Aerobic Soil Biodegradation of Bisphenol (BPA) Alternatives

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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

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Environmental Science & Technology

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

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Bisphenol AF Compared to BPA

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Youn Jeong Choi and Linda S. Lee*

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Purdue University, Department of Agronomy, Ecological Science and Engineering

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Interdisciplinary Graduate Program, West Lafayette, IN 47907-2054

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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)

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Prepared for Environmental Science & Technology

October 20, 2017

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1 ACS Paragon Plus Environment


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Graphical Abstract

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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kinetics



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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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


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

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

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

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

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

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

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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

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

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

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

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

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

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

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

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for which data are lacking or non-existent.

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

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

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

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

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

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



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

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

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studies were initiated within 3 months of sampling.

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

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

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

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

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

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

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

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

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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

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

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

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

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

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

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

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deionized water were also autoclave-sterilized.

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

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

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


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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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added uniformly to all vials.

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

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

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

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



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

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

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summarized in Table S2.

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Isotopically labeled internal standards and external calibration curves were used to

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

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

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

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

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

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

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

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

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

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

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

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additional details).

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

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

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

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

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

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

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


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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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



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

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

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

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

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

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to ensure adequate O2 levels.

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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


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

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

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

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

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

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

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

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

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

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not quantified in this study.

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

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

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

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

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

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

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

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

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

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

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

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

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



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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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The most prevalent transformation reaction identified in aerobic conditions with

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

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

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

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

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

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

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


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

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

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

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Fig. S3 compound #21).

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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



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

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cleavage with subsequent degradation to smaller metabolites and possibly mineralization except

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

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

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

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

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

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

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

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

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

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

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

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

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

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bisphenols.

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ACKNOWLEDGEMENTS

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

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Purdue Agronomy Department.

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SUPPORTING INFORMATION

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

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

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

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

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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.


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433 434


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



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

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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



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.


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Aerobic Soil Biodegradation of Bisphenol (BPA) Alternatives

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