol. 80(3), 808-818.

562

Peng, X.X. Jia, X.S., 2013. Optimization of parameters for anaerobic co-metabolic degradation of TBBPA. Bioresource Technol. 148, 386-393.

RI PT

561

Peng, X.X., Wang, Z.N., Huang, J.F., Pittendrigh, B.R., Liu, S.W., Jia, X.S., Wong, P.K.,

564

2017. Efficient degradation of tetrabromobisphenol A by synergistic integration of Fe/Ni

565

bimetallic catalysis and microbial acclimation. Water Res. 122, 471-480.

SC

563

Peng, X.X., Zhang, Z.L., Luo, W.S., Jia, X.S., 2013. Biodegradation of tetrabromobisphenol

567

A by a novel Comamonas sp strain, JXS-2-02, isolated from anaerobic sludge.

568

Bioresource Technol. 128, 173-179.

570

Richardson, R.E. 2013. Genomic insights into organohalide respiration. Curr. Opin. Biotechnol. 24(3), 498-505.

TE D

569

M AN U

566

Richardson, R.E., 2016. Organohalide-Respiring Bacteria as Members of Microbial

572

Communities: Catabolic Food Webs and Biochemical Interactions. Springer Berlin

573

Heidelberg.

575

Ronen, Z. Abeliovich, A., 2000. Anaerobic-aerobic process for microbial degradation of

AC C

574

EP

571

tetrabromobisphenol A. Appl. Environ. Microbiol. 66(6), 2372-2377.

576

Seshadri, R., Adrian, L., Fouts, D.E., Eisen, J.A., Phillippy, A.M., Methe, B.A., Ward, N.L.,

577

Nelson, W.C., Deboy, R.T., Khouri, H.M., Kolonay, J.F., Dodson, R.J., Daugherty, S.C.,

578

Brinkac, L.M., Sullivan, S.A., Madupu, R., Nelson, K.T., Kang, K.H., Impraim, M.,

579

Tran, K., Robinson, J.M., Forberger, H.A., Fraser, C.M., Zinder, S.H., Heidelberg, J.F.,

27

ACCEPTED MANUSCRIPT 580

2005. Genome sequence of the PCE-dechlorinating bacterium Dehalococcoides

581

ethenogenes. Science 307(5706), 105-108. Stasinakis, A.S., Kordoutis, C.I., Tsiouma, V.C., Gatidou, G., Thomaidis, N.S., 2010.

583

Removal of selected endocrine disrupters in activated sludge systems: Effect of sludge

584

retention time on their sorption and biodegradation. Bioresource Technol. 101(7),

585

2090-2095.

RI PT

582

Sun, F.F., Kolvenbach, B.A., Nastold, P., Jiang, B.Q., Ji, R., Corvini, P.F.X., 2014.

587

Degradation and metabolism of tetrabromobisphenol A (TBBPA) in submerged soil and

588

soil-plant systems. Environ. Sci. Technol. 48(24), 14291-14299.

M AN U

SC

586

Tas, N., van Eekert, M.H.A., de Vos, W.M., Smidt, H., 2010. The little bacteria that can -

590

diversity, genomics and ecophysiology of 'Dehalococcoides' spp. in contaminated

591

environments. Microb. Biotechnol. 3(4), 389-402.

TE D

589

Uhnakova, B., Ludwig, R., Peknicova, J., Homolka, L., Lisa, L., Sulc, M., Petrickova, A.,

593

Elzeinova, F., Pelantova, H., Monti, D., Kren, V., Haltrich, D., Martinkova, L., 2011.

594

Biodegradation of tetrabromobisphenol A by oxidases in basidiomycetous fungi and

595

estrogenic activity of the biotransformation products. Bioresource Technol. 102(20),

596

9409-9415.

AC C

EP

592

597

Voordeckers, J.W., Fennell, D.E., Jones, K., Haggblom, M.M. 2002. Anaerobic

598

biotransformation of tetrabromobisphenol A, tetrachlorobisphenol A, and bisphenol A in

599

estuarine sediments. Environ. Sci. Technol. 36(4), 696-701.

28

ACCEPTED MANUSCRIPT 600

Wang, S., Chng, K.R., Chen, C., Bedard, D.L., He, J., 2015a. Genomic characterization of

601

Dehalococcoides mccartyi strain JNA that reductively dechlorinates tetrachloroethene

602

and polychlorinated biphenyls. Environ. Sci. Technol. 49 (24), 14319-14325. Wang, X.W., Hu, X.F., Zhang, H., Chang, F., Luo, Y.M., 2015b. Photolysis kinetics,

604

mechanisms, and pathways of tetrabromobisphenol a in water under simulated solar light

605

irradiation. Environ. Sci. Technol. 49(11), 6683-6690.

RI PT

603

Xiong, J.K., Li, G.Y., An, T.C., 2015. Development of methodology for the determination of

607

carbon isotope ratios using gas chromatography/combustion/isotope ratio mass

608

spectrometry and applications in the biodegradation of phenolic brominated flame

609

retardants and their degradation products. Rapid Commun. Mass Sp. 29(1), 54-60.

M AN U

SC

606

Xiong, L., Kameshwar, A.K., Chen, X., Guo, Z., Mao, C., Chen, S., Qin, W. 2016. The

611

ACEII recombinant Trichoderma reesei QM9414 strains with enhanced xylanase

612

production and its applications in production of xylitol from tree barks. Microb. Cell

613

Fact. 15(1), 215.

TE D

610

Xiu, Z.M., Gregory, K.B., Lowry, G.V., Alvarez, P.J.J., 2010. Effect of bare and coated

615

nanoscale zerovalent iron on tceA and vcrA gene expression in Dehalococcoides spp.

616

Environ. Sci. Technol. 44(19), 7647-7651.

AC C

EP

614

617

Yang, C., Kublik, A., Weidauer, C., Seiwert, B., Adrian, L. 2015. Reductive dehalogenation

618

of oligocyclic phenolic bromoaromatics by Dehalococcoides mccartyi Strain CBDB1.

619

Environ. Sci. Technol. 49(14), 8497-8505.

29

ACCEPTED MANUSCRIPT 620

Yu, H., Tang, H.Z., Zhu, X.Y., Li, Y.Y., Xu, P., 2015. Molecular mechanism of nicotine

621

degradation by a newly isolated strain, Ochrobactrum sp strain SJY1. Appl. Environ.

622

Microbiol. 81(1), 272-281. Zhang, C., Zeng, G.M., Yuan, L., Yu, J., Li, J.B., Huang, G.H., Xi, B.D., Liu, H.L., 2007.

624

Aerobic degradation of bisphenol A by Achromobacter xylosoxidans strain B-16 isolated

625

from compost leachate of municipal solid waste. Chemosphere 68(1), 181-190.

RI PT

623

Zhang, C.F., Li, Z.L., Suzuki, D., Ye, L.Z., Yoshida, N., Katayama, A., 2013. A

627

humin-dependent Dehalobacter species is involved in reductive debromination of

628

tetrabromobisphenol A. Chemosphere 92(10), 1343-1348.

M AN U

SC

626

Zhou, N.A., Kjeldal, H., Gough, H.L., Nielsen, J.L. 2015. Identification of putative genes

630

involved in bisphenol a degradation using differential protein abundance analysis of

631

Sphingobium sp BiD32. Environ. Sci. Technol. 49(20), 12232-12241.

TE D

629

Zu, L., Li, G.Y., An, T.C., Wong, P.K., 2012. Biodegradation kinetics and mechanism of

633

2,4,6-tribromophenol by Bacillus sp GZT: A phenomenon of xenobiotic methylation

634

during debromination. Bioresource Technol, 110, 153-159.

EP

632

Zu, L., Xiong, J.K., Li, G.Y., Fang, Y.J., An, T.C. 2014. Concurrent degradation of

636

tetrabromobisphenol A by Ochrobactrum sp T under aerobic condition and estrogenic

637

transition during these processes. Ecotox. Environ. Safe. 104, 220-225.

638

AC C

635

30

ACCEPTED MANUSCRIPT Table 1. Primers for quantitative real-time polymerase chain reaction

Gene

Genbank number

16sRNA

HM5431 85.1

tbbpaA

KY 483638

Gene Primer sequence (5’→3’) Forward Reverse Forward

GAGAGGAAGGTGGGGATGAC CGAACTGAGATGGCTTTTGG GCTTCAGAAACCGATGCTCT

Reverse

CCTCTTCAAACGCTTTCCAC

Product length (bp)

Efficiency (%)

131

99.6

59.85

118

93

AC C

EP

TE D

M AN U

SC

640

Annealing temperature (°C) 60.86 60.77 59.58

RI PT

639

31

ACCEPTED MANUSCRIPT 641

Figure captions:

642

Fig. 1 Phylogenetic tree of bromophenol dehalogenase (ORF 04638) from Strain T and other related dehalogenases based on amino acid sequences by the neighbor-joining method on the program

644

MEGA5.2. The numbers shown next to the nodes indicated the bootstrap values of 1000 in

645

percentage.

646

RI PT

643

Fig. 2 Comparison of (a) temporal decrease of TBBPA, (b) formation of bromide and (c) loss of total organic carbon (TOC) during the course of biodegradation of TBBPA by

648

constructed strain with other bacteria. Error bars represent standard deviations of

649

triplicate cultures.

M AN U

SC

647

Fig. 3 The growth (a) and relative tbbpaA gene expression fold changes (b) of constructed

651

strain and Strain T after the exposure to TBBPA. Fold changes of target genes for

652

each time were normalized to initial conditions (time 0 h). The 16s rRNA was used as

653

the reference gene. All data points represent average values from triplicate samples,

654

and error bars represent one standard deviation.

657

EP

656

Fig. 4 The addition of different energy sources (a), TBP and BPA (b) on the biodegradation efficiency of TBBPA by constructed strain. Dates are the means of triplicates.

AC C

655

TE D

650

32

ACCEPTED MANUSCRIPT 658

Fig. 1

659 84 100

Haloacid dehalogenase Ochrobactrum sp. EGD-AQ16 (WP 021588323) Haloacid dehalogenase Ochrobactrum intermedium (WP 025091885) Bromophenol dehalogenase Ochrobactrum sp. T

RI PT

Haloacid dehalogenase Ochrobactrum anthropic (WP 061347506)

Haloacid dehalogenase Ochrobactrum rhizosphaerae (WP 024899099)

100

Haloacid dehalogenase Ochrobactrum anthropic (WP 036587187) 96

Haloacid dehalogenase Ochrobactrum anthropic (WP 040128237)

0.01

SC

660

AC C

EP

TE D

M AN U

661

33

ACCEPTED MANUSCRIPT 662

Fig. 2 100 Ochrobactrum sp. T pET-30a(+)-a6 pET-30a(+)

C/C0

75 50

(a)

0

24

48

96

(b)

SC

20 10

0

24

(c) 30 20

48

72

96

48 Time (h)

72

96

TE D

10

M AN U

0 40

TOC removal (%)

72

30

-

Br concentration (µΜ )

0 40

RI PT

25

0

664

AC C

663

24

EP

0

34

ACCEPTED MANUSCRIPT Fig. 3 (a)

7

3x10

7

2x10

7

1x10

0

0

24

48 72 Time (h)

666

96

(b)

0.9 0.6 0.3 0.0 0

24

48 72 Time (h)

96

AC C

EP

TE D

M AN U

SC

667

1.2

RI PT

Constructed strain Strain T

-1

Cell concentration (cfu mL )

7

4x10

Log10 (gene expression fold change)

665

35

ACCEPTED MANUSCRIPT Fig. 4

C/C0 (%)

100 75

75

50 25 0

50 Glucose Ethanol NH4NO3

24

TBP BPA TBBPA TBBPA+BPA TBBPA+TBP

25 0 48

Time (h)

72

96

(b)

0

24

48

Time (h)

72

96

AC C

EP

TE D

M AN U

SC

(a) 0

669

100

Acetic acid Citric acid TBBPA control

RI PT

668

36

ACCEPTED MANUSCRIPT

Highlights
Gene tbbpaA encoding TBBPA degrading enzyme was cloned and overexpressed
Higher bromophenol dehalogenase activity was achieved in the constructed strain

AC C

EP

TE D

M AN U

SC

RI PT


The expression level of tbbpaA gene was upregulated upon the exposure to TBBPA