Aerobic Soil Biodegradation of Bisphenol (BPA) - ACS Publications

Oct 20, 2017 - Purdue University, Department of Agronomy, Ecological Science and ... Pressures to ban bisphenol A (BPA) has led to the use of alternat...
2 downloads 4 Views 1MB Size
Subscriber access provided by Gothenburg University Library

Article

Aerobic Soil Biodegradation of Bisphenol (BPA) Alternatives Bisphenol S and Bisphenol BPAF Compared to BPA Youn Jeong Choi, and Linda S Lee Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b03889 • Publication Date (Web): 07 Nov 2017 Downloaded from http://pubs.acs.org on November 8, 2017

Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a free service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are accessible to all readers and citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.

Environmental Science & Technology is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.

Page 1 of 22

Environmental Science & Technology

1 2 3

Aerobic Soil Biodegradation of Bisphenol (BPA) Alternatives Bisphenol S and

4

Bisphenol AF Compared to BPA

5

Youn Jeong Choi and Linda S. Lee*

6

Purdue University, Department of Agronomy, Ecological Science and Engineering

7

Interdisciplinary Graduate Program, West Lafayette, IN 47907-2054

8 9 10 11 12

*

Corresponding author at: Department of Agronomy Purdue University, West Lafayette, IN 47907, USA. Tel.: +1 765 494 8612; fax: +1 765 496 2926. E-mail address: [email protected] (L.S. Lee)

13 14 15 16 17

Prepared for Environmental Science & Technology

October 20, 2017

18 19 20

1 ACS Paragon Plus Environment

Environmental Science & Technology

21

Graphical Abstract

22

23

2 ACS Paragon Plus Environment

Page 2 of 22

Page 3 of 22

24 25

Environmental Science & Technology

ABSTRACT Pressures to ban bisphenol A (BPA) has led to the use of alternate chemicals such as BPA

26

analogues bisphenol S and bisphenol AF in production of consumer products; however,

27

information on their environmental fate is scarce. In this study, aerobic degradation of BPA,

28

BPAF and BPS at 100 μg/kg soil and 22  2 ℃ was monitored for up to 180 d in a forest soil and

29

an organic farm soil. At each sampling point, soils were extracted three times and analyzed by

30

liquid chromatography high resolution mass or time of flight mass spectrometry. Based on

31

compound mass recovered from soils compared to the mass applied, BPS had short half-lives of

32

< 1 d in both soils similar to BPA. BPAF was much more persistent with observed half-lives of

33

32.6 and 24.5 days in forest and farm soils, respectively. To our knowledge, this is the first report

34

on BPAF degradation. For all three compounds, half-lives were longer in the higher organic

35

carbon (OC) forest soil which correlates well to sorption studies showing higher sorption with

36

higher OC. Metabolites identified for all three bisphenols support degradation pathways that

37

include meta-cleavage as well as ortho-cleavage, which has not been previously shown.

38 39

Keywords. microbial degradation, metabolite identification, ortho-cleavage, meta-cleavage.

40

kinetics

3 ACS Paragon Plus Environment

Environmental Science & Technology

41 42

INTRODUCTION Bisphenol A (BPA) has been in commercial use since 1957 for the production of a

43

variety of products and as a pre-cursor in the synthesis of other chemicals. Its greatest use has

44

been in the production of polycarbonate, which aids in making material both structurally strong

45

and transparent at a low cost.1 However, increased concerns regarding its estrogen disrupting

46

activity and the discovery that BPA leaches out of products such as baby bottles2,3 and food and

47

beverage containers,4–7 the public began to demand BPA-free products. In the past decade, BPA

48

sales or use have been banned by many countries and several cities within the U.S.A. including a

49

complete ban in the U.S.A. of use in baby bottles and children’s cups in 2012,8 which increased

50

production of alternatives. However, some BPA alternatives such as BPS and BPAF which have

51

widespread consumer and commercial use, are presence in various media, share a similar

52

structure to BPA and also exemplify endocrine disrupting characteristics.9,10

53

BPA and its alternatives may enter into the environment in several ways. One major

54

source is wastewater treatment plants (WWTPs) 11 through direct effluent discharge12,13 or land

55

application of biosolids from processed sludges.14,15 Even though WWTP processes are designed

56

as a clean-up process of municipal or industrial wastes, many compounds are not completely

57

removed in the process. Recent studies revealed the presence of BPA and its substitutes in

58

sediments and water near industrial sites and WWTP discharge points. For example, BPAF,

59

BPS, and BPA were found at concentrations as high as 2010 ng/g, 1970 ng/g, and 13,370 ng/g in

60

sediment, respectively, and as high as 246 ng/L, 20 ng/L, and 75 ng/L in river water,

61

respectively.16,17 In addition, removal of hydrophobic recalcitrant chemicals from the wastewater

62

stream is often simply due to sorption to the sludge solids, which in turn can be land applied for

63

their nutrient value. Nie et al. (2009) reported BPA concentrations in WWTP solids of 101-127

4 ACS Paragon Plus Environment

Page 4 of 22

Page 5 of 22

Environmental Science & Technology

64

g/kg, which was at least 700 times greater than found in the liquid phase.18 Choi et al.19 also

65

found high sorption of BPAF and moderate sorption of BPA sorption anaerobic sludge solids

66

maintained under methanogenic conditions where no degradation of BPs was observed.

67

BPA has been shown to readily degrade aerobically in water, sediment, and soils through

68

microbial processes20–23 with reported aerobic half-lives (t1/2) of 0.6~8 d in river water and river

69

sediments21,22,24–28 and 0.81~7 d in soil.23,29,30 In the marine environment, BPA persistence

70

increased with t1/2 values in the 4-20 d range23,29,31–33 and the longest in sea water due to a 30-d

71

lag phase.26 While ample research on BPA degradation exists in various media, data are limited

72

or nonexistent for the BPA alternatives. For example, no data on degradation in environmental

73

media exists for BPAF. For BPS, only two studies, which were with river water and sea water,

74

have been done for which no aerobic degradation was observed.27,33 Biodegradation is an

75

important fate process controlling the persistence of organic contaminants once released into the

76

soil environment such as in the case of effluent irrigation and land application of biosolids. This

77

study focused on quantifying the aerobic soil biodegradation of BPS and BPAF relative to BPA

78

for which data are lacking or non-existent.

79 80 81

MATERIALS AND METHODS Soils. Two surface clay loam soils were used in this study: one sampled in a forested area

82

close to the Purdue campus (FRST-50) and one sampled from the Purdue Student Organic Farm

83

(PSF-51). Although both soils are clay loams, they were impacted by different land uses as well

84

as the forest soil is more acidic (by 0.4 pH units), and has almost twice the organic carbon (OC)

85

content and cation exchange capacity (CEC). Differences in land use may also have led to

86

different microbial consortia, which can impact degradation; however, microbial community

5 ACS Paragon Plus Environment

Environmental Science & Technology

87

analysis was not included in this study. Selected soil properties are detailed in Table 1. Soils

88

were moist sieved (2-mm maximum particle size), stored at 4 ℃ prior to use, and degradation

89

studies were initiated within 3 months of sampling.

90

Chemicals. BPA {4,4'-(propane-2,2-diyl)diphenol}, BPAF {4-[1,1,1,3,3,3-Hexafluoro-2-

91

(4 hydroxyphenyl) propan-2-yl]phenol}, and BPS {4,4'-Sulfonyldiphenol} were obtained from

92

Sigma Chemicals, St. Louis MO, USA and stored at room temperature (See SI for

93

physicochemical properties of target chemicals, Table S1). Deuterated BPA (d8-BPA) for use as

94

an internal standard was purchased from Cambridge chemicals. Acetonitrile (ACN) and

95

methanol (MeOH) were purchased as >99% purity, HPLC grade from Sigma-Aldrich and

96

Macron, respectively. Stock solutions of target chemicals were prepared in pure methanol and

97

stored at 4℃ individually. Talc used as a compound carrier was purchased from Mallinckrodt.

98 99

Degradation studies. Aerobic biodegradation studies were conducted using methods similar to those described by Mashtare et al.34 Briefly, soil (10 g air-dried weight) was added to

100

sterile 125-mL amber glass serum bottles capped with butyl rubber aluminum crimp caps,

101

adjusted to approximately 75 % of field capacity (Table 1) using sterile water, and pre-incubated

102

for 5 days at 22  2 ℃ to establish a steady-state microbial activity.35 Soils for sterile controls

103

were autoclave-sterilized using a method similar to that described by Wolf et al.36 After pre-

104

incubation, a set of soil microcosms were autoclaved (hereafter referred to as autoclave-sterilized

105

controls) three times at 103.4 KPa and 121 °C for 2 h on day 1, 2 and 4. All glassware and

106

deionized water were also autoclave-sterilized.

107

All experiments were conducted with individual bisphenol chemicals (BPs) in triplicate

108

for microbially active systems and in duplicate for autoclave-sterilized controls. Individual target

109

compounds were added to soil microcosms through a talc-carrier to target an initial soil

6 ACS Paragon Plus Environment

Page 6 of 22

Page 7 of 22

Environmental Science & Technology

110

concentration of 100 g/kg. BP-coated talc was prepared by mixing 10 ml of an individual

111

chemical stock solution (10 mg/L) dissolved in MeOH with 10 g of talc in a petri dish followed

112

by evaporating MeOH and homogenizing dry BP-coated talc. Single compound amended talc

113

(100 mg) was added to each microcosm resulting in the mass of talc not exceeding 1% of the soil

114

weight. In previous studies, no significant influence of the talc on chemical degradation

115

including phenolic-based compounds37,38 or differences compared to using ethanol as the target

116

compound carrier.39 Talc was considered to allow for a more even distribution in the soil, thus is

117

often selected as the carrier of choice for low solubility compounds. Compound concentrations

118

were monitored for 180 d with sampling times selected based on expected degradation patterns

119

and adjusted accordingly depending on observed degradation trends. Headspace O2 and CO2

120

levels were measured by sampling 5-mL of headspace in a subset of microcosms using a

121

monoject 6-mL needle syringe at designated incubation times to confirm aerobic conditions and

122

biological activity were maintained.. Headspace samples were injected directly onto an Agilent

123

7890A gas chromatograph (GC) equipped with a thermal conductivity detector (TCD).

124

At each sampling time, triplicate microcosms were extracted three times sequentially

125

with 25-ml of MeOH each time. After each extraction, bottles were equilibrated end-over-end at

126

35 rpm for ∼24 h at room temperature (22 ± 2 °C), and centrifuged at 1700 rpm for 60 min.

127

Aliquots (1 ml) of individual extracts were added to an HPLC vial and d8-BPA (0.5 ml) was

128

added uniformly to all vials.

129

HPLC-MS/MS Analysis. BPA, BPAF and BPS were quantified using a Shimadzu liquid

130

chromatography system (HPLC) system coupled to an Applied Biosystems Sciex API3000

131

tandem mass spectrometer (MS/MS). Data were acquired using the negative ion multiple

132

reaction monitoring (MRM) mode. Chromatographic separation was performed on a Kinetex

7 ACS Paragon Plus Environment

Environmental Science & Technology

133

C18 column (100 × 2.0 mm, dp-5 μm) using a 15-μL injection volume and an 80/20 v/v

134

MeOH/0.15% acetic acid phase at 0.3 mL/min using. Retention times and MS/MS conditions

135

summarized in Table S2.

136

Isotopically labeled internal standards and external calibration curves were used to

137

quantify all three BPs. Extract subsamples (1 ml) were transferred to an HPLC vial and 0.5-ml of

138

d8-BPA in MeOH was added to a final concentration of 150 g/L. Isotopic dilution coupled to

139

external calibration curves of no less than 5 concentrations were used to quantify target

140

compounds. The limit of detection (LOD) was considered three times the signal-to-noise ratio

141

(S/N) yielding LOD values of 1.70, 0.05, and 0.48 g/L for BPA, BPAF and BPS, respectively.

142

Limits of quantitation (LOQs) were designated as 10×S/N. The resulting method limits of

143

quantification (MLOQ) calculated from the quantifiable measured concentrations in soil extracts

144

were 19.2, 0.60, and 5.44 g/kg soil for BPA, BPAF and BPS, respectively. Isotopic mass-

145

labeled compounds were not available for BPS or BPAF; therefore, to confirm that using d8-

146

BPA to correct for matrix effects was adequate or matrix effects were minimal, additional tests

147

were done from which no significant matrix effects were observed in either soil (see SI for

148

additional details).

149

TOF analysis for metabolites. Metabolite analysis was done on a Shimadzu Ultra High

150

Pressure liquid chromatography system (uHPLC) system coupled to an applied Biosystems Sciex

151

API5600 Triple TOF. Data were acquired by non-targeted screening with an electrospray source

152

operating in negative ionization mode using Information Dependent Acquisition (IDA) MS/MS

153

spectra. uHPLC-TOF MS conditions for screening metabolites are summarized in Table S3.

154

Samples were injected (50 μL) onto a Kinetex C18 column (100 × 2.0 mm, dp-5 μm) and eluted

155

with a gradient of 0.15% acetic acid water and 20 mM ammonium acetate in methanol (see Fig.

8 ACS Paragon Plus Environment

Page 8 of 22

Page 9 of 22

Environmental Science & Technology

156

S1 in SI for additional details). Data was processed using PeakView software and MultiQuant

157

software. Using non-target comparative screening, only peaks with m/z observed in the BP-

158

amended soils and completely absent in the soil controls were used to come up with a list of

159

molecular masses for potential metabolites that may be specific to BP degradation. This list of

160

exact masses were compared to the masses identified in the MS data. For any match between the

161

expected and observed masses, MS/MS data were analyzed by calculating and comparing

162

expected fragmentation of the metabolites in the proposed list with the collected observed

163

MS/MS fragmentation patterns (see Fig. S7 for work flow scheme).

164

Kinetic analysis. Half-lives and degradation rates were obtained using the Computer

165

Aided Kinetic Evaluation (CAKE) R-based software. Data were fitted using both a Single First

166

Order (SFO) kinetic model, and a bi-exponential model referred to as the Double First Order in

167

Parallel (DFOP) kinetic model for comparison. Additional details are provided in the SI.

168 169 170

RESULTS & DISCUSSION Recoveries and sterile controls. Recoveries were 90.4 ± 0.6 % and 101.6 ± 5.7 % for

171

BPA, 104.4 ± 7.3% and 110.1 ± 11.8 % for BPAF, and 89.1 ± 8.7 % and 89.5 ± 10.2 % for BPS

172

in FRST and PSF soils, respectively. No statistically significant changes in BPA, BPAF and BPS

173

concentrations were observed over time in autoclave-sterilized soils (Fig. S2). Recoveries

174

averaged over time for each compound in the autoclave-sterilized FRST and PSF soils are as

175

follows: 95.6 ± 18.3% and 96.8 ± 11.4 % for BPA; 95.3 ± 17.1 % and 91.6 ± 16.1 % for BPAF;

176

and 92.9 ± 16.5 % and 102.5 ± 18.0 % for BPS, respectively. BPA, BPS and BPAF were not

177

detected in either soil prior to addition of the bisphenols in the degradation studies.

9 ACS Paragon Plus Environment

Environmental Science & Technology

178

Degradation in live soils. Oxygen content in soil microcosms remained ≥ 85% of

179

ambient O2 levels for the first month. In live soils, % CO2 increased continuously from 0.6 to

180

6.1 % indicating active aerobic microbial degradation. Soil incubation of BPA and BPS was

181

complete within one month whereas a 3-month period was needed for BPAF to be degraded to

182

levels approaching MLOQ (Figure 1). For the longer BPAF study, bottles were aerated monthly

183

to ensure adequate O2 levels.

184

The degradation profiles for BPA, BPS, and BPAF are shown in Figure 1 along with fits

185

to the SFO (solid line) and DFOP (dashed line) kinetic models. Model outputs for both kinetic

186

models are summarized in Table 2 along with observed half-lives (t1/2). Generally, the DFOP

187

model provided better fits overall (R2 values 0.88 to 0.96) and resulted in t1/2 values that agreed

188

better with observed values; however, reasonable fits also resulted from the simpler SFO model

189

(R2 values 0.79 to 0.96) except for BPA. Degradation of all three BPs followed similar trends

190

between soils (Figure 1) with BPA and BPS degrading faster than BPAF and consistently slower

191

degradation observed in the FRST soil. For BPA and BPS, ≥ 50% degraded in less than one day

192

in both soils whereas observed BPAF half-lives (t1/2) were 24.5 d and 32.6 d in the farm and

193

forest soils, respectively (Table 2, Figure 1). BPA concentrations fell below MLOQ by day 3

194

(open circles in Figure 1) and below LOD by day 11 (squares in Figure 1). BPS concentrations

195

were below MLOQ by day 11. For BPAF, concentrations remained above the MLOQ throughout

196

the 180-d incubation period. BPA half-lives in this study are similar to those previously reported

197

for aerobic soils of ≤ 3.3 days 29,40, 0.81 to 5.50 days30 and 7 days23.

198

The higher organic carbon (OC) in the forest soil of 2.7% versus 1.5 % in the farm soil

199

may have reduced bioavailability to microbes resulting in slower degradation in the forest soil.

200

Sorption of organic compounds is often associated with lower bioavailability, thus slower

10 ACS Paragon Plus Environment

Page 10 of 22

Page 11 of 22

Environmental Science & Technology

201

degradation and longer half-lives.41,42 OC-dependent sorption of BPAF, BPS, and BPA was

202

shown by Choi and Lee (2017)43 across 4 soils with average measured log OC-normalized-

203

sorption coefficients (log Koc) of 3.44 ± 0.28, 2.84 ± 0.28, and 2.57 ± 0.10, respectively. The

204

order of log Koc across these three BPs is also inversely correlated to the observed t1/2 values

205

adding support to the role of sorption on bioavailability, thus degradation rates and t1/2 values.

206

Ionizable compounds such as the BPs are subject to pH-dependent sorption as well; however, all

207

three BPs are essentially neutral in the pH range of the soils studied. Differences in degradation

208

between these two soils were expected given that the soils were sampled from lands with

209

distinctly different uses, which is expected to lead to different microbial communities although

210

not quantified in this study.

211

Metabolite Identification. Metabolite identification using LC/QTOF MS was conducted

212

by matching the observed MS and MS/MS spectrum, precursor ion spectra, and accurate mass

213

measurement of the observed transformation products (detailed in the SI, Fig. S7). Schymanski

214

et al. (2014) proposed levels of certainty in metabolite identification ranging from a structure

215

confirmed using MS/MS and a reference standard (highest level refer to as ‘Level 1’) to only

216

getting an exact mass of interest using MS (lowest level refer to as ‘Level 5’).44 Metabolites

217

identified from degradation of BPA, BPS and BPAF by aerobic soil microbes in samples from at

218

least one sampling time are summarized in Table 3. Formulas that are bolded represent

219

metabolites that were qualified with both MS and MS/MS fragmentation data (Level 2 certainty

220

with diagnostic evidence) while those not bolded and italicized metabolites were tentatively

221

identified with MS data but for which the MS/MS data was insufficient to confirm structures

222

(Level 3 certainty). Higher levels of certainty could not be obtained in this study because

223

reference standards for most of the metabolites identified are not commercially available; a

11 ACS Paragon Plus Environment

Environmental Science & Technology

224

library that included the metabolites was not available; and compounds were not known for

225

synthesis prior to this study. MS data and where available MS/MS fragmentation data for

226

metabolites that were confirmed as tentative candidate with level 3 certainty and those identified

227

with certainty at a Level 2 with diagnostic evidence (referred to as unconfirmed) are provided in

228

Figures S4 and S5, respectively. A chromatogram with relative retention times of metabolites

229

identified and target compounds are provided in Fig. S6. All metabolites listed were observed in

230

soil extracts of both soils amended with BPs and absent from extracts of soil controls (soils with

231

no addition of BPs), but not at all sampling times (Tables S5 and S6). For BPA, in particular, but

232

also BPS, a high abundance of interfering peaks were found from both endogenous and

233

structurally similar compounds that were in the soil extracts reducing the ability to clearly

234

identify metabolites of interest specific to bisphenol degradation, thus only a few metabolites

235

were identified for BPA and BPS compared to BPAF. It is also important to note that some

236

degradation steps can be so fast that intermediate metabolite concentrations are too low to

237

observe. We did see peaks with MS values from which a ring hydrolyzed metabolite mass could

238

be derived, but peaks were of low intensity and MS/MS fragmentation was not triggered.

239

The most prevalent transformation reaction identified in aerobic conditions with

240

confirmed intermediates is hydroxylation, substitution, and rearrangement (summarized in Fig.

241

S3). Of the metabolites identified in our study, only hydroxylated metabolites E and F for BPA

242

metabolites (Table 3, Table S4 and Fig. S4) were similar to those in bacteria isolate studies

243

(structure 6 and 7 in Fig. S3) with no similar parallel metabolites for BPS and BPAF observed.

244

Most of the other metabolites found are best explained by ring cleavage following hydroxylation

245

with metabolites reflecting ortho-cleavage for all three BPs (D, H, J, L, and U in Table 3, Figures

246

S4 and S5), which has not been shown previously, as well as meta-cleavage (G, I, M, N, and T in

12 ACS Paragon Plus Environment

Page 12 of 22

Page 13 of 22

Environmental Science & Technology

247

Table 3, Figures S4 and S5) and post cleavage (K, O, and S in Table 3, Figures S4 and S5).

248

However, some meta-cleavage metabolites we identified are different than those presumed to be

249

intermediates by Ogato et al.47 (N and T in Table 3, Fig. S4 and S5, bracketed intermediates in

250

Fig. S3 compound #21).

251

Degradation pathway. In studies using different bacteria isolates,22,45–47 several BPA

252

degradation pathways were proposed from confirmed metabolites (Fig. S3 in SI). Most of the

253

pathways proposed involve skeletal rearrangements and no ring cleavage (Pathways A45, B22,

254

and C46 in Fig. S3). However, Ogato et al.47 observed hydroxylation and meta-cleavage of ring

255

(Pathway D in Fig. S3) with no changes to the alkyl group connecting the two phenolic rings.

256

The latter was also observed for other bisphenol structures including BPS indicating that the

257

alkyl linkage did not affect susceptibility of the ring to hydroxylation and meta-cleavage.

258

Catechol has been shown to be readily degraded by ortho- and meta-cleavage

259

dioxygenases which are common in soil bacteria.48,49 Based on what has been for ortho-cleavage

260

in catechol degradation by intradiol dioxygenases and what we observed for BPA, BPAF, and

261

BPS, we have proposed a BP degradation pathway involving ortho-cleavage in Fig. 2 (pathway

262

a). Also, using the limited number of metabolites identified from BPA and BPAF degradation

263

that can be explained by applying what has been observed for meta-cleavage of catechol by

264

extradiol dioxygenases, we proposed a meta-cleavage pathway for BPs (pathway b in Fig. 2).

265

Although ring hydroxylation precedes ring cleavage,50 we only observed hydroxylated BPA (E

266

and F in Table 3). The pathways shown in Fig. 2 are focused on cleavage of just one of the

267

aromatic rings in BP since this is what aligns with the metabolites we identified with the

268

exception of F and U (Table 3) in which both rings were hydroxylated. Also metabolite V from

269

BPAF, which could follow after the initial cleavage (likely meta-cleavage due to the structural

13 ACS Paragon Plus Environment

Environmental Science & Technology

270

similarity), is not represented in Fig. 2. It is reasonable to assume that both rings are undergoing

271

cleavage with subsequent degradation to smaller metabolites and possibly mineralization except

272

in the case of BPAF, which likely leaves polyfluorinated alkyl metabolite residuals.

273 274 275

ENVIRONMENTAL IMPLICATIONS Rapid microbial degradation (half-lives less than one day) under aerobic conditions for

276

BPA and BPS suggests that they will likely not accumulate in soils from land-application of

277

biosolids or wastewater effluent irrigation. This would seem contradictory to the repeated

278

occurrences in soils, but this is simply likely due to extensive inputs from the massive production

279

and use of BPA1. With replacement of BPA to alternatives such as BPS and BPAF, inputs of

280

BPA should decrease in the environment while the alternatives will increase. Given BPAF’s

281

slow microbial degradation and high sorption affinity, increasing inputs of BPAF will result in

282

an increased levels of BPAF if it continues to replace BPA. In addition, the two highly

283

electronegative CF3 groups in BPAF make skeletal rearrangement and hydroxylation prior to

284

cleavage less favorable, which leads to metabolites of concern also remaining in soils for

285

extended periods of time. This slow degradation process of BPAF allow us to detect BPAF

286

metabolites, which aided in proposing an additional and new degradation pathway for

287

bisphenols.

288 289

ACKNOWLEDGEMENTS

290

This work was funded in part by support provided by the Purdue Research Foundation and the

291

Purdue Agronomy Department.

292

14 ACS Paragon Plus Environment

Page 14 of 22

Page 15 of 22

Environmental Science & Technology

293

SUPPORTING INFORMATION

294

Supporting information is available that includes physicochemical properties of BPA, BPAF and

295

BPS, additional soil properties detail, additional information of chromatographic analysis

296

(HPLC-MS/MS, UPLC-QToF), a soil matrix evaluation, the kinetic analysis using CAKE

297

software, and information on identified metabolites, including spectral data.

298 299 300 301 302 303 304 305 306 307 308 309 310 311 312 313 314 315 316 317 318 319 320 321 322 323 324 325 326 327 328 329 330 331 332

REFERENCES (1) U. S. Environmental Protection Agency. Bisphenol A Action Plan. 2010, 1–22. (2) Nam, S. H.; Seo, Y. M.; Kim, M. G. Bisphenol A migration from polycarbonate baby bottle with repeated use. Chemosphere 2010, 79 (9), 949–952. (3) Kubwabo, C.; Kosarac, I.; Stewart, B.; Gauthier, B. R.; Lalonde, K.; Lalonde, P. J. Migration of bisphenol A from plastic baby bottles, baby bottle liners and reusable polycarbonate drinking bottles. Food Addit. Contam. Part A. Chem. Anal. Control. Expo. Risk Assess. 2009, 26 (6), 928–937. (4) Cooper, J. E.; Kendig, E. L.; Belcher, S. M. Assessment of bisphenol A released from reusable plastic, aluminium and stainless steel water bottles. Chemosphere 2011, 85 (6), 943–947. (5) Biles, J. E.; McNeal, T. P.; Begley, T. H.; Hollifield, H. C. Determination of bisphenol-A in reusable polycarbonate food-contact plastics and migration to food-simulating liquids. J. Agric. Food Chem 1997, 45 (9), 3541–3544. (6) Viñas, P.; Campillo, N.; Martínez-Castillo, N.; Hernández-Córdoba, M. Comparison of two derivatization-based methods for solid-phase microextraction-gas chromatographymass spectrometric determination of bisphenol A, bisphenol S and biphenol migrated from food cans. Anal. Bioanal. Chem. 2010, 397 (1), 115–125. (7) Brotons, J. A.; Olea-Serrano, M. F.; Villalobos, M.; Pedraza, V.; Olea, N. Xenoestrogens released from lacquer coatings in food cans. Environ. Health Perspect. 1995, 103 (6), 608–612. (8) FDA. Indirect Food Additives: Polymers 21 CFR Part 177 FDA-2012-F-0031-0007, 2012. (9) Perera, L.; Li, Y.; Coons, L. A.; Houtman, R.; van Beuningen, R.; Goodwin, B.; Auerbach, S. S.; Teng, C. T. Binding of bisphenol A, bisphenol AF, and bisphenol S on the androgen receptor: Coregulator recruitment and stimulation of potential interaction sites. Toxicol. Vitr. 2017, 44, 287–302. (10) Ruan, T.; Liang, D.; Song, S.; Song, M.; Wang, H.; Jiang, G. Evaluation of the in vitro estrogenicity of emerging bisphenol analogs and their respective estrogenic contributions in municipal sewage sludge in China. Chemosphere 2015, 124, 150–155. (11) Fürhacker, M.; Scharf, S.; Weber, H. Bisphenol A: emissions from point sources. Chemosphere 2000, 41 (5), 751–756. (12) Mohapatra, D. P.; Brar, S. K.; Tyagi, R. D.; Surampalli, R. Y. Occurrence of bisphenol A in wastewater and wastewater sludge of CUQ treatment plant. J. Xenobiotics 2011, 1 (1), 9–16. 15 ACS Paragon Plus Environment

Environmental Science & Technology

333 334 335 336 337 338 339 340 341 342 343 344 345 346 347 348 349 350 351 352 353 354 355 356 357 358 359 360 361 362 363 364 365 366 367 368 369 370 371 372 373 374 375 376 377 378

(13) Musolff, A.; Leschik, S.; Reinstorf, F.; Strauch, G.; Schirmer, M. Micropollutant loads in the urban water cycle. Environ. Sci. Technol. 2010, 44 (13), 4877–4883. (14) Kinney, C. a; Furlong, E. T.; Zaugg, S. D.; Burkhard, M. R.; Werner, S. L.; Cahill, J. D.; Jorgensen, G. R. Survey of organic wastewater contaminants in biosolids destined for land application. Environ. Sci. Technol. 2006, 40 (23), 7207–7215. (15) Yu, X.; Xue, J.; Yao, H.; Wu, Q.; Venkatesan, A. K.; Halden, R. U.; Kannan, K. Occurrence and estrogenic potency of eight bisphenol analogs in sewage sludge from the U.S. EPA targeted national sewage sludge survey. J. Hazard. Mater. 2015, 299, 733–739. (16) Yang, Y.; Lu, L.; Zhang, J.; Yang, Y.; Wu, Y.; Shao, B. Simultaneous determination of seven bisphenols in environmental water and solid samples by liquid chromatographyelectrospray tandem mass spectrometry. J. Chromatogr. A 2014, 1328, 26–34. (17) Liao, C.; Liu, F.; Moon, H.-B.; Yamashita, N.; Yun, S.; Kannan, K. Bisphenol analogues in sediments from industrialized areas in the United States, Japan, and Korea: spatial and temporal distributions. Environ. Sci. Technol. 2012, 46 (21), 11558–11565. (18) Nie, Y.; Qiang, Z.; Zhang, H.; Adams, C. Determination of endocrine-disrupting chemicals in the liquid and solid phases of activated sludge by solid phase extraction and gas chromatography-mass spectrometry. J. Chromatogr. A 2009, 1216 (42), 7071–7080. (19) Choi, Y. Distribution and degradation of Bisphenol A (BPA) substitutes BPAF and BPS compared to BPA in aerobic soil and anaerobic, Purdue University, 2016. (20) Voordeckers, J. W.; Fennell, D. E.; Jones, K.; Häggblom, M. M. Anaerobic biotransformation of tetrabromobisphenol A, tetrachlorobisphenol A, and bisphenol A in estuarine sediments. Environ. Sci. Technol. 2002, 36 (4), 696–701. (21) Chang, B. V; Yuan, S. Y.; Chiou, C. C. Biodegradation of bisphenol-A in river sediment. J. Environ. Sci. Health. A. Tox. Hazard. Subst. Environ. Eng. 2011, 46 (9), 931–937. (22) Ronen, Z.; Abeliovich, a. Anaerobic-aerobic process for microbial degradation of tetrabromobisphenol A. Appl. Environ. Microbiol. 2000, 66 (6), 2372–2377. (23) Ying, G.-G.; Kookana, R. S. Sorption and degradation of estrogen-like-endocrine disrupting chemicals in soil. Environ. Toxicol. Chem. 2005, 24 (10), 2640–2645. (24) Dorn, P. B.; Chou, C.; Gentempo, J. J. Degradation of Bisphenol A in natural waters. Chemosphere 1987, 16 (7), 1501–1507. (25) Klecka, G. M.; Gonsior, S. J.; West, R. J.; Goodwin, P.; Markham, D. Biodegradation of bisphenol A in aquatic environments: river die-away. Environ. Toxicol. Chem. 2001, 20 (12), 2725–2735. (26) Kang, J.-H.; Kondo, F. Bisphenol A degradation in seawater is different from that in river water. Chemosphere 2005, 60 (9), 1288–1292. (27) Ike, M.; Chen, M. Y.; Danzl, E.; Sei, K.; Fujita, M. Biodegradation of a variety of bisphenols under aerobic and anaerobic conditions. Water Sci. Technol. 2006, 53 (6), 153. (28) Sarmah, A. K.; Northcott, G. L. Laboratory degradation studies of four endocrine disruptors in two environmental media. Environ. Toxicol. Chem. 2008, 27 (4), 819–827. (29) Li, J.; Jiang, L.; Liu, X.; Lv, J. Adsorption and aerobic biodegradation of four selected endocrine disrupting chemicals in soil-water system. Int. Biodeterior. Biodegrad. 2013, 76, 3–7. (30) Xu, J.; Wu, L.; Chang, A. C. Degradation and adsorption of selected pharmaceuticals and personal care products (PPCPs) in agricultural soils. Chemosphere 2009, 77 (10), 1299– 1305. (31) Robinson, B. J.; Hellou, J. Biodegradation of endocrine disrupting compounds in harbour 16 ACS Paragon Plus Environment

Page 16 of 22

Page 17 of 22

379 380 381 382 383 384 385 386 387 388 389 390 391 392 393 394 395 396 397 398 399 400 401 402 403 404 405 406 407 408 409 410 411 412 413 414 415 416 417 418 419 420 421 422 423 424

Environmental Science & Technology

(32) (33) (34) (35) (36)

(37) (38) (39)

(40) (41) (42) (43) (44)

(45) (46)

(47)

(48) (49) (50)

seawater and sediments. Sci. Total Environ. 2009, 407 (21), 5713–5718. Ying, G. G.; Kookana, R. S. Degradation of five selected endocrine-disrupting chemicals in seawater and marine sediment. Environ. Sci. Technol. 2003, 37 (7), 1256–1260. Danzl, E.; Sei, K.; Soda, S.; Ike, M.; Fujita, M. Biodegradation of bisphenol A, bisphenol F and bisphenol S in seawater. Int. J. Environ. Res. Public Health 2009, 6 (4), 1472–1484. Mashtare, M. L.; Green, D. a; Lee, L. S. Biotransformation of 17α- and 17β-estradiol in aerobic soils. Chemosphere 2013, 90 (2), 647–652. U.S. EPA. Fate , Transport , and Transformation Test Guidelines OPPTS 835.4100 Aerobic Soil Metabolism. 2008. Wolf, D. C.; Dao, T. H.; Scott, H. D.; Lavy, T. L. (1989) Influence of Sterilization Methods on Selected Soil Microbiological, Physical, and Chemical Properties. J. Environ. Qual. 1989, 18 (1), 39–44. Mashtare, M. L.; Khan, B.; Lee, L. S. Evaluating stereoselective sorption by soils of 17αestradiol and 17β-estradiol. Chemosphere 2011, 82 (6), 847–852. Dasu, K.; Liu, J.; Lee, L. S. Aerobic soil biodegradation of 8:2 fluorotelomer stearate monoester. Environ. Sci. Technol. 2012, 46 (7), 3831–3836. Khan, B.; Lee, L. S. Soil temperature and moisture effects on the persistence of synthetic androgen 17alpha-trenbolone, 17beta-trenbolone and trendione. Chemosphere 2010, 79 (8), 873–879. Fent, G.; Hein, W. J.; Moendel, M. J.; Kubiak, R. Fate of 14C-bisphenol A in soils. Chemosphere 2003, 51 (8), 735–746. Feng, Y.; Park, J. H.; Voice, T. C.; Boyd, S. A. Bioavailability of soil-sorbed biphenyl to bacteria. Environ. Sci. Technol. 2000, 34 (10), 1977–1984. Scow, K. M.; Johnson, C. R. Effect of Sorption on Biodegradation of Soil Pollutants. Adv. Agron. 1996, 58 (C), 1–56. Choi, Y. J.; Lee, L. S. Partitioning Behavior of Bisphenol Alternatives BPS and BPAF Compared to BPA. Environ. Sci. Technol. 2017, 51 (7), 3725–3732. Schymanski, E. L.; Jeon, J.; Gulde, R.; Fenner, K.; Ruff, M.; Singer, H. P.; Hollender, J. Identifying small molecules via high resolution mass spectrometry: Communicating confidence. Environ. Sci. Technol. 2014, 48 (4), 2097–2098. Spivacks, J.; Leib, T. K.; Lobos, J. H. Novel Pathway for Bacterial Metabolism of Bisphenol A. J. Biol. Chem. 1994, 269 (10), 7323–7329. Kolvenbach, B.; Schlaich, N.; Raoui, Z.; Prell, J.; Zühlke, S.; Schäffer, A.; Guengerich, F. P.; Corvini, P. F. X. Degradation pathway of bisphenol A: Does ipso substitution apply to phenols containing a quaternary α-carbon structure in the para position? Appl. Environ. Microbiol. 2007, 73 (15), 4776–4784. Ogata, Y.; Goda, S.; Toyama, T.; Sei, K.; Ike, M. The 4-tert-butylphenol-utilizing bacterium Sphingobium fuliginis OMI can degrade bisphenols via phenolic ring hydroxylation and meta-cleavage pathway. Environ. Sci. Technol. 2013, 47 (2), 1017– 1023. Hayaishi, O.; Katagiri, M.; Rothberg, S. Studies on oxygenases; pyrocatechase. J. Biol. Chem. 1957, 229 (2), 905–920. Kojima, Y.; Itada, N.; Hayaishi, O. Metapyrocatechase : a New Catechol-cleaving Enzyme. J. Biol. Chem. 1961, 236 (8), 2223–2228. Fuchs, G.; Boll, M.; Heider, J. Microbial degradation of aromatic compounds — from one strategy to four. Nat. Rev. Microbiol. 2011, 9 (11), 803–816. 17 ACS Paragon Plus Environment

Environmental Science & Technology

425 426

Page 18 of 22

Table 1. Selected properties of soils used in the aerobic degradation studies. Additional details can be found in Table S4 of the SI. Soil Texture (%)d

427 428 429 430 431 432

Soil ID

Organic carbon (%)a

Soil pHb

CECc

FRST-50

2.7

5.8

PSF-51

1.5

6.2

Moisture content (%)f

Sand

Silt

Clay

75% field capacitye

11.3

36

36

28

26.9

20.2

6.9

36

34

30

16.5

15.6

a

Percent soil organic carbon = % organic matter/1.72, organic matter content determined by loss on ignition (LOI) method; bpH of a 1 g:1 mL soil:water slurry; cCation exchange capacity determined by the ammonium acetate method, cmol/kg; eParticle size analysis determined by hydrometer method; dSoil textural classification following USDA-NRCS by hydrometer method; e 75% of the moisture content determined at 0.01 MPa (field capacity); f Moisture content (%) of soils right after sampling = (ambient soil wt.-oven dried soil wt.)/ ambient soil wt.* 100.

18 ACS Paragon Plus Environment

Page 19 of 22

433 434

Environmental Science & Technology

Table 2. Model outputs from the CAKE model fits to the data using Single First Order (SFO) and Double First Order in Parallel (DFOP) kinetic models. BPA Observed t1/2 (d)

BPAF

BPS

FRST

PSF

FRST

PSF

FRST

PSF

0.75

0.39

32.6

24.5

0.69

0.59

0.471 (0.073)b

1.55 (0.39)

0.0202 (0.0022)

0.0262 (0.0031)

0.632 (0.059)

1.05 (0.08)

1.47 4.89 0.788

0.447 1.49 0.810

34.3 114 0.860

26.4 87.8 0.890

1.10 3.64 0.944

0.661 2.20 0.963

30.0 (5.17)

6.74 (8.76)

0.783 (0.510)

0.0378 (0.0124)

0.997 (0.341)

1.09 (0.122)

a

SFO k1e (SD)b DT50 (t1/2) DT90 R2 DFOP k1 (SD)

0.171 0.175 0.0109 0.00129 0.180 0.0228 (0.027) (0.102) (0.0014) (0.01025) (0.159) (0.181) 0.421 0.700 0.258 0.863 0.765 0.985 Gg (SD) (0.042) (0.096) (0.053) (0.200) (0.225) (0.036) c DT50 (t1/2) 0.863 0.179 36.3 22.6 0.935 0.649 c DT90 10.3 6.27 184 221 5.04 2.24 2d R 0.949 0.878 0.884 0.901 0.950 0.963 a b Observed t1/2 (d) interpolated from the two closest measured values; SD is standard deviation; t1/2 is half-life (d); c DT50 and DT90 are CAKE outputs for the times (d) required for the time 0 concentration to decline by 50% and 90%, respectively; dR2 is the coefficient of determination; e k1 (d-1) is the first-order rate constant for the SFOP and the rate constant for compartment 1 of the DFOP model; fk2 (d-1) is the rate constant for compartment 2 of DFOP model; gg and 1-g are the fractions of the total concentration that is subject to k1 and k2, respectively in the DFOP model fit. k2f (SD)

435 436 437 438 439 440 441

19 ACS Paragon Plus Environment

Environmental Science & Technology

442 443 444 445 446 447 448

Table 3. Structure and molecular weight of the metabolites of BPA, BPS and BPAF found in the microcosms for both soils amended with target compounds and also absent from soil controls. Formulas representing metabolites that were qualified with both MS and MS/MS fragmentation data are bolded whereas metabolites that were identified but information was insufficient to confirm are italicized and not bolded. Parenthetical letters in the formula mass column refer to the graphs in Fig. S4 for confirmed metabolites and Fig. S5 for unconfirmed metabolites. See Tables S5 and S6 for a summary of which sampling time these metabolites were observed.

449

452

BPA Metabolites Formula Structure Mass C15H18O6 294.1103 (D)

BPAF Metabolites Formula Structure Mass C15H10F6O5 384.0432 (L) C14H12F6O5 374.0589 (M)

C15H16O3 244.1099 (E)

C14H10F6O4 356.0483 (N)

C15H16O4 260.1049 (F)

C11H8F6O3 302.0378 (O)

C15H16O5 276.0998 (G, H, I)

C9H6F6O 44.0323 (P) C11H8F6O2 286.0429 (S)

BPS Metabolites Formula Structure Mass C12H10O8S 314.0096 (U) C12H10O7S 298.0147 (J) C8H8O4S 200.0143 (K)

C14H10F6O5 372. 0432 (T) C13H10F6O5 360.0432 (V) 450 451

453 20 ACS Paragon Plus Environment

Page 20 of 22

Page 21 of 22

Environmental Science & Technology

454 120

120

BPA_FRST

100

100

80

80

60

60

40

40

20

20 0

0 8

6

4

2

0

10

Remaining (%)

120

12

14

10

120

BPAF_FRST

100

100

80

80

60

60

40

40

20

20

12

14

16

BPAF_PSF

0 20

0

40

60

80

40

20

0

100 120 140 160 180

120

60

80

100 120 140 160 180

120

BPS_FRST

100

100

80

80

60

60

40

40

20

20

BPS_PSF

Measured < MLOQ SFO DFOP

0

0 0

456 457 458 459

8

6

4

2

0

16

0

455

BPA_PSF

2

4

6

8 10 12 14 16 18 20 22 24 26 28 30

0

2

4

6

8 10 12 14 16 18 20 22 24 26 28 30

Time (Days)

Time (Days)

Figure 1. Aerobic degradation of BPA, BPAF and BPS in a forested (FRST) soil and farm (PSF) soil. Vertical bars are standard deviations. Open circles are < MLOQ (method limit of quantitation). Solid and dashed lines are SFO (single first-order) and DFOP (double first-order parallel model fits to the data, respectively.

460 461

21 ACS Paragon Plus Environment

Environmental Science & Technology

462

463 464 465 466 467 468

Figure 2. Proposed cleavage pathway postulated from oxidative degradation of catechol by (a) intradiol dioxygenase (ortho-cleavage) and (b) extradiol dioxygenage (meta-cleavage). In the molecular structures shown, X represents carbon or sulfur linkages to two phenol groups and Y represents CH3, CF3, or O groups to bridging atom. Letters under structures refer to metabolites shown in Table 3, and Figures S4 and S5.

22 ACS Paragon Plus Environment

Page 22 of 22