Stimulation of Tetrabromobisphenol A Binding to Soil Humic

May 25, 2016 - (27) The chemical properties of the two HA are listed in Table S1. Prior to their use in the experiment, both HAs were dissolved in 0.1...
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Stimulation of Tetrabromobisphenol A Binding to Soil Humic Substances by Birnessite and the Chemical Structure of the Bound Residues Fei Tong,† Xueyuan Gu,*,† Cheng Gu,† Jinyu Xie,† Xianchuan Xie,‡ Bingqi Jiang,† Yongfeng Wang,† Tanya Ertunc,§ Andreas Schaff̈ er,§ and Rong Ji† †

State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment and ‡Center for Hydrosciences Research, State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, PRC § Institute for Environmental Research, RWTH Aachen University, D-52056 Aachen, Germany S Supporting Information *

ABSTRACT: Studies have shown the main fate of the flame retardant tetrabromobisphenol A (TBBPA) in soils is the formation of bound residues, and mechanisms on it are lessunderstood. This study investigated the effect of birnessite (δMnO2), a naturally occurring oxidant in soils, on the formation of bound residues. 14C-labeled TBBPA was used to investigate the pH dependency of TBBPA bound-residue formation to two soil humic acids (HAs), Elliott soil HA and Steinkreuz soil HA, in the presence of δ-MnO2. The binding of TBBPA and its transformation products to both HAs was markedly increased (3- to 17-fold) at all pH values in the presence of δ-MnO2. More bound residues were formed with the more aromatic Elliott soil HA than with Steinkreuz soil HA. Gel-permeation chromatography revealed a uniform distribution of the bound residues within Steinkreuz soil HA and a nonuniform distribution within Elliott soil HA. 13C NMR spectroscopy of 13C-TBBPA residues bound to 13C-depleted HA suggested that in the presence of δ-MnO2, binding occurred via ester and ether and other types of covalent bonds besides HA sequestration. The insights gained in this study contribute to an understanding of the formation of TBBPA bound residues facilitated by δ-MnO2.

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INTRODUCTION Tetrabromobisphenol A (TBBPA) is the most widely utilized brominated flame retardant in the world.1 Due to its structural similarities to the thyroid hormone thyroxin and to bisphenol A, a suspected endocrine disruptor, TBBPA is also regarded to have endocrine-disrupting properties.1 Common routes of TBBPA entry into the environment are through the disposal of waste electric and electronic equipment (WEEE) and as effluents of municipal wastewater-treatment plants.2,3 TBBPA has been widely detected in various environmental matrices, such as air, sewage sludge, sediment, and soil. 1 Its concentrations in soils range from nanograms to micrograms per gram but was as high as 450 mg (kg of soil)−1 at a contaminated site in Israel.4 TBBPA can be adsorbed by soils,5 with soil organic matter (SOM) contributing most to sorption. TBBPA may also undergo biotic degradation by soil microbes6−8 or abiotic transformation by soil metal oxides.9 Alternatively, it can be immobilized via the formation of nonextractable residues (bound residues).7 Recent studies7,8,10 have shown the main fate of TBBPA in soils is forming bound residues with SOM. © XXXX American Chemical Society

Besides SOM, other soil components, such as active oxides (e.g., birnessite) or soil enzymes (e.g., laccase11), may also influence the fate of TBBPA in soils by readily reacting with it.9 Indeed, birnessite (δ-MnO2) is the strongest naturally occurring oxidant (E0 = 1.2 V for MnO2) in the soil environment,12 which can catalytically oxidize both phenolic and nonphenolic organic contaminants, leading to the mineralization or humification of these organics in soil.9,13−15 It has been reported that dihydroxybenzenes (such as catechol) could be mineralized by birnessite because the two hydroxyl-bound carbons in the benzene ring have high potential to form −COOH group bound aliphatic fragments through ring cleavage and then subsequently release CO2 via decarboxylation.13 In contrast, ring carbons of monophenols are less susceptible to mineralization than dihydroxybenzenes.16,17 Abiotic oxidation mediated by birnessite may be a Received: December 22, 2015 Revised: May 18, 2016 Accepted: May 25, 2016

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DOI: 10.1021/acs.est.5b06265 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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study. It was isolated from a 13C-depleted compost using a method similar to that described for the Steinkreuz soil HA, dissolved in 0.1 M NaOH under a N2 atmosphere, and adjusted to pH 7 before its use in the experiment. Preparation of the 13 C-depleted compost was described elsewhere.16 The chemical property of the 13C-depleted HA is listed in Table S1. Isotope-Labeled TBBPA. Uniformly 14C-ring-labeled TBBPA (2.442 × 1012 Bq mol−1, 99% radiochemical purity and 97% chemical purity) was synthesized from 14C-ringlabeled phenol.28 The synthesis of a uniformly 13C-ring-labeled TBBPA from 13C-ring-labeled phenol was similar to that described for 14C-ring-labeled TBBPA. Reaction of 14C-TBBPA with Birnessite in the Presence of HA. The oxidation of TBBPA by birnessite in the presence of HA was conducted in closed glass bottles (60 mL) with a rubber stopper from which a 5 mL vial containing 0.5 mL NaOH (1 M) was suspended. A total of three reaction systems were used in this study: (1) 14C-TBBPA and δ-MnO2; (2) 14CTBBPA and HA; and (3) 14C-TBBPA, HA, and δ-MnO2. The 1.6 mL of reaction mixture consisted of 10 mg of birnessite, 0.16 μg of 14C-TBBPA (total radioactivity: 740 Bq), and 0.8 mg C of HA. Before the addition of TBBPA, the reaction solutions were adjusted to pH 5, 6, 7, or 8 with diluted HCl or NaOH according to the preliminary experiment on pH adjustment and balanced within 24 h until pH was stable. The variation of pH during the incubation was negligible. The bottles containing the reaction mixtures were gently shaken on a rotary shaker (10 rpm) for 48 h. The suspended vial served as the 14CO2 trap, which absorbed the 14CO2 released during the reaction. The 14 CO2 trap was replaced by a fresh one at scheduled intervals, and the amount of 14CO2 in the collected trap was determined by liquid scintillation counting (LSC, see below). All experiments were performed in duplicate. At the end of the reaction (48 h), birnessite was separated by centrifugation. The supernatant was acidified to pH 1 to precipitate the HA. Ethyl acetate (EtOAc) was then added (1:1, v/v) and the tubes were shaken for 3 h (120 rpm). The acidic condition did not lead to ester hydrolysis such as for EtOAc. The EtOAc extraction was repeated three times, after which the upper layers were pooled for LSC determination of the extractable residues of 14C-TBBPA. In addition, the parent TBBPA molecules in the EtOAc extracts were analyzed using high-performance liquid chromatography (HPLC) (see the Supporting Information). The aqueous supernatant was collected for LSC determination of the amount of watersoluble residues of 14C-TBBPA. In addition, concentration of Mn2+ in the aqueous supernatant was measured by inductively coupled plasma optical emission spectroscopy (ICP-OES). The HA pellet was dissolved in 0.1 M anoxic NaOH and stored under a N2 atmosphere to avoid autoxidation of HA under alkaline conditions.29 The radioactivity within the HA was determined by LSC and the results interpreted as the content of nonextractable residues of 14C-TBBPA. In this study, the EtOAc unextractable residue is defined as bound residue according to the International Union of Pure and Applied Chemistry (IUPAC).30 The molecular size distribution of the 14 C-TBBPA residues within the HA was analyzed by highperformance gel permeation chromatography (HP-GPC; see below) in combination with LSC (HP-GPC−14C-LSC). In addition, the birnessite particles separated after the reaction were air-dried and then combusted with an oxidizer (biological oxidizer OX500; Zinsser Analytical). The generated 14CO2 was

possible pathway of mineralization for TBBPA in soils, although it has not been fully examined so far. Oxidation of phenolic compounds mediated by birnessite can also lead to degradation, polymerization, and humus formation.9,18−20 For example, it has been reported that the oxidation of TBBPA by birnessite in a 50/50 (v/v) water− methanol matrix yielded seven reaction products.9 The oxidative intermediates of TBBPA may be incorporated into the polymeric structures of humic substances.18,19 It is important to understand the chemical nature of the bound residue. However, the degradation behavior of TBBPA in the presence of both birnessite and humic substance are unknown. In the oxidation reaction between Mn oxides and organics, for example, phenol oxidation is generally initiated by a surface complex between the adsorbate and the oxide21 and then a phenoxyl radical, resulting from a one-electron transfer within the complex, and radical-induced reaction products are formed.12 When humic acids (HAs) are included in the system, they react with the phenoxyl radical to form covalently bound residues.16,22,23 Analogously, birnessite may also oxidize phenolic moieties in HA, forming HA phenoxy radicals that subsequently react with TBBPA to form covalent bonds. Therefore, the objective of this work was to (i) investigate the effect of HAs on the mineralization of TBBPA mediated by birnessite as a function of pH; (ii) more importantly, examine the bound-residue formation of TBBPA with the HAs in the presence of birnessite as a function of pH; and, (iii) above all, investigate the chemical nature of the TBBPA bound residues.

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MATERIALS AND METHODS Preparation of Birnessite and Humic Acids. Birnessite (δ-MnO2) was synthesized according to the method described by McKenzie.24 In brief, 170 mL of concentrated HCl was added to a boiling solution of 1 mol of KMnO4 in 2.5 L of H2O. The formed precipitate was washed repeatedly with 1 mM HCl and rinsed with deionized water until the pH of the supernatant was 4.0. The solid was freeze-dried for further use. X-ray diffraction of the birnessite particles is shown in Figure S1. The surface area of birnessite, measured using the N2−BET method (ASAP 2020, Micromeritics), was 43.13 m2 g−1, comparable to the 40.0 m2 g−1 determined by Lee et al.25 using a similar method of synthesis. A total of two HAs extracted from soils were used in this study. The first Elliott prairie soil HA (1S102H) was obtained from the International Humic Substances Society (IHSS). Both elemental composition and 13C nuclear magnetic resonance (NMR) estimation of the carbon distribution of this HA are provided on the Web site of the IHSS.26 According to the 13C NMR data, the proportion of aromatic carbon to the total carbon content of Elliott soil HA is 50%. The second HA was extracted from a Steinkreuz forest soil (Steigerwald, Germany) using 0.1 M NaOH, followed by precipitation at pH 1, dialysis, and lyophilization. A detailed description of this HA was published elsewhere.27 The aromatic carbon of Steinkreuz soil HA accounts for only 7% of the total carbon content, as estimated from the 13C NMR data.27 The chemical properties of the two HA are listed in Table S1. Prior to their use in the experiment, both HAs were dissolved in 0.1 M NaOH (1 g C L−1) under N2 atmosphere, and the pH of the resulting suspensions was adjusted to pH 7. 13 C-Depleted HA. To weaken the background signals of HA in the 13C NMR analysis (see below), we also used a 13Cdepleted HA (with 99.95% of 12C atom enrichment) in this B

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GPC column (PSS MCX 1000 Å, 8 mm × 300 mm; Polymer Standards Service GmbH) connected to an HPLC system (HP 1100; Agilent Technology) equipped with a diode array detector. The eluent was 0.6% K2CO3 (pH 11) at a flow rate of 0.1 mL min−1. The temperature of the GPC column was set at 40 °C. To avoid the possible hydrolysis of ester bonds under such condition, we checked phenyl acetate as a reference sample by equilibrating under same condition (0.6% K2CO3, pH 11, and room temperature) for 24 h, and the result showed no observable hydrolysis of phenyl acetate. The absorbance of the HA at 254 nm was recorded simultaneously. Each sample was analyzed for 150 min. The molecular masses of the compounds separated by the GPC column ranged from 200 Da to 100 kDa, determined based on calibrations using molecular size standards of polystyrenesulfonate sodium salts.16 During the elution of the HP-GPC system, eluent fractions were sampled every 2 min and their radioactivity counted by LSC.16,35 Solution 13C NMR Spectroscopy. Solution 13C NMR was run on a Bruker AVANCE III 600 MHz spectrometer (1H at 600 MHz; 13C at 150 MHz) equipped with a 14.09 T magnetic field. The 13C-depleted HA with bound residues of 13C-TBBPA was dissolved in 0.1 M anoxic deuterated NaOH (0.6 mL) in 5 mm NMR tubes (541-PP-7; Wilmad-Labglass, Vineland, NJ) under Ar gas protection. An inverse gated decoupling method was used to record ∼12000 scans. The 13C NMR spectrum of pure 13C-TBBPA was analyzed in CDCl3 solution, also using an inverse gated decoupling method. An internal standard of 1% tetramethylsilane was included.

absorbed in 10 mL of alkaline scintillation cocktail and then quantified by LSC. The total recovery of radioactivity in all the experiments ranged from 94 to 99%. Reaction of 13C-TBBPA with Birnessite in the Presence of 13C-Depleted HA. To investigate the chemical nature of the TBBPA residues bound to HA, we reacted 13C-TBBPA with birnessite in the presence of 13C-depleted HA, and the bound residues were then characterized by 13C NMR. 13C-depleted HA was used to ensure a weak background 13C signal during 13 C NMR spectroscopy, and 13C-TBBPA was used to greatly enhance the 13C signal of the bound TBBPA residues. The ratio of 13C-TBBPA to 13C-depleted HA and birnessite was thereby also increased to enhance the signal intensity. Briefly, 15 mL of reaction mixture consisting of 100 mg of birnessite and 20 mg of 13C-depleted HA was adjusted to pH 8.0 with diluted HCl or NaOH, followed by the addition of 0.64 μg of 14C-TBBPA (total radioactivity: 2960 Bq) and 3 mg of 13C-TBBPA. The reaction conditions and the treatment of the reaction suspension after the reaction were similar to the reaction of 14 C-TBBPA with birnessite in the presence of the two HA. The final HA pellet was dissolved in 0.1 M anoxic deuterated NaOH (0.6 mL) and then analyzed by solution 13C NMR spectroscopy (see below). Besides this, a control in the absence of 13Cdepleted HA was also conducted for reference. The aqueous supernatant of the reaction suspension was directly extracted by EtOAc in the control. Alkaline Hydrolysis of TBBPA Residues Bound to 13CDepleted HA. To determine the specific amount of ester- and ether-linked bound residues of TBBPA in the 13C-depleted HA after NMR analysis, we conducted alkaline hydrolysis of this HA according to the method described by Martens,31 which was used to recover the phenolic acids from plants and soils. Previous studies32−34 suggested that treatment with 1 M NaOH at room temperature could extract ester-linked phenolic acids, and treatment with 4 M NaOH and heat (microwave digestion or autoclave) could extract total phenolic acids. Martens31 suggested combining these extractions in sequence to distinguish the ester and ether linkage of phenolic acids. In short, the HA pellet was dissolved in 1 M NaOH (4 mL) under a N2 atmosphere and heated for 4 h at 90 °C to cleave the ester bonds.10,31 The HA suspension was extracted with EtOAc (1:1, v/v) after acidification of the suspension to pH 1. The upper layer of the extract was subjected to LSC determination; the results were interpreted as the amount of ester-linked bound residues of TBBPA. The HA pellet was redissolved in 4 M NaOH (4 mL) under a N2 atmosphere and heated for 15 min at 120 °C in an autoclave to cleave the ether bonds.10,31 After acidification and EtOAc extraction of the HA suspension, the radioactivity in the EtOAc extract, representing the content of ether-linked bound residues of TBBPA, was counted by LSC. Next, the remaining HA pellet was separated and freeze-dried for NMR analysis (see below). Determination of Radioactivity. Radioactivity was quantified by LSC (LS6500; Beckman Coulter) with a detection limit of 0.5 Bq. Quenching was corrected by external standards; the chemiluminescence in the samples was negligible. For LSC, the 14CO2 trap solution, a portion of the EtOAc extract, the aqueous solution after EtOAc extraction, and the final HA dissolved in NaOH were each mixed with 3 mL of scintillation cocktail (Gold Star Multipurpose; Meridian Biotechnologies Ltd.) and then counted. HP-GPC−14C-LSC. The molecular size distribution of 14CTBBPA residues bound to the two HA was analyzed using a

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RESULTS AND DISCUSSION Mineralization of TBBPA by Birnessite. In this study, the mineralization of TBBPA, expressed as the δ-MnO2-induced evolution of CO2, was monitored at defined intervals by determining the radioactivity in the 14CO2 trap solution (Figure 1). The time course of CO2 evolution during the reaction (Figure 1B) showed that the mineralization of TBBPA was still in progress after 48 h of incubation. A previous study showed that ring carbons of monophenols are more difficult to be mineralized by birnessite than dihydroxybenzenes.16 In the absence of HA, the δ-MnO2induced CO2 evolution in 48 h accounted for below 2% of the total radioactivity in the studied pH range (5−8), which might be because hydroxyl radicals generated by birnessite were coupled into the benzene ring, forming dihydroxybenzenes, and then leading to ring cleavage.13,19 For example, a dihydroxybenzene (molecular weight = 268) has been identified as one of TBBPA oxidization products by birnessite.9 The low mineralization rate indicates that oxidation by birnessite is not a major pathway for TBBPA mineralization in soils. It has been reported that the mineralization of TBBPA in natural soils was below 1.3% after 200 days of anoxic incubation8 and 19.6% after 143 days of oxic incubation,10 suggesting it is relatively stable in the soil environment even with biotic processes. In our study, mineralization increased slightly as the pH rose from 5 to 6 but then showed a decreasing trend with further increases in pH (Figure 1A) (p < 0.05, i.e., mineralization at pH 6 was statistically significantly higher than that at pH 5 or 7). On one side, higher pH generates deprotonated TBBPA (pKa 7.5 and 8.536), possibly leading to the increase of the mineralization because deprotonated TBBPA is far more reactive than the protonated form.37 On the other side, negatively charged, deprotonated phenols will be repelled by C

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the interaction between TBBPA and HA. The water-soluble radioactivity increased to approximately 5% (Elliott soil HA) and 10% (Steinkreuz soil HA). Moreover, the formation of bound residues of TBBPA (or its transformation products) and the HA was enhanced to approximately 35% (Elliott soil HA) and 25% (Steinkreuz soil HA), i.e., 3- to 17-fold higher than in the absence of δ-MnO2. The results of treatment of TBBPA + δ-MnO2 were showed in Figure S2 as a reference. It was found that the EtOAc-extractable radioactivity in the aqueous supernatant was similar to that with HA present (about 40%), and more than 50% of the total 14C was found sorbed on δ-MnO2, which decreased up to 25% with the presence of HA (Figure S2). These results provided direct evidence that δMnO2 plays an important role in the formation of HA-bound TBBPA residues. δ-MnO2 can induce the oxidative degradation of many organic compounds, especially those with phenol groups, because of the high reduction potential of δ-MnO2.12 Lin et al.9 reported 90% oxidation of TBBPA by birnessite in 60 min. The first step of the reaction was the formation of a phenoxy radical and its resonance forms. Similar results were obtained in this study. After 48 h of reaction, the parent TBBPA molecules were not detectable in the EtOAc extracts (i.e., below 3% of the initially applied TBBPA; see the Supporting Information) of the reaction systems containing: (1) 14C-TBBPA and δ-MnO2 and (2) 14C-TBBPA, HA, and δ-MnO2. This implies that most of the extractable residues of TBBPA at the end of the reaction were TBBPA transformation products. In the presence of HA, the active TBBPA phenoxy radicals induced by δ-MnO2 have a much higher probability for reacting with the humic molecules. Their incorporation into the latter results in the formation of bound residues. Besides, the phenolic moieties of HA also provide the possibilities to produce phenoxy radicals and initiate the oxidation reactions. Compared to TBBPA, the phenoxy radicals of HA are expected to have more complicated resonance forms. High aromaticity of HA will facilitate the stabilization of radicals and have higher probability for reacting with TBBPA. The slight increase in the amount of bound residues with an increase in pH (Figure 2E,F) was not in accordance with the decrease in either δ-MnO2 reduction potential14,47 or TBBPA adsorption capability5 with increasing pH but consistent with the decrease of TBBPA reduction potential, suggesting that deprotonated TBBPA is more active.37 Furthermore, the amount of binding of TBBPA residues to the Elliott soil HA was higher than to the Steinkreuz soil HA (Figure 2E,F). This was probably due to the much higher aromaticity of the Elliott soil HA than the Steinkreuz soil HA (50% versus 7%) because phenoxy radicals may readily react with the aromatic moieties of HA, which help the stabilization of radical resonance.16 In the absence of δ-MnO2, the fraction of bound residues of TBBPA was small (below 8% for Elliott soil HA and below 5% for Steinkreuz soil HA). Similarly, Cecchi et al. showed that sorption of phenolic acids was dramatically reduced in soils from which metal oxides were removed.48 It has been reported that dihydroxybenzenes (such as catechol) can form large amounts of bound residues within HA due to its high autoxidation potential, which means that catechol can be easily oxidized by O2 in the atmosphere.16 The autoxidation of TBBPA was unlikely because, as shown in a previous study, even in the presence of δ-MnO2 under oxic conditions, the oxidation of TBBPA does not result in the formation of polymerization products (such as dimers and tetramers).9

Figure 1. Mineralization of TBBPA (expressed as CO2 evolution) by birnessite (δ-MnO2) in the absence and presence of Elliott soil HA or Steinkreuz soil HA as a function of pH in 48 h (A) and time at pH 5 (B). Values are the means and standard deviations of two separate experiments.

the surface of δ-MnO2 (isoelectric point 2.538 and site density 3.1 × 10−5 mol m−2)39, resulting in less adsorption and less oxidation. Moreover, the reduction potential of δ-MnO2 decreases with increasing pH,12 which is also the case for TBBPA. Thus, in general, the differences in TBBPA mineralization at different pH values are an integrated result of these various effects. In previous studies, HAs were shown to either promote, inhibit, or not affect the oxidation of organic chemicals by Mn oxides.12 The overall effect of HA on chemical oxidation may be specific to the nature and concentration of the involved compounds.12 The inhibitory effect of HA is mostly due to their competition with organic chemicals for sorption sites40,41 and their reductive dissolution of MnO2.42,43 Complexation of HA with Mn(II) in solution can prevent its competitive adsorption on the oxide surface, thus promoting the oxidation rate.44,45 In this study, the presence of both HAs had almost no effect on the mineralization of TBBPA except for slightly enhanced mineralization achieved with Elliott soil HA at pH 5 (p < 0.05) (Figure 1A). This finding is similar to that of a previous study, in which HA did not alter the degradation rate of bisphenol A by δ-MnO2.46 Binding of TBBPA Residues to HA. After 48 h of incubation, the TBBPA residues in the HA solution were fractionated by EtOAc extraction, which yielded EtOAcextractable, water-soluble, and bound (EtOAc unextractable) residues. In the absence of δ-MnO2, approximately 90% of the residual radioactivity could be extracted by EtOAc, versus below 4% of the water-soluble residues and below 8% of the residues retained in the HA pellets (EtOAc unextractable) (Figure 2), indicating a weak interaction between TBBPA and HA as a whole. However, in the presence of δ-MnO2, the EtOAc-extractable radioactivity decreased significantly to approximately 40%, suggesting that MnO2 greatly changed D

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Figure 2. Ethyl acetate extractable residues (A,B), water-soluble residues (i.e., radioactivity in the acidic aqueous supernatant after the extraction) (C,D), bound residues (E,F) of TBBPA in Elliott soil HA or Steinkreuz soil HA after a 48 h incubation in the absence and presence of δ-MnO2. Values are the mean and standard deviations of two separate experiments.

Therefore, the TBBPA bound residues that formed in the absence of δ-MnO2 may have resulted from the sequestration of TBBPA molecules within the HA.49 However, it should also be noted that sample treatment after the reaction, including the acidification of the HA, may have resulted in a coil structure of the HA50 and thus led to TBBPA sequestration artifacts. Another possible effect of the absence of δ-MnO2 was the formation of water-soluble residues of TBBPA due to its binding to the HA fraction dissolved in the acidic supernatant (evident by its brown color after EtOAc extraction). In the presence of δ-MnO2, the fraction of water-soluble residues of TBBPA increased significantly for both HAs, implying that besides HA-bound residues, TBBPA transformation products of

smaller size and higher polarity were formed. In the study of Lin et al.,9 two products with a single benzene ring were predominant in the reactions. Similar results51 showed that the major MnO2-mediated bisphenol A transformation product was also with a single benzene ring. Our observation of an increase in the amount of water-soluble residues with increasing pH may also have reflected the greater reactivity of deprotonated TBBPA than the protonated one.37 Molecular Size Distribution of TBBPA Residues Bound to HA. HP-GPC−14C-LSC was used to investigate the molecular size distribution of TBBPA residues bound to the two HA after their reaction in the presence of δ-MnO2 (Figure 3). Because the carbon content of TBBPA residues was below E

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HA,35 such as phenolic or carboxylic groups, may also lead to the different distribution of TBBPA bound residues in HA. Chemical Nature of the Bound Residues. To gain insight into the chemical nature of the HA-bound TBBPA residues, we performed a similar reaction containing 13CTBBPA, 13C-depleted HA, 14C-TBBPA, and δ-MnO2 at pH 8. The bound residues that formed with the 13C-depleted HA were further analyzed by solution 13C NMR spectroscopy (Figure 4). In this particular study, 13C-depleted HA is superior

Figure 3. Gel-permeation chromatography−liquid scintillation counting spectra of TBBPA residues bound to Elliott soil HA (A) or Steinkreuz soil HA (B) after the reaction of TBBPA with δ-MnO2 at pH 6 for 48 h. The spectra were obtained by combining two parallel samples.

0.01% of the total carbon content of the HA, the UV signal was believed to be all contributed by HA, and the 14C radioactivity corresponds to the TBBPA bound residues. The GPC spectra indicated that TBBPA residues bound to all size fractions of each HA because the measured 14C radioactivity was distributed over the whole absorbance range (Figure 3). Moreover, the GPC spectra showed that the molecular size distribution of TBBPA residues bound to the Elliott soil and Steinkreuz soil HA differed from each other, which implied that HA from different sources differed in their reactivity with TBBPA. The mismatch between the UV absorbance signals of the Elliott soil HA and the 14C radioactivity (Figure 3A) suggested that the distribution of HA-bound TBBPA residues was not uniform. In contrast, the UV absorbance signals of the Steinkreuz soil HA paralleled those of the 14C radioactivity, which indicated the uniform distribution of the bound TBBPA residues within this HA (Figure 3B).16,35,52 In this study, higher radioactivity in larger HA fraction was observed in the Steinkreuz soil HA, and reverse distribution was found for the Elliott soil HA. The discrepancy in the molecular size distribution of the TBBPA bound residues may be attributed to the inherent differences of the two HAs, such as the size distribution of aromaticity or active reaction sites. Because, on the one hand, high aromaticity of HA may facilitate bound residue formation,16 the different distribution of aromaticity over HA fractions with various molecular weights between the two HAs may contribute to the difference in GPC spectra. On the other hand, the active functional groups of HA are important in oxidative coupling reactions mediated by δMnO2 between TBBPA and HA (see the next section). Thus, the different size distribution of the active reaction sites of

Figure 4. 13C NMR spectra of 13C-depleted HA (A), 13C-TBBPA (B), 13 C-TBBPA residues in the supernatant of the reaction suspension after reaction of 13C-TBBPA with δ-MnO2 in the absence of HA at pH 8 for 48 h (C), 13C-TBBPA residues bound to the 13C-depleted HA after reaction of 13C-TBBPA with δ-MnO2 at pH 8 for 48 h (D), and 13 C-TBBPA bound residues within the 13C-depleted HA after alkaline hydrolysis (E). The solvents used for the analyses were 0.1 M deuterated NaOH (A), deuterated chloroform (CDCl3) (B,C), and 0.1 M deuterated NaOH (D,E).

to the Elliott and Steinkreuz HA because it avoids the problem of their high natural 13C abundances.16,53 In addition, 13CTBBPA was used to enhance the 13C signal of the TBBPA bound residues. The 13C NMR spectrum of the 13C-depleted HA had a low 13C-background, with the sole signal at 160 ppm probably originating from carbonate16 (Figure 4A). A total of four types of aromatic carbon were evident in the spectrum of pure 13C-TBBPA (Figure 4B): C2 (the carbon numbered 2 in the spectrum), C3, C4, and C1, with chemical shifts of 109.8, 130.3, 144.3, and 147.6 ppm, respectively. No obvious signals from the alkyl carbon occurred in this spectrum because 13CTBBPA was ring-labeled. The spectrum of the residues of 13CTBBPA in the aqueous phase of the reaction suspension of δMnO2 in the absence of HA showed two adjacent peaks around 130 ppm (Figure 4C) that were significantly different from the spectrum of pure 13C-TBBPA, which may be contributed by one single-ring transformation product of TBBPA mediated by δ-MnO2 (Figure S3B) according to the result of calculation of F

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Figure 5. Proposed transformation mechanisms of TBBPA oxidized by δ-MnO2 in the presence and absence of humic substances. TBBPA can be oxidized by δ-MnO2 to aromatic products, mainly those with a single benzene ring, as proposed previously;9 the reaction is accompanied by the release of trace amounts of CO2. In the presence of humic substances, TBBPA residues are bound to the various size fractions of humic substances. The benzene-ring structure of TBBPA is for the most part preserved. The mechanism includes the sequestration of TBBPA (or its transformation products and intermediates) within the humic substances and the formation of ester and ether bonds or a C−C bond between these components (Figures S3 and S4).

Alkaline hydrolysis was applied to distinguish between ester and ether bonds in the bound residues of TBBPA in 13Cdepleted HA.10,31 Briefly, treatment with 1 M NaOH at 90 °C was used to cleave the operationally defined ester bonds. Then sequential treatment with 4 M NaOH at 120 °C was used to cleave the operationally defined ether bonds.10,31 The results showed that ester- and ether-linked bound residues accounted for 32.7% and 4.3%, respectively, of the total TBBPA bound residues (quantified by 14C radioactivity; see the Materials and Methods section). This to some extent supported the 13C NMR spectroscopy findings and implied that TBBPA might preferentially form ester rather than ether bonds with HA. A newly published report also suggested the predominance of ester over ether linkages in bound residues of TBBPA within an oxic sandy soil.10 However, hydrolysis with 4 M NaOH may not completely break all of the ether bonds, which would lead to an underestimation of their contribution to the formation of TBBPA bound residues. After alkaline hydrolysis, the 13C-depleted HA was further analyzed by 13C NMR (Figure 4E). It was found that five peaks with the same chemical shifts as that in Figure 4D still existed in Figure 4E, and the signal intensities of all of the peaks were much weaker than those in Figure 4D, indicating that each peak in Figure 4D was contributed by both ester and ether bonds formed in the TBBPA bound residues with 13C-depleted HA. The remaining 13C signal in Figure 4E suggested that other mechanisms, such as other covalent bonds (e.g., C−C bond or metal bridges between compounds and humic molecules) or sequestration by HA, play important roles in bound-residue formation. For example, in the study of Lin et al., the products of TBBPA oxidation by δ-MnO2 contained a C−C bond.9 In addition, Mn (II), which is released from δ-MnO2 into the solution by reductive dissolution (Table S2), might act as a

chemical shifts (see below). The spectrum of the bound residues within 13C-depleted HA formed in the presence of δMnO2 (Figure 4D) contained five peaks, with the predominant signals located at 114, 118, 130, 145, and 156 ppm. A comparison of panels B and C of Figure 4 shows that in the majority of the HA-bound residues, the benzene ring remained largely intact. In the chemical bonding of organics with HA to form bound residues, the most frequently discussed bonding mechanisms are the formation of ester or ether bonds.10,16,31,54,55 In this study, we used MestRenova (version 9.0, Mestrelab Research S.L.) to predict the possible bound structures of TBBPA on the basis of the 13C NMR spectra.54,56 The predicted C shifts for the possible degradation products and bound residues are shown in Figure S3. MestRenova could accurately predict the chemical shifts of aromatic carbons of TBBPA (Figures S3A and 4B). In the presence of HA, the chemical shifts of TBBPA bound to HA through ester and ether bonds predicted (Figure S3D,E) coincided well with the signals in Figure 4D. In addition, two main transformation products (Figure S3B,C)9 might be incorporated into HA. The possible chemical shifts of the two transformation products and the cationic intermediate bound to HA via an ether bond (Figure S3F,H,J) or an ester bond (Figure S3G,I,K) were also in relatively good agreement with the signals in Figure 4D. Oxidation by δ-MnO2 may also lead to the formation of a single-ring radical of TBBPA9 that then binds to a radical of the humic molecules in a reaction induced by δ-MnO2 via C−C bonding (Figure S3L), which is in good agreement with the 13C NMR result (Figure 4D). Thus, overall, the δ-MnO2-induced formation of an ester, ether, or C−C bond between TBBPA bound residues and HA can be reasonably assumed. G

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soils. Actually, similar phenol oxidation reactions have been observed in soils containing manganese oxides.64,65 Organic contaminants covalently bound to humic substances are integral components of SOM,66,67 and they therefore contribute to detoxification of the contaminants.68 However, one should note that ester bonds, such as those involved in the formation of TBBPA bound residues, might be hydrolyzed, leading to the release of HA-bound TBBPA under certain conditions. In addition, bound residues of TBBPA might also be reintroduced into the soil solution and eventually assimilated by plants due to the activity of microorganisms, alteration of soil redox conditions, agricultural practices, or the introduction of certain chemicals that change the chemistry of the soil.8,49,55

bridge ion to facilitate the binding of phenolic compounds to HA. However, this mechanism should have played only a minor role in our system because (1) our previous study showed that divalent metals (such as Cd (II), Cu (II), and Pb (II)) had a weak capability to facilitate TBBPA adsorption onto a Ferrosols soil;5 and (2) prior to the fractionation of the bound residues, the samples had been acidified to pH 1, such that any TBBPA residues bound to HA via metal bridges should have been dissociated. Physical sequestration by HA is another important mechanism in the formation of bound residues of organic contaminants.49 In fact, in both HAs, a small fraction of the bound residues of TBBPA formed in the absence of δ-MnO2 (Figure 2E,F). This was probably the result of the sequestration of TBBPA within the HA. In the presence of δ-MnO2, smaller transformation products than TBBPA induced by its oxidative transformation may have more easily diffused into microsites of HA.57,58 Thus, the remaining 13C signal in Figure 4E may result from the binding of residues of TBBPA with 13C-depleted HA via C−C bonds (Figure S3L) or the physical entrapment of TBBPA or its products within the HA (Figure S3A−C). On the basis of the results of the 13C NMR spectra and the alkaline hydrolysis of HA, a mechanism for the oxidative binding of TBBPA to HA can be proposed (Figure 5). The formation of phenoxy radicals is the initiating reaction during the δ-MnO2-induced oxidation of TBBPA.9,14,46,59−61 The phenoxy radical can also undergo β-carbon scission to form a phenyl radical and a cationic intermediate with a single benzene ring.10 In the presence of HA, within a large pH range, these radicals and the cationic intermediate more readily react with HA to incorporate themselves into HA via chemical bonding (for example, through the formation of ester bonds between the carboxylic groups of HA and the phenoxy radial or the cationic intermediate of TBBPA)62 or ether bonds between the phenolic groups of HA and the phenoxy radial or the cationic intermediate,22 or by other covalent bonds, such as C−C bonds (detailed reactions have been proposed in Figure S4). Meanwhile, the phenolic moieties of HA may also initiate the reaction by formation of phenoxy radicals. More binding of TBBPA residues to Elliott soil HA with higher aromaticity than to Steinkreuz soil HA probably reflects that the high aromaticity of HA can assist the stabilization of radicals, thus facilitating the formation of the bound residue of TBBPA. Of course, bound residue may also be formed by HA sequestration. Meanwhile, the dissolution of transformation products of TBBPA induced by MnO2 in water increased because of their increased solubility. The mineralization of TBBPA was hardly influenced by either Elliott soil HA or Steinkreuz soil HA, probably due to the low mineralization of TBBPA by δ-MnO2 alone (Figure 1A).

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ASSOCIATED CONTENT

* Supporting Information S

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.5b06265. Additional details on the HPLC analysis method of parent TBBPA, chemical properties of HAs (Table S1), dissolved Mn concentrations in aqueous solution after reaction with TBBPA (Table S2), X-ray diffraction of birnessite (Figure S1), distribution of 14C radioactivity in the reaction suspension of MnO2 with TBBPA in the absence and presence of HA (Figure S2), postulated structures of the interactions of TBBPA or its transformation products with HA and the expected chemical shifts (Figure S3), and possible pathways for the formation of an ester or ether bond or a C−C bond between HA and TBBPA (Figure S4). (PDF)

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AUTHOR INFORMATION

Corresponding Author

*Tel and fax: +86 25 89680361; e-mail: [email protected]. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The authors thank the Natural Science Foundation of China (grant nos. 21237001, 21577062, 21277068, 21177057, and 41203062) for financial support.

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REFERENCES

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ENVIRONMENTAL IMPLICATIONS This study demonstrated that δ-MnO2 is able to significantly enhance the binding of TBBPA to soil humic substances by 3to as much as 17-fold over a relatively broad pH range. It also provided molecular-level information on the chemical nature of the bound residues. The high aromaticity of humic substances facilitates bound-residue formation. Our results suggest that in oxic soils rich in Mn oxides, the formation of bound residues may contribute significantly to the removal of TBBPA. Under alkaline soil conditions, TBBPA is relatively mobile,5 or it may be assimilated by plants.7,63 However, the δ-MnO2-stimulated formation of TBBPA bound residues occurs even at pH 8, thereby decreasing the mobility and bioavailability of TBBPA in H

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