Chemosphere 125 (2015) 86–93
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Transformation of bisphenol A in water distribution systems: A pilot-scale study Cong Li a, Zilong Wang a, Y. Jeffrey Yang b, Jingqing Liu a, Xinwei Mao a, Yan Zhang a,⇑ a b
College of Civil Engineering and Architecture, Zhejiang University, Hangzhou, China U.S. EPA, National Risk Management Research Laboratory, Cincinnati, USA
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
The halogenations of BPA in a pilot-
Cl CH 3 OH
HO
scale water distribution systems was studied. The chlorinated mechanism of BPA in water distribution systems was proposed. THMs and HAA formation in BPA chlorinated in water distribution systems were found.
CH 3 Br
minor
Cl CH 3 OH
HO CH 3
chlorination in WDS
CH3
Cl
OH
HO
major
Cl
CH3
stage 1
Cl
CH 3 OH
HO CH 3 Cl Cl
Cl CH 3
OH
HO CH 3 Cl
Cl
Cl CH 3 HO
OH CH 3
Cl
Cl
R
CH 3
stage 2
OH
HO Cl
CH 3 R
CH3 HO
Cl
OH CH3 Cl
halogenation
benzene ring opening
CHCl3 CHBrCl2 CHBr2Cl
stage 3
CHBr3 HAAs
a r t i c l e
i n f o
Article history: Received 6 August 2014 Received in revised form 18 November 2014 Accepted 22 November 2014 Available online 27 December 2014 Handling Editor: Shane Snyder Keywords: Bisphenol A Water distribution system Halogenations Mechanism
a b s t r a c t Halogenations of bisphenol A (BPA) in a pilot-scale water distribution system (WDS) of a cement-lined ductile cast iron pipe were investigated. The water in the pilot-scale WDS was chlorinated with a free chlorine concentration of 0.7 mg L1 using sodium hypochlorite, and with an initial BPA concentration of 100 lg L1 was spiked in the WDS. Halogenated compounds in the BPA experiments were identified using EI/GC/MS and GC. Several BPA congeners, including 2-chlorobisphenol A (MCBPA), dichlorobisphenol A (D2-CBPA), 2,20 ,6-trichlorobisphenol A (T3CBPA), 2,20 ,6,60 -tetrachlorobisphenol A (T4CBPA), 2-bromobisphenol A (MBBPA), and bromochlorobisphenol A (MBMCBPA) were found. Moreover, further halogenation yielded other reaction intermediates, including 2,4,6-trichlorophenol (T3CP), dichlorobisphenol A, bromodichlorophenol, and dibromochlorophenol. After halogenation for 120 min, most of the abovementioned reaction intermediates disappeared and were replaced by trihalomethanes (THMs). Based on these experimental findings, the halogenation process of BPA oxidation in a WDS includes three stages: (1) halogenation on the aromatic ring; (2) chlorine or bromine substitution followed by cleavage of the a-C bond on the isopropyl moiety with a positive partial charge and a b0 -C bond on the benzene moiety with a negative partial charge; and (3) THMs and a minor HAA formation from phenolic intermediates through the benzene ring opening with a chlorine and bromine substitution of the hydrogen on the carbon atoms. The oxidation mechanisms of the entire transformation from BPA to THM/HAA in the WDS were proposed. Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction ⇑ Corresponding author. Tel./fax: +86 571 87952015. E-mail address: [email protected] (Y. Zhang). http://dx.doi.org/10.1016/j.chemosphere.2014.11.047 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.
Bisphenol A (BPA) is a well-known endocrine disruption chemical. In despite of relatively low estrogen receptor binding activity
C. Li et al. / Chemosphere 125 (2015) 86–93
in vitro (Hu and Aizawa, 2003), the relatively high estrogenic activity was observed in mouse fetuese (500-fold greater than that of octylphenol) (Nagel et al., 1997). BPA can inhibit the development of the reproductive, embryo and nervous system leading to apparent health concerns. Animal experiments show that BPA can cause leukaemia and increase the lymphoma rate. BPA has attracted more attention recently and is listed in the EPA’s drinking water contaminant candidate list (CCL) (Li et al., 2005; Li et al., 2008; Deborde and von Gunten, 2008; Vikesland et al., 2013; U.S. EPA. Bisphenol-a Action Plan). BPA has been used in high quantities in the manufacture of plastics and beverage cans and may be leaked to the environment in the processes of production, transportation and usage. Therefore, BPA has frequently been detected in surface waters, drinking water, and in the tissues of animals worldwide (Hendriks et al., 1994; Azevedod et al., 2001; Bolz et al., 2001; Benotti et al., 2009; Klecˇka et al., 2009;Kleywegt et al., 2011; Chen et al., 2013). In China, various concentrations of BPA were found in the Shanghai Huangpu River (Ma et al., 2006), the Pearl River basin (Li et al., 2010), the Qiantang River (Zhang et al., 2004), and water from water treatment processes and in water distribution pipes (Zhao and Guo, 2007; Shao et al., 2008; Li et al., 2010). Furthermore, the chlorinated byproducts of BPA in drinking water were firstly reported in the recent paper (Fan et al., 2013). A number of investigations of the chlorinated reaction in beaker experiments were conducted using a purified water matrix. In a beaker experiment, Hu et al. (2002) found that chlorobisphenol A, 2,4,6-trichlorophenol and some polychlorinated phenoxyphenols (PCPPs) were the main byproducts. Jensen et al. (1999) found that amide nitrogens, such as that contained in the acetaminophen structure, were not favourably N-chlorinated. Gallard and von Gunten (2002) found that yields of chloroform formation depends on the concentration of phenol. The halogenation of BPA in an actual WDS is affected by more factors than that in a beaker experiment, including drinking water disinfectants, natural organic matter (NOM), inorganic ions (such as Br and Fe3+), pipe scales, and pipe wall chlorine demand among others. Moreover, there is no report about the by-products difference of BPA halogenations in the presence of Br- in the water distribution systems. Therefore, the halogenations and degradation of BPA in a WDS was studied in this paper. In this study, we report our experimental results on the BPA halogenation for a pilot-scale WDS and analyse the degradation mechanisms of BPA in a WDS. Moreover, in the presence of Br-, the products of BPA halogenations in WDS were firstly reported. 2. Materials and methods 2.1. Materials BPA experiments were performed on a pilot-scale water distribution system (WDS) (see Supplementary Information Fig. 1). This pilot-scale water distribution system has one loop made of cement-lined ductile cast iron pipe used in this study. Each loop is approximately 80 m in length and 150 mm in diameter. Water used in the experiments came from the municipal drinking water network on the Zhejiang University campus in Hangzhou City and was circulated in the loop. Results of water quality measurements, monitored every 2 d (see in Supplementary Information Table 1). A mixture of four THMs standards including trichloromethane (TCM), bromodichloromethane (DCBM), dibromochloromethane (DBCM), tribromomethane (TBM) in methyl tert-butyl ether (MTBE) containing 2000 lg mL1 was purchased from Supelco(Supelco Park, PA, USA). A mixture of six haloacetic acids (HAAs) standards including monochloroacetic acid(MCAA), dichloroacetic acid(DCAA), trichloroacetic acid(TCAA), monobromoacetic acid
87
(MBAA), dibromoacetic acid (DBAA) and bromochloroacetic acid (BCAA) in methyl tert-butyl ether(MTBE) containing 2000 lg mL1 was purchased from Supelco(Supelco Park, PA, USA). Water samples were collected using a device equipped with a 12-position vacuum manifold set(CNW); CNW HC-18 SPE cartridges; pressured gas blowing concentrators (ND200, Shanghai Joyn Electronic CO., Ltd.); EI/GC/MS(GC2010 Plus, MS-QP2010 SE, SHIMADZU); Gas chromatography (GC)(450, VARIAN); and a DR2800 ultraviolet spectrophotometer (HACH). Chemicals and agents used in the experiments included sodium thiosulfate (CP, the Chinese medicine group chemical reagent co., LTD); a free chlorine powder pillow reagent (HACH); and BPA, sodium hypochlorite, acetonitrile, nitric acid, and acetonitrile(AR, Aladdin). 2.2. Experimental methods To identify BPA halogenation in a WDS, two parallel experiments were performed. One set of experiments, denoted as ‘‘BPA runs’’, was based on BPA-spiked water. The other set, denoted as ‘‘blank runs’’, was performed in identical experimental conditions but without the addition of BPA. Therefore, the effects on BPA halogenation and the byproducts of BPA chlorination were identified by comparison. 2.2.1. Halogenation experiment BPA halogenations in chlorinated drinking water were tested in the A loop of the WDS. Before the experiments, the loop was flashed with chlorinated drinking water for 20 min. The water temperature and flow velocity in the WDS were adjusted using automated controls. The experiment conditions were stabilised at pH = 7.3 ± 0.3, a flow velocity of 1.0 m s1, a water temperature of 25 °C ± 1 °C, and an initial free chlorine concentration of 0.7 mg L1 adjusted using a sodium hypochlorite solution. The initial concentration of BPA in this study was 100 lg L1. While water was circulated in the WDS, 1 L samples were taken at 10 min, 30 min, 60 min, 90 min, 120 min, 240 min and 24 h. To these samples were added 2 mL of a 100 mg L1 sodium thiosulfate solution to stop the reaction. 2.2.2. Sample analysis 10 mL water samples were taken from the 1 L samples collected for GC analysis. Before the GC analysis, BPA and its reaction byproducts were extracted by solid-phase extraction (SPE). After the pH adjustment to 3.5 by adding 3 mL HNO3 (0.32 M), the samples passed through the C18 cartridges at a rate of 5 mL min1 using the CNW 12-position vacuum manifold set. Then, the SPE cartridges were dried by passing a gentle nitrogen stream for 2 min. Finally, BPA and its byproducts were eluted using 5 mL acetonitrile, and the effluents were concentrated with a nitrogen stream for approximately 70 min to be analysed by EI/GC/MS. Most byproducts of the BPA halogenation in the WDS were detected by the EI/GC/MS (SHIMADZU, GCMS-QP2010 SE), equipped with a GC splitless injector operating at 260 °C and with a HP-5 MS capillary column (30 m 0.25 mm 0.25 lm, Agilent). Helium was used as a carrier gas with a 1 mL min1 flow rate. The column temperature was raised from 60 °C to 280 °C at a rate of 10 °C min1 and then maintained for 8 min. The mass spectrometer with electron ionisation (70 eV) was operated at 180 °C in an ionisation chamber with a scan range from m/z 50 to 500. THMs and HAAs were detected by GC (Varian, GC-450) with a splitless injector and a SP-Sil 8 CB capillary column (15 m 0.25 mm 0.25 lm, Varian). Headspace injection of THMs was used at 150 °C in the injectors. Column flow was 1 mL min and the temperature of the ECD detector was 290 °C for THMs detection. Headspace injection of HAAs was used at 175 °C in the injectors. The column temperature was raised from 40 °C to 140 °C at a rate of 10 °C min1, from
88
C. Li et al. / Chemosphere 125 (2015) 86–93
140 °C to 190 °C at a rate of 25 °C min1 and then maintained for 8 min at 40 °C and for 0.5 min at 190 °C. Column flow was 1 mL min1 and the temperature of the ECD detector was 300 °C for HAAs detection. 3. Results and discussion 3.1. Formation of BPA congeners Compared to the blank runs (Fig. 1a), the total ion chromatogram (TIC) for the water samples showed that additional nine compounds generated in the BPA halogenations. As shown in Fig. 1b, the new peaks numbered 1–9 were absent in the TIC of the blank samples. The samples’ mass spectra are shown in Fig. 2a–i. Peak 1 in Fig. 1b at the retention time of 19 min is the same as the pure BPA solution from Aladdin. Peak 2 is MCBPA with a retention time of 19.42 min, which is approximate to the BPA peak. The spectra also show a molecular ion (M+) at m/z 262 and a fragment ion at m/z 247. This ion is assigned to C14H12ClO+2 formed by MCBPA losing a methyl. Peaks 3–5 have mass spectra similar to that of BPA and MCBPA. The molecular ions of these peaks were m/z 296, 330, and 366 with retention times of 21.25 min, 21.45 min, and 23.05 min, respectively. These peaks are assigned as DCBPA, T3CBPA and T4CBPA. The mass spectra of Peak 6 show a molecular ion of m/z 307, which is 79 more than that of BPA, with a retention time between MCBPA and DCBPA. Peak 6 is identified as the compound bromobisphenol A (MBBPA). Similar to MCBPA, the fragment ion at m/z 292 is assigned to C14H12BrO+2, which is formed by MBBPA losing a methyl functional group. Peak 7 has a molecular ion at m/z 341, approximately m/z 34 more than that of MBBPA, which is designated as bromochlorobisphenol A (MBMCBPA). The retention time of Peak 1 is 10 min, and its mass spectra are consistent with that of pure 2, 4, 6-trichlorophenol (T3CP). The retention time of Peak 4 is 13.45 min, and its molecular ion is m/z 220. The fragment ion at m/ z 205 is generated by losing CH3, which is consistent with the a-C
on the isopropyl of BPA that can cleave when it reacts with HClO (Hu et al., 2002). In this study, the BPA congeners with more than two bromine atoms were not found due to low concentrations of bromide ion in the drinking water (26 lg L1). In the first 60 min of the halogenation, BPA congeners were the main reaction byproducts (Fig. 3). Over 95% of the BPA was degraded in the first 10 min. The concentration of BPA congeners reached its highest concentration at approximately 10 min and then decreased. Most of the BPA congeners disappeared after 60 min. These experiment results show that in the first 60 min (especially the first 30 min), BPA halogenation in the WDS is the stage of chlorine-substitution and brominesubstitution. 3.2. Formation of halogenated phenols Hu et al. (2002) showed that the a-C bond on the isopropyl of BPA can cleave in the chlorination reactions. The benzene moiety is then substituted with chlorine, and the isopropyl moiety is substituted by a hydroxyl. The chlorine oxidation reaction is consistent with our experimental results. However, in this study, the drinking water in Hangzhou City included chlorine and bromine ions. The mass spectra of three new peaks (Peak 2, Peak 3 and Peak 5) are shown in Fig. 4. Fig. 4 shows that the mass spectra for Peak 1 with a retention time at 10 min are consistent with that of T3CP. For Peaks 2 and 3, their retention time is 11.19 min and 12.38 min, respectively. In Fig. 4a, the molecular ion of Peak 2 is m/z 241. The relative masses of Br and Cl are 80 and 35, respectively. Their difference, 45, is consistent with the Br substitution of the Cl atom in T3CP, upon which Peak 2 is deduced to be bromodichlorophenol. In addition, the molecular ion of Peak 3 with m/z 286 is dibromochlorophenol. Peak 5 is m/z 266, and the fragment ion appears at m/z 251. Its relative molecular weight is 46 more than the molecular weight of D2CBPA (C9H10Cl2O2), but is very similar to the relative
Fig. 1. (a) Total ion GC–MS chromatogram (TIC) of blank sample. Scan range: 0–25 min; (b) the TIC of BPA halogenation in WDS at 10 min. Scan range: 0–25 min.
C. Li et al. / Chemosphere 125 (2015) 86–93
89
Fig. 2. Mass spectra of BPA halogenation byproducts in WDS at 10 min. (a) BPA (peak 1 in Fig. 1); (b) MCBPA (peak 2 in Fig. 1); (c) DCBPA (peak 3 in Fig. 1); (d) T3CBPA (peak 4 in Fig. 1); (e) T4CBPA (peak 5 in Fig. 1); (f) MBBPA (peak 6 in Fig. 1); (g) MBMCBPA (peak 7 in Fig. 1); (h) T3CP (peak 8 in Fig. 1); (i) C9H10Cl2O2 (peak 9 in Fig. 1).
90
C. Li et al. / Chemosphere 125 (2015) 86–93
Fig. 5 shows that from 30 min to 60 min, with further halogenation, the a-C bond on the isopropyl of BPA cleaves to form a large amount of halogenated phenolic compounds. This distinct BPAoxidation stage is labelled as being the halogenated phenols formation. 3.3. Formation of trihalomethanes (THMs) and halogenated acetic acids (HAAs)
Fig. 3. The variation of BPA congeners with time.
mass difference between Br and Cl. As its retention time at 14.5 min is approximate to that of D2CBPA(13.45 min), Peak 5 is interpreted to be MBMCBPA (C9H10BrClO2). With the concentration of BPA congeners decreasing, a large amount of halogenated phenol appeared. The halogenated phenols marked as T3CP reached a maximum concentration at 60 min after reactions in the WDS, but mostly disappeared at 120 min. The changes in their concentration with time are shown in Fig. 5.
BPA chlorination leading to the formation of THMs is obvious by comparing water samples from the BPA chlorination and the water samples with no BPA chlorination. The results are shown in Fig. 6a. In Fig. 6a, the differences in the THMs formation between the BPA chlorination and the blank runs were very small in the first 120 min. The difference becomes significant in the final stage (120–240 min), and the trichloromethane (CHCl3) formation is the largest. The formation rate of trichloromethane minimally increases with the presence of more bromine atoms in the THMs formation. This is apparently related to trace bromine concentration, which is a factor that limits the reaction rate in the drinking water. After 120 min, the rate of the chloroform (CHCl3) and bromodichloromethame (CHBrCl2) formation suddenly increased. Approximately 6.5 lg L1 CHCl3 formed in this stage compared to 14.5 lg L1 that formed at 120 min (Fig. 6a). This increase is consistent with the appearance of THMs in the TIC spectra, indicating that the THMs formation arises from halogenated phenols. The concentrations of the CHCl3, CHBrCl2, CHBr2Cl and CHBr3 compounds were stabilised at 21.84 lg L1, 11.02 lg L1, 6.05 lg L1 and 1.75 lg L1, respectively, after 24 h of chlorination reaction.
Fig. 4. (a) TIC of BPA halogenations in WDS at 60 min. Scan range: 0–25 min; (b) mass spectra of bromodichlorophenol (peak 2 in Fig. 5a); (c) dibromochlorophenol (peak 3 in Fig. 5a); (d) C9H10BrClO2 (peak 5 in Fig. 5a).
C. Li et al. / Chemosphere 125 (2015) 86–93
91
The disappearance of halogenated phenols accompanied by the formation of THMs, particularly CHCl3 and CHBrCl2, strongly indicated a transformation of the phenolic compounds to halogenated methanes through halogenation. In the BPA chlorination reaction, halogenations open the benzene ring with a cleavage of the C–C bonds, followed by a chlorine and bromine substitution for hydrogen on the H–C bonds. Haloacetic acids (HAAs), being the disinfection byproducts, were detected at lower concentrations. Both THM and HAA formations through phenol halogenations (e.g., chlorination) have been reported in published experiments and synthesis (Gallard and von Gunten, 2002; Rule et al., 2005; Ge et al., 2006; Liu et al., 2006; Guo and Lin, 2009; Zhang and Liu, 2010). 3.4. BPA chlorination sequence and implications
Fig. 5. The variation of halogenated phenols with time.
Fig. 6. The concentration curve of THMs with time. The concentration curve of HAAs with time.
Three types of HAAs, including dichloroacetic acid (DCAA), trichloroacetic acid (TCAA), and bromoacetic acid (MBAA), are shown in Fig. 6b. Fig. 6b shows that the HAA formation had a similar regulation as that of the THMs. After 120 min, the rate of HAA formation increased suddenly in the BPA chlorination reaction (Fig. 6b). Moreover, in all of the HAAs, the concentration of DCAA had the greatest increase. After 24 h of reaction time, the final concentrations of DCAA, TCAA and MBAA were stabilised at 6.58 lg L1, 2.29 lg L1 and 3.34 lg L1, respectively.
BPA transformation in pure water with disinfection oxidants has been reported (e.g., Hu et al., 2002; Deborde and von Gunten, 2008; Gallart-Ayala et al., 2010; Vikesland et al., 2013). In the WDS, the reaction pathways and byproducts are different. A BPA oxidation pathway is proposed in Fig. 7 based on the sequential appearance of the byproducts and their mass spectra identifications, as discussed in earlier sections. Under the experimental conditions, the sequential transformation process can be divided into three stages. Stage 1 from 0 min to 60 min is the BPA congener generation. Electrophilic substitution by chlorine to the phenol moiety occurs at ortho and para, namely the 3, 5 or 30 , 50 positions of the phenol moiety (Deborde and von Gunten, 2008; Vikesland et al., 2013). These types of mono-, di-, tri-, and tetra-halogenations are shown in Fig. 7. Hu et al. (2002) detailed the BPA chlorination pathways and mechanisms in beaker tests. In this study, in the WDS, BPA degraded over 95% in the first 10 min. Chlorine-substitution and bromine-substitution in the phenolic moiety on the benzene ring are the primary mechanisms yielding chlorobisphenol A and bromobisphenol A. Their concentrations reached the maximum at approximately 10 min and decreased to almost zero at approximately 60 min. Stage 2 from approximately 30 min to 120 min is the formation of halogenated phenols. It overlaps with Stage 1 in which, as previously mentioned, a small amount of T3CP formed at approximately 30 min. In the phenol formation, the a-C bond on the BPA isopropyl function group begins to cleave, yielding high concentrations of the halogenated phenolic compounds C9H10Cl2O2 and C9H10BrClO2. These immediate byproducts in the reaction reach their maximum concentration in approximately 60–90 min. After 120 min, most of the byproducts were further chlorinated. In Stage 3, benzene rings opened in further halogenation. The C–C bonds were broken and the hydrogen on the carbon atoms was substituted with chlorine and bromine, yielding THMs as the main byproducts. The nature of fast-reacting BPA halogenations obtained in the WDS experiment is consistent with the electrophilic substitution mechanisms reported previously (e.g., Hu et al., 2002; Gallard et al., 2003, 2004). Moreover, halogen substitution at the ortho and para positions forms chlorinated and brominated bisphenols as shown in Fig. 7. Vikesland et al. (2013) proposed the reaction follows the second-order kinetics reaction, and the rate of halogenated congeners depends on both BPA and the disinfectant concentration. For this reason, highly chlorinated BPAs dominated in the experiments, whereas brominated BPA congeners were absent in Stage 1 reactions, despite the significantly higher rate constants in the brominated BPA reactions as mentioned by Gallard et al. (2003). Furthermore, the trace bromine concentration in the DWS experiments is the factor that limits the reaction rate. Additionally, in the WSD experiments, the BPA halogenation in Stages 1–3 yields THMs as shown in Fig. 7 and minor HAAs. The formation of THMs in Stage 3 starts after 120 min and stabilises after 24 h. In the pilot-scale of BPA oxidation with purified water,
92
C. Li et al. / Chemosphere 125 (2015) 86–93
Fig. 7. The transformation process of BPA in WDS.
most reports (e.g., Gallard et al., 2004; Deborde et al., 2008; Vikesland et al., 2013) described the BPA oxidation as ending in Stage 1with the formation of highly halogenated BPA such as T4BPA. A few experiments using drinking water or natural waters showed the formation of chlorophenols or phenolic compounds (Hu et al., 2002; Dupuis et al., 2012). In this study, the WDS experimental results show subsequent chlorination transfers of chlorophenols to THMs and minor HAAs from the benzene ring opening and halogen substitutions. Chlorophenols or bromophenols that formed as an immediate reaction in Stage 2 proceeded further in several pathways depending on the types and dosages of disinfectants, the presence of natural organic matters (NOM) as electron scavengers (Peller et al., 2009; Wenk et al., 2011), and, conceivably, the chlorine competition in a water pipe between oxidation in the bulk phase and water demand. The complete degradation path of BPA ? halogenated phenols ? THM (and minor HAAs) is specific for the experimental conditions of this study. As Zhai and Zhang (2011) stated, the immediate reaction product formation is often only one step to control the DBP formation. This is particularly important for BPA because BPA reacts quickly with disinfectants and the resulting immediate and end byproducts, such as chlorophenol and THM/HAA, are known to be highly toxic. 4. Conclusions BPA was investigated for its reactions with and degradation in chlorinated drinking water that was collected from the pilot-scale WDS. BPA and its disinfection derivatives were identified using the IE/MS/GC and GC analysis of water samples. A comparison of the results between the BPA runs in the WDS and blank runs are provided below:
from the congeners to THM were reported in this WDS experiment, which are different from the halogenated BPA congeners as final products in the pilot-scale experiments using a pure water matrix. In addition, the chlorophenol that was immediately formed in Stage 2 was transformed to halogenated THMs. It was also observed that the presence of trace bromide ion in water results in the formation of brominated BPA congeners (MBBPA and MBMCBPA), as well as the THMs without bromine ion (CHCl3 and CHBrCl2) in the final stage of the BPA halogenation. Furthermore, the BPA halogenations follow the second-order kinetic laws as proposed by Vikesland et al. (2013). In this study, these results on quick BPA halogenations and the subsequent transformation to THMs could assist in the characterisation of BPA’s fate and transport in drinking water distribution pipes.
Acknowledgments This work was kindly supported by the National Natural Science Foundation of China (No. 51208455), the Zhejiang Provincial Natural Science Foundation of China (No. LY12E08017), the Doctoral Fund of the Ministry of Education of China (No. 2011010112033), the Zhejiang Provincial Innovative Research Team (2010R50037), and the National Major Program on Pollution Control and Management of Water Body (Grant No. 2012ZX07403-003).
Appendix A. Supplementary material In drinking water distribution systems, BPA reacts quickly with chlorine sequentially producing chlorinated and brominated BPA congeners, halogenated phenols, and finally THM/HAA in three stages. The sequential BPA halogenations
Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2014.11.047.
C. Li et al. / Chemosphere 125 (2015) 86–93
References Azevedod, D.D., Lacorte, S., Viana, P.V., 2001. Occurrence of nonylphenol and bisphenol A in surface water from Portugal. J. Brazil. Chem. Soc. 12, 532–537. Benotti, M.J., Trenholm, R.A., van Derford, B.J., Holady, J.C., Stanford, B.D., Snyder, S.A., 2009. Pharmaceuticals and endocrine distrupting compounds in U.S. drinking water. Environ. Sci. Technol. 43, 597–603. Bolz, U., Hagenmaier, H., Korner, W., 2001. Phenolic xenoestrogens in surface water, sediments, and sewage sludge from Baden-Wiirttemberg, South-west Germany. Environ. Pollut. 115, 291–301. Chen, H.W., Liang, C.H., Wu, Z.M., Chang, E.E., Lin, T.F., Chiang, P.C., Wang, G.S., 2013. Occurrence and assessment of treatment efficiency of nonylphenol, octyphenol and bisphenol-A in drinking water in Taiwan. Sci. Total Environ. 449, 20–48. Deborde, M., von Gunten, U., 2008. Reactions of chlorine with inorganic and organic compounds during water treatment – kinetics and mechanisms: a critical review. Water Res. 42, 13–51. Deborde, M., Rabouan, S., Mazellier, P., Duguet, J.-P., Legube, B., 2008. Oxidation of bisphenol A by ozone in aqueous solution. Water Res. 42, 4299–4308. Dupuis, A., Migeot, V., Cariot, A., Albouy-Llaty, M., Legube, B., Rabouan, S., 2012. Qualification of bisphenol A, 353-nonylphenol and their chlorinated derivatives in drinking water plants. Environ. Sci. Pollut. Res. 19, 4193–4205. Fan, Z.L., Hu, J.Y., An, W., Yang, M., 2013. Detection and occurrence of chlorinated byproducts of bisphenol A, nonylphenol, and estrogens in drinking water of China: comparison to the parent compounds. Environ. Sci. Technol. 47, 10841– 10850. Gallard, H., von Gunten, U., 2002. Chlorination of phenols: kinetics and formation of chloroform. Environ. Sci. Technol. 36, 884–890. Gallard, H., Pellizzari, F., Croué, J.P., Legube, B., 2003. Rate constrants of reactions of bromine with phenols in aquous solution. Water Res. 27, 2883–2892. Gallard, H., Leclercq, A., Croué, J.-P., 2004. Chlorination of bisphenol A: kinerics and by-products formation. Chemosphere 56, 465–473. Gallart-Ayala, H., Moyano, E., Galceran, M.T., 2010. On-line solid phase extraction fast liquid chromatography-tandem mass spectrometry for the analysis of bisphenol A and its chlorinated derivatives in water samples. J. Chromatogr. A1217, 3511–3518. Ge, F., Zhu, L.Z., Chen, H., 2006. Effects of pH on the chlorination process of phenols in drinking water. J. Hazard. Mater. 1–3, 99–105. Guo, G., Lin, F., 2009. The bromination kinetics of phenolic compounds in aqueous solution. J. Hazard. Mater. 170, 645–651. Hendriks, A.J., Maasd, J.L., Noordsij, A., 1994. Monitoring response of XADconcentrated water in the Rhine Delta: a major part of the toxic compounds remains unidentified. Water Res. 28, 581–598. Hu, J.Y., Aizawa, T., Ookubo, S., 2002. Products of aqueous chlorination of bisphenol A and their estrogenic activity. Environ. Sci. Technol. 36, 1980–1987. Hu, J.Y., Aizawa, T., 2003. Quantitative structure-activity relationships for estrogen receptor binding affinity of phenolic chemicals. Water Res. 37, 1213–1222. Jensen, J.S., Lam, Y.F., Heriz, G.R., 1999. Role of Amide Nitrogen in water chlorination: proton NMR evidence. Environ. Sci. Technol. 33, 3568–3573. Kleywegt, S., Pileggi, V., Yang, P., Hao, C., Zhao, X., Rocks, C., Thach, S., Cheung, P., Whitehead, B., 2011. Pharmaceuticals, hormones and bisphenol A in untreated
93
source and finished drinking water in Ontario, Canada – and treatment efficiency. Sci. Total Environ. 409, 1481–1488. Klecˇka, G.M., Staples, C.A., Clark, K.E., van der Hoeven, N., Thomas, D.E., Hentges, S.G., 2009. Exposure analysis of Bisphenol A in surface water systems in North America and Europe. Environ. Sci. Technol. 43, 6145–6150. Li, C., Li, X., Graham, N., 2008. The aqueous chlorination of bisphenol A and steroidestrogens by ferrate. Water Res. 42, 109–120. Li, C., Li, X., Graham, N., 2005. A study of the preparation and reactivity of potassium ferrate. Chemosphere 61, 537–543. Li, X.L., Zou, S.C., Luan, T.G., 2010. Distribution of the three phenolic endocrine disrupting chemicals in the water of the Pearl River Estuary, Guang dong. J. Safety Environ. 10, 96–101 (in Chinese). Liu, Y., Zhao, S.L., Li, Q.B., 2006. Electrochemical oxidation of phenol in chloride-ion system. Environ. Sci. Technol. 11, 21–22. Ma, X.Y., Gao, N.Y., Li, Q.S., 2006. SPE-HPLC determination of trace endocrine disrupting chemicals in raw water. Water Wastewater Eng. 32, 6–10 (in Chinese). Nagel, S.C., vom Saal, F.S., Thayer, K.A., Dhar, M.G., Boechler, M., Welshons, W.V., 1997. Relative binding affinity serum modified access (RBA-SMA) assay predicts the relative in vivo bioactivity of the xenoestrogens bisphenol A and octylphenol. Environ. Health Persp. 105, 70–76. Peller, J.R., Mezyk, S.P., Cooper, W.J., 2009. Bisphenol A reactions with hydroxyl radicals: diverse pathways determined between water and tertiary treated wastewater solutions. Res. Che. Intermed. 35, 21–34. Rule, K.L., Ebbett, V.R., Vikesland, P.J., 2005. Formation of chloroform and chlorinated organics by free-chlorine-mediated oxidation of triclosan. Environ. Sci. Technol. 39, 3176–3185. Shao, X.L., Ma, J., Wen, G., 2008. Investigation of endocrine disrupting chemicals in a drinking water work located in Songhua River Basin. Environ. Sci. 29, 2723– 2728 (in Chinese). U.S. EPA. Bisphenol-a Action Plan. (accessed 24.10.12). Vikesland, P.J., Fiss, E.M., Wigginton, K.R., McNeil, K., Arnold, W.A., 2013. Halogenation of bisphenol-A, triclosan, and phenols in chlorinated water containing iodide. Environ. Sci. Technol. 47, 6764–6772. Wenk, J., von Gunten, U., Canonica, S., 2011. Effect of dissolved organic matter on the transformation of contaminants induced by excited triplet states and the hydroxyl radical. Environ. Sci. Technol. 45, 1334–1340. Zhang, H.F., Hu, J.Y., Chang, H., 2004. Determination of bisphenol A in surface and drinking water in Hangzhou by SPE-LC-MS method. Environ. Chem. 23, 584– 586 (in Chinese). Zhai, H., Zhang, X., 2011. Formation and decomposition of new and unknown polar brominated disinfection byproducts during chlorination. Environ. Sci. Technol. 45, 2194–2201. Zhang, S.L., Liu, D., 2010. Oxidation of phenol by sodium hypochlorite in ammonia nitrogen system. Environ. Sci. Technol. 11, 70–74. Zhao, Y.H., Guo, D.S., 2007. Analysis to the content of di—n—butyl phthalate. Nonylphenol, and bisphenol A in the water of Yellow River. Environ. Monitor. China 23, 19–21 (in Chinese).
Contents lists available at ScienceDirect
Chemosphere journal homepage: www.elsevier.com/locate/chemosphere
Transformation of bisphenol A in water distribution systems: A pilot-scale study Cong Li a, Zilong Wang a, Y. Jeffrey Yang b, Jingqing Liu a, Xinwei Mao a, Yan Zhang a,⇑ a b
College of Civil Engineering and Architecture, Zhejiang University, Hangzhou, China U.S. EPA, National Risk Management Research Laboratory, Cincinnati, USA
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
The halogenations of BPA in a pilot-
Cl CH 3 OH
HO
scale water distribution systems was studied. The chlorinated mechanism of BPA in water distribution systems was proposed. THMs and HAA formation in BPA chlorinated in water distribution systems were found.
CH 3 Br
minor
Cl CH 3 OH
HO CH 3
chlorination in WDS
CH3
Cl
OH
HO
major
Cl
CH3
stage 1
Cl
CH 3 OH
HO CH 3 Cl Cl
Cl CH 3
OH
HO CH 3 Cl
Cl
Cl CH 3 HO
OH CH 3
Cl
Cl
R
CH 3
stage 2
OH
HO Cl
CH 3 R
CH3 HO
Cl
OH CH3 Cl
halogenation
benzene ring opening
CHCl3 CHBrCl2 CHBr2Cl
stage 3
CHBr3 HAAs
a r t i c l e
i n f o
Article history: Received 6 August 2014 Received in revised form 18 November 2014 Accepted 22 November 2014 Available online 27 December 2014 Handling Editor: Shane Snyder Keywords: Bisphenol A Water distribution system Halogenations Mechanism
a b s t r a c t Halogenations of bisphenol A (BPA) in a pilot-scale water distribution system (WDS) of a cement-lined ductile cast iron pipe were investigated. The water in the pilot-scale WDS was chlorinated with a free chlorine concentration of 0.7 mg L1 using sodium hypochlorite, and with an initial BPA concentration of 100 lg L1 was spiked in the WDS. Halogenated compounds in the BPA experiments were identified using EI/GC/MS and GC. Several BPA congeners, including 2-chlorobisphenol A (MCBPA), dichlorobisphenol A (D2-CBPA), 2,20 ,6-trichlorobisphenol A (T3CBPA), 2,20 ,6,60 -tetrachlorobisphenol A (T4CBPA), 2-bromobisphenol A (MBBPA), and bromochlorobisphenol A (MBMCBPA) were found. Moreover, further halogenation yielded other reaction intermediates, including 2,4,6-trichlorophenol (T3CP), dichlorobisphenol A, bromodichlorophenol, and dibromochlorophenol. After halogenation for 120 min, most of the abovementioned reaction intermediates disappeared and were replaced by trihalomethanes (THMs). Based on these experimental findings, the halogenation process of BPA oxidation in a WDS includes three stages: (1) halogenation on the aromatic ring; (2) chlorine or bromine substitution followed by cleavage of the a-C bond on the isopropyl moiety with a positive partial charge and a b0 -C bond on the benzene moiety with a negative partial charge; and (3) THMs and a minor HAA formation from phenolic intermediates through the benzene ring opening with a chlorine and bromine substitution of the hydrogen on the carbon atoms. The oxidation mechanisms of the entire transformation from BPA to THM/HAA in the WDS were proposed. Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction ⇑ Corresponding author. Tel./fax: +86 571 87952015. E-mail address: [email protected] (Y. Zhang). http://dx.doi.org/10.1016/j.chemosphere.2014.11.047 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.
Bisphenol A (BPA) is a well-known endocrine disruption chemical. In despite of relatively low estrogen receptor binding activity
C. Li et al. / Chemosphere 125 (2015) 86–93
in vitro (Hu and Aizawa, 2003), the relatively high estrogenic activity was observed in mouse fetuese (500-fold greater than that of octylphenol) (Nagel et al., 1997). BPA can inhibit the development of the reproductive, embryo and nervous system leading to apparent health concerns. Animal experiments show that BPA can cause leukaemia and increase the lymphoma rate. BPA has attracted more attention recently and is listed in the EPA’s drinking water contaminant candidate list (CCL) (Li et al., 2005; Li et al., 2008; Deborde and von Gunten, 2008; Vikesland et al., 2013; U.S. EPA. Bisphenol-a Action Plan). BPA has been used in high quantities in the manufacture of plastics and beverage cans and may be leaked to the environment in the processes of production, transportation and usage. Therefore, BPA has frequently been detected in surface waters, drinking water, and in the tissues of animals worldwide (Hendriks et al., 1994; Azevedod et al., 2001; Bolz et al., 2001; Benotti et al., 2009; Klecˇka et al., 2009;Kleywegt et al., 2011; Chen et al., 2013). In China, various concentrations of BPA were found in the Shanghai Huangpu River (Ma et al., 2006), the Pearl River basin (Li et al., 2010), the Qiantang River (Zhang et al., 2004), and water from water treatment processes and in water distribution pipes (Zhao and Guo, 2007; Shao et al., 2008; Li et al., 2010). Furthermore, the chlorinated byproducts of BPA in drinking water were firstly reported in the recent paper (Fan et al., 2013). A number of investigations of the chlorinated reaction in beaker experiments were conducted using a purified water matrix. In a beaker experiment, Hu et al. (2002) found that chlorobisphenol A, 2,4,6-trichlorophenol and some polychlorinated phenoxyphenols (PCPPs) were the main byproducts. Jensen et al. (1999) found that amide nitrogens, such as that contained in the acetaminophen structure, were not favourably N-chlorinated. Gallard and von Gunten (2002) found that yields of chloroform formation depends on the concentration of phenol. The halogenation of BPA in an actual WDS is affected by more factors than that in a beaker experiment, including drinking water disinfectants, natural organic matter (NOM), inorganic ions (such as Br and Fe3+), pipe scales, and pipe wall chlorine demand among others. Moreover, there is no report about the by-products difference of BPA halogenations in the presence of Br- in the water distribution systems. Therefore, the halogenations and degradation of BPA in a WDS was studied in this paper. In this study, we report our experimental results on the BPA halogenation for a pilot-scale WDS and analyse the degradation mechanisms of BPA in a WDS. Moreover, in the presence of Br-, the products of BPA halogenations in WDS were firstly reported. 2. Materials and methods 2.1. Materials BPA experiments were performed on a pilot-scale water distribution system (WDS) (see Supplementary Information Fig. 1). This pilot-scale water distribution system has one loop made of cement-lined ductile cast iron pipe used in this study. Each loop is approximately 80 m in length and 150 mm in diameter. Water used in the experiments came from the municipal drinking water network on the Zhejiang University campus in Hangzhou City and was circulated in the loop. Results of water quality measurements, monitored every 2 d (see in Supplementary Information Table 1). A mixture of four THMs standards including trichloromethane (TCM), bromodichloromethane (DCBM), dibromochloromethane (DBCM), tribromomethane (TBM) in methyl tert-butyl ether (MTBE) containing 2000 lg mL1 was purchased from Supelco(Supelco Park, PA, USA). A mixture of six haloacetic acids (HAAs) standards including monochloroacetic acid(MCAA), dichloroacetic acid(DCAA), trichloroacetic acid(TCAA), monobromoacetic acid
87
(MBAA), dibromoacetic acid (DBAA) and bromochloroacetic acid (BCAA) in methyl tert-butyl ether(MTBE) containing 2000 lg mL1 was purchased from Supelco(Supelco Park, PA, USA). Water samples were collected using a device equipped with a 12-position vacuum manifold set(CNW); CNW HC-18 SPE cartridges; pressured gas blowing concentrators (ND200, Shanghai Joyn Electronic CO., Ltd.); EI/GC/MS(GC2010 Plus, MS-QP2010 SE, SHIMADZU); Gas chromatography (GC)(450, VARIAN); and a DR2800 ultraviolet spectrophotometer (HACH). Chemicals and agents used in the experiments included sodium thiosulfate (CP, the Chinese medicine group chemical reagent co., LTD); a free chlorine powder pillow reagent (HACH); and BPA, sodium hypochlorite, acetonitrile, nitric acid, and acetonitrile(AR, Aladdin). 2.2. Experimental methods To identify BPA halogenation in a WDS, two parallel experiments were performed. One set of experiments, denoted as ‘‘BPA runs’’, was based on BPA-spiked water. The other set, denoted as ‘‘blank runs’’, was performed in identical experimental conditions but without the addition of BPA. Therefore, the effects on BPA halogenation and the byproducts of BPA chlorination were identified by comparison. 2.2.1. Halogenation experiment BPA halogenations in chlorinated drinking water were tested in the A loop of the WDS. Before the experiments, the loop was flashed with chlorinated drinking water for 20 min. The water temperature and flow velocity in the WDS were adjusted using automated controls. The experiment conditions were stabilised at pH = 7.3 ± 0.3, a flow velocity of 1.0 m s1, a water temperature of 25 °C ± 1 °C, and an initial free chlorine concentration of 0.7 mg L1 adjusted using a sodium hypochlorite solution. The initial concentration of BPA in this study was 100 lg L1. While water was circulated in the WDS, 1 L samples were taken at 10 min, 30 min, 60 min, 90 min, 120 min, 240 min and 24 h. To these samples were added 2 mL of a 100 mg L1 sodium thiosulfate solution to stop the reaction. 2.2.2. Sample analysis 10 mL water samples were taken from the 1 L samples collected for GC analysis. Before the GC analysis, BPA and its reaction byproducts were extracted by solid-phase extraction (SPE). After the pH adjustment to 3.5 by adding 3 mL HNO3 (0.32 M), the samples passed through the C18 cartridges at a rate of 5 mL min1 using the CNW 12-position vacuum manifold set. Then, the SPE cartridges were dried by passing a gentle nitrogen stream for 2 min. Finally, BPA and its byproducts were eluted using 5 mL acetonitrile, and the effluents were concentrated with a nitrogen stream for approximately 70 min to be analysed by EI/GC/MS. Most byproducts of the BPA halogenation in the WDS were detected by the EI/GC/MS (SHIMADZU, GCMS-QP2010 SE), equipped with a GC splitless injector operating at 260 °C and with a HP-5 MS capillary column (30 m 0.25 mm 0.25 lm, Agilent). Helium was used as a carrier gas with a 1 mL min1 flow rate. The column temperature was raised from 60 °C to 280 °C at a rate of 10 °C min1 and then maintained for 8 min. The mass spectrometer with electron ionisation (70 eV) was operated at 180 °C in an ionisation chamber with a scan range from m/z 50 to 500. THMs and HAAs were detected by GC (Varian, GC-450) with a splitless injector and a SP-Sil 8 CB capillary column (15 m 0.25 mm 0.25 lm, Varian). Headspace injection of THMs was used at 150 °C in the injectors. Column flow was 1 mL min and the temperature of the ECD detector was 290 °C for THMs detection. Headspace injection of HAAs was used at 175 °C in the injectors. The column temperature was raised from 40 °C to 140 °C at a rate of 10 °C min1, from
88
C. Li et al. / Chemosphere 125 (2015) 86–93
140 °C to 190 °C at a rate of 25 °C min1 and then maintained for 8 min at 40 °C and for 0.5 min at 190 °C. Column flow was 1 mL min1 and the temperature of the ECD detector was 300 °C for HAAs detection. 3. Results and discussion 3.1. Formation of BPA congeners Compared to the blank runs (Fig. 1a), the total ion chromatogram (TIC) for the water samples showed that additional nine compounds generated in the BPA halogenations. As shown in Fig. 1b, the new peaks numbered 1–9 were absent in the TIC of the blank samples. The samples’ mass spectra are shown in Fig. 2a–i. Peak 1 in Fig. 1b at the retention time of 19 min is the same as the pure BPA solution from Aladdin. Peak 2 is MCBPA with a retention time of 19.42 min, which is approximate to the BPA peak. The spectra also show a molecular ion (M+) at m/z 262 and a fragment ion at m/z 247. This ion is assigned to C14H12ClO+2 formed by MCBPA losing a methyl. Peaks 3–5 have mass spectra similar to that of BPA and MCBPA. The molecular ions of these peaks were m/z 296, 330, and 366 with retention times of 21.25 min, 21.45 min, and 23.05 min, respectively. These peaks are assigned as DCBPA, T3CBPA and T4CBPA. The mass spectra of Peak 6 show a molecular ion of m/z 307, which is 79 more than that of BPA, with a retention time between MCBPA and DCBPA. Peak 6 is identified as the compound bromobisphenol A (MBBPA). Similar to MCBPA, the fragment ion at m/z 292 is assigned to C14H12BrO+2, which is formed by MBBPA losing a methyl functional group. Peak 7 has a molecular ion at m/z 341, approximately m/z 34 more than that of MBBPA, which is designated as bromochlorobisphenol A (MBMCBPA). The retention time of Peak 1 is 10 min, and its mass spectra are consistent with that of pure 2, 4, 6-trichlorophenol (T3CP). The retention time of Peak 4 is 13.45 min, and its molecular ion is m/z 220. The fragment ion at m/ z 205 is generated by losing CH3, which is consistent with the a-C
on the isopropyl of BPA that can cleave when it reacts with HClO (Hu et al., 2002). In this study, the BPA congeners with more than two bromine atoms were not found due to low concentrations of bromide ion in the drinking water (26 lg L1). In the first 60 min of the halogenation, BPA congeners were the main reaction byproducts (Fig. 3). Over 95% of the BPA was degraded in the first 10 min. The concentration of BPA congeners reached its highest concentration at approximately 10 min and then decreased. Most of the BPA congeners disappeared after 60 min. These experiment results show that in the first 60 min (especially the first 30 min), BPA halogenation in the WDS is the stage of chlorine-substitution and brominesubstitution. 3.2. Formation of halogenated phenols Hu et al. (2002) showed that the a-C bond on the isopropyl of BPA can cleave in the chlorination reactions. The benzene moiety is then substituted with chlorine, and the isopropyl moiety is substituted by a hydroxyl. The chlorine oxidation reaction is consistent with our experimental results. However, in this study, the drinking water in Hangzhou City included chlorine and bromine ions. The mass spectra of three new peaks (Peak 2, Peak 3 and Peak 5) are shown in Fig. 4. Fig. 4 shows that the mass spectra for Peak 1 with a retention time at 10 min are consistent with that of T3CP. For Peaks 2 and 3, their retention time is 11.19 min and 12.38 min, respectively. In Fig. 4a, the molecular ion of Peak 2 is m/z 241. The relative masses of Br and Cl are 80 and 35, respectively. Their difference, 45, is consistent with the Br substitution of the Cl atom in T3CP, upon which Peak 2 is deduced to be bromodichlorophenol. In addition, the molecular ion of Peak 3 with m/z 286 is dibromochlorophenol. Peak 5 is m/z 266, and the fragment ion appears at m/z 251. Its relative molecular weight is 46 more than the molecular weight of D2CBPA (C9H10Cl2O2), but is very similar to the relative
Fig. 1. (a) Total ion GC–MS chromatogram (TIC) of blank sample. Scan range: 0–25 min; (b) the TIC of BPA halogenation in WDS at 10 min. Scan range: 0–25 min.
C. Li et al. / Chemosphere 125 (2015) 86–93
89
Fig. 2. Mass spectra of BPA halogenation byproducts in WDS at 10 min. (a) BPA (peak 1 in Fig. 1); (b) MCBPA (peak 2 in Fig. 1); (c) DCBPA (peak 3 in Fig. 1); (d) T3CBPA (peak 4 in Fig. 1); (e) T4CBPA (peak 5 in Fig. 1); (f) MBBPA (peak 6 in Fig. 1); (g) MBMCBPA (peak 7 in Fig. 1); (h) T3CP (peak 8 in Fig. 1); (i) C9H10Cl2O2 (peak 9 in Fig. 1).
90
C. Li et al. / Chemosphere 125 (2015) 86–93
Fig. 5 shows that from 30 min to 60 min, with further halogenation, the a-C bond on the isopropyl of BPA cleaves to form a large amount of halogenated phenolic compounds. This distinct BPAoxidation stage is labelled as being the halogenated phenols formation. 3.3. Formation of trihalomethanes (THMs) and halogenated acetic acids (HAAs)
Fig. 3. The variation of BPA congeners with time.
mass difference between Br and Cl. As its retention time at 14.5 min is approximate to that of D2CBPA(13.45 min), Peak 5 is interpreted to be MBMCBPA (C9H10BrClO2). With the concentration of BPA congeners decreasing, a large amount of halogenated phenol appeared. The halogenated phenols marked as T3CP reached a maximum concentration at 60 min after reactions in the WDS, but mostly disappeared at 120 min. The changes in their concentration with time are shown in Fig. 5.
BPA chlorination leading to the formation of THMs is obvious by comparing water samples from the BPA chlorination and the water samples with no BPA chlorination. The results are shown in Fig. 6a. In Fig. 6a, the differences in the THMs formation between the BPA chlorination and the blank runs were very small in the first 120 min. The difference becomes significant in the final stage (120–240 min), and the trichloromethane (CHCl3) formation is the largest. The formation rate of trichloromethane minimally increases with the presence of more bromine atoms in the THMs formation. This is apparently related to trace bromine concentration, which is a factor that limits the reaction rate in the drinking water. After 120 min, the rate of the chloroform (CHCl3) and bromodichloromethame (CHBrCl2) formation suddenly increased. Approximately 6.5 lg L1 CHCl3 formed in this stage compared to 14.5 lg L1 that formed at 120 min (Fig. 6a). This increase is consistent with the appearance of THMs in the TIC spectra, indicating that the THMs formation arises from halogenated phenols. The concentrations of the CHCl3, CHBrCl2, CHBr2Cl and CHBr3 compounds were stabilised at 21.84 lg L1, 11.02 lg L1, 6.05 lg L1 and 1.75 lg L1, respectively, after 24 h of chlorination reaction.
Fig. 4. (a) TIC of BPA halogenations in WDS at 60 min. Scan range: 0–25 min; (b) mass spectra of bromodichlorophenol (peak 2 in Fig. 5a); (c) dibromochlorophenol (peak 3 in Fig. 5a); (d) C9H10BrClO2 (peak 5 in Fig. 5a).
C. Li et al. / Chemosphere 125 (2015) 86–93
91
The disappearance of halogenated phenols accompanied by the formation of THMs, particularly CHCl3 and CHBrCl2, strongly indicated a transformation of the phenolic compounds to halogenated methanes through halogenation. In the BPA chlorination reaction, halogenations open the benzene ring with a cleavage of the C–C bonds, followed by a chlorine and bromine substitution for hydrogen on the H–C bonds. Haloacetic acids (HAAs), being the disinfection byproducts, were detected at lower concentrations. Both THM and HAA formations through phenol halogenations (e.g., chlorination) have been reported in published experiments and synthesis (Gallard and von Gunten, 2002; Rule et al., 2005; Ge et al., 2006; Liu et al., 2006; Guo and Lin, 2009; Zhang and Liu, 2010). 3.4. BPA chlorination sequence and implications
Fig. 5. The variation of halogenated phenols with time.
Fig. 6. The concentration curve of THMs with time. The concentration curve of HAAs with time.
Three types of HAAs, including dichloroacetic acid (DCAA), trichloroacetic acid (TCAA), and bromoacetic acid (MBAA), are shown in Fig. 6b. Fig. 6b shows that the HAA formation had a similar regulation as that of the THMs. After 120 min, the rate of HAA formation increased suddenly in the BPA chlorination reaction (Fig. 6b). Moreover, in all of the HAAs, the concentration of DCAA had the greatest increase. After 24 h of reaction time, the final concentrations of DCAA, TCAA and MBAA were stabilised at 6.58 lg L1, 2.29 lg L1 and 3.34 lg L1, respectively.
BPA transformation in pure water with disinfection oxidants has been reported (e.g., Hu et al., 2002; Deborde and von Gunten, 2008; Gallart-Ayala et al., 2010; Vikesland et al., 2013). In the WDS, the reaction pathways and byproducts are different. A BPA oxidation pathway is proposed in Fig. 7 based on the sequential appearance of the byproducts and their mass spectra identifications, as discussed in earlier sections. Under the experimental conditions, the sequential transformation process can be divided into three stages. Stage 1 from 0 min to 60 min is the BPA congener generation. Electrophilic substitution by chlorine to the phenol moiety occurs at ortho and para, namely the 3, 5 or 30 , 50 positions of the phenol moiety (Deborde and von Gunten, 2008; Vikesland et al., 2013). These types of mono-, di-, tri-, and tetra-halogenations are shown in Fig. 7. Hu et al. (2002) detailed the BPA chlorination pathways and mechanisms in beaker tests. In this study, in the WDS, BPA degraded over 95% in the first 10 min. Chlorine-substitution and bromine-substitution in the phenolic moiety on the benzene ring are the primary mechanisms yielding chlorobisphenol A and bromobisphenol A. Their concentrations reached the maximum at approximately 10 min and decreased to almost zero at approximately 60 min. Stage 2 from approximately 30 min to 120 min is the formation of halogenated phenols. It overlaps with Stage 1 in which, as previously mentioned, a small amount of T3CP formed at approximately 30 min. In the phenol formation, the a-C bond on the BPA isopropyl function group begins to cleave, yielding high concentrations of the halogenated phenolic compounds C9H10Cl2O2 and C9H10BrClO2. These immediate byproducts in the reaction reach their maximum concentration in approximately 60–90 min. After 120 min, most of the byproducts were further chlorinated. In Stage 3, benzene rings opened in further halogenation. The C–C bonds were broken and the hydrogen on the carbon atoms was substituted with chlorine and bromine, yielding THMs as the main byproducts. The nature of fast-reacting BPA halogenations obtained in the WDS experiment is consistent with the electrophilic substitution mechanisms reported previously (e.g., Hu et al., 2002; Gallard et al., 2003, 2004). Moreover, halogen substitution at the ortho and para positions forms chlorinated and brominated bisphenols as shown in Fig. 7. Vikesland et al. (2013) proposed the reaction follows the second-order kinetics reaction, and the rate of halogenated congeners depends on both BPA and the disinfectant concentration. For this reason, highly chlorinated BPAs dominated in the experiments, whereas brominated BPA congeners were absent in Stage 1 reactions, despite the significantly higher rate constants in the brominated BPA reactions as mentioned by Gallard et al. (2003). Furthermore, the trace bromine concentration in the DWS experiments is the factor that limits the reaction rate. Additionally, in the WSD experiments, the BPA halogenation in Stages 1–3 yields THMs as shown in Fig. 7 and minor HAAs. The formation of THMs in Stage 3 starts after 120 min and stabilises after 24 h. In the pilot-scale of BPA oxidation with purified water,
92
C. Li et al. / Chemosphere 125 (2015) 86–93
Fig. 7. The transformation process of BPA in WDS.
most reports (e.g., Gallard et al., 2004; Deborde et al., 2008; Vikesland et al., 2013) described the BPA oxidation as ending in Stage 1with the formation of highly halogenated BPA such as T4BPA. A few experiments using drinking water or natural waters showed the formation of chlorophenols or phenolic compounds (Hu et al., 2002; Dupuis et al., 2012). In this study, the WDS experimental results show subsequent chlorination transfers of chlorophenols to THMs and minor HAAs from the benzene ring opening and halogen substitutions. Chlorophenols or bromophenols that formed as an immediate reaction in Stage 2 proceeded further in several pathways depending on the types and dosages of disinfectants, the presence of natural organic matters (NOM) as electron scavengers (Peller et al., 2009; Wenk et al., 2011), and, conceivably, the chlorine competition in a water pipe between oxidation in the bulk phase and water demand. The complete degradation path of BPA ? halogenated phenols ? THM (and minor HAAs) is specific for the experimental conditions of this study. As Zhai and Zhang (2011) stated, the immediate reaction product formation is often only one step to control the DBP formation. This is particularly important for BPA because BPA reacts quickly with disinfectants and the resulting immediate and end byproducts, such as chlorophenol and THM/HAA, are known to be highly toxic. 4. Conclusions BPA was investigated for its reactions with and degradation in chlorinated drinking water that was collected from the pilot-scale WDS. BPA and its disinfection derivatives were identified using the IE/MS/GC and GC analysis of water samples. A comparison of the results between the BPA runs in the WDS and blank runs are provided below:
from the congeners to THM were reported in this WDS experiment, which are different from the halogenated BPA congeners as final products in the pilot-scale experiments using a pure water matrix. In addition, the chlorophenol that was immediately formed in Stage 2 was transformed to halogenated THMs. It was also observed that the presence of trace bromide ion in water results in the formation of brominated BPA congeners (MBBPA and MBMCBPA), as well as the THMs without bromine ion (CHCl3 and CHBrCl2) in the final stage of the BPA halogenation. Furthermore, the BPA halogenations follow the second-order kinetic laws as proposed by Vikesland et al. (2013). In this study, these results on quick BPA halogenations and the subsequent transformation to THMs could assist in the characterisation of BPA’s fate and transport in drinking water distribution pipes.
Acknowledgments This work was kindly supported by the National Natural Science Foundation of China (No. 51208455), the Zhejiang Provincial Natural Science Foundation of China (No. LY12E08017), the Doctoral Fund of the Ministry of Education of China (No. 2011010112033), the Zhejiang Provincial Innovative Research Team (2010R50037), and the National Major Program on Pollution Control and Management of Water Body (Grant No. 2012ZX07403-003).
Appendix A. Supplementary material In drinking water distribution systems, BPA reacts quickly with chlorine sequentially producing chlorinated and brominated BPA congeners, halogenated phenols, and finally THM/HAA in three stages. The sequential BPA halogenations
Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2014.11.047.
C. Li et al. / Chemosphere 125 (2015) 86–93
References Azevedod, D.D., Lacorte, S., Viana, P.V., 2001. Occurrence of nonylphenol and bisphenol A in surface water from Portugal. J. Brazil. Chem. Soc. 12, 532–537. Benotti, M.J., Trenholm, R.A., van Derford, B.J., Holady, J.C., Stanford, B.D., Snyder, S.A., 2009. Pharmaceuticals and endocrine distrupting compounds in U.S. drinking water. Environ. Sci. Technol. 43, 597–603. Bolz, U., Hagenmaier, H., Korner, W., 2001. Phenolic xenoestrogens in surface water, sediments, and sewage sludge from Baden-Wiirttemberg, South-west Germany. Environ. Pollut. 115, 291–301. Chen, H.W., Liang, C.H., Wu, Z.M., Chang, E.E., Lin, T.F., Chiang, P.C., Wang, G.S., 2013. Occurrence and assessment of treatment efficiency of nonylphenol, octyphenol and bisphenol-A in drinking water in Taiwan. Sci. Total Environ. 449, 20–48. Deborde, M., von Gunten, U., 2008. Reactions of chlorine with inorganic and organic compounds during water treatment – kinetics and mechanisms: a critical review. Water Res. 42, 13–51. Deborde, M., Rabouan, S., Mazellier, P., Duguet, J.-P., Legube, B., 2008. Oxidation of bisphenol A by ozone in aqueous solution. Water Res. 42, 4299–4308. Dupuis, A., Migeot, V., Cariot, A., Albouy-Llaty, M., Legube, B., Rabouan, S., 2012. Qualification of bisphenol A, 353-nonylphenol and their chlorinated derivatives in drinking water plants. Environ. Sci. Pollut. Res. 19, 4193–4205. Fan, Z.L., Hu, J.Y., An, W., Yang, M., 2013. Detection and occurrence of chlorinated byproducts of bisphenol A, nonylphenol, and estrogens in drinking water of China: comparison to the parent compounds. Environ. Sci. Technol. 47, 10841– 10850. Gallard, H., von Gunten, U., 2002. Chlorination of phenols: kinetics and formation of chloroform. Environ. Sci. Technol. 36, 884–890. Gallard, H., Pellizzari, F., Croué, J.P., Legube, B., 2003. Rate constrants of reactions of bromine with phenols in aquous solution. Water Res. 27, 2883–2892. Gallard, H., Leclercq, A., Croué, J.-P., 2004. Chlorination of bisphenol A: kinerics and by-products formation. Chemosphere 56, 465–473. Gallart-Ayala, H., Moyano, E., Galceran, M.T., 2010. On-line solid phase extraction fast liquid chromatography-tandem mass spectrometry for the analysis of bisphenol A and its chlorinated derivatives in water samples. J. Chromatogr. A1217, 3511–3518. Ge, F., Zhu, L.Z., Chen, H., 2006. Effects of pH on the chlorination process of phenols in drinking water. J. Hazard. Mater. 1–3, 99–105. Guo, G., Lin, F., 2009. The bromination kinetics of phenolic compounds in aqueous solution. J. Hazard. Mater. 170, 645–651. Hendriks, A.J., Maasd, J.L., Noordsij, A., 1994. Monitoring response of XADconcentrated water in the Rhine Delta: a major part of the toxic compounds remains unidentified. Water Res. 28, 581–598. Hu, J.Y., Aizawa, T., Ookubo, S., 2002. Products of aqueous chlorination of bisphenol A and their estrogenic activity. Environ. Sci. Technol. 36, 1980–1987. Hu, J.Y., Aizawa, T., 2003. Quantitative structure-activity relationships for estrogen receptor binding affinity of phenolic chemicals. Water Res. 37, 1213–1222. Jensen, J.S., Lam, Y.F., Heriz, G.R., 1999. Role of Amide Nitrogen in water chlorination: proton NMR evidence. Environ. Sci. Technol. 33, 3568–3573. Kleywegt, S., Pileggi, V., Yang, P., Hao, C., Zhao, X., Rocks, C., Thach, S., Cheung, P., Whitehead, B., 2011. Pharmaceuticals, hormones and bisphenol A in untreated
93
source and finished drinking water in Ontario, Canada – and treatment efficiency. Sci. Total Environ. 409, 1481–1488. Klecˇka, G.M., Staples, C.A., Clark, K.E., van der Hoeven, N., Thomas, D.E., Hentges, S.G., 2009. Exposure analysis of Bisphenol A in surface water systems in North America and Europe. Environ. Sci. Technol. 43, 6145–6150. Li, C., Li, X., Graham, N., 2008. The aqueous chlorination of bisphenol A and steroidestrogens by ferrate. Water Res. 42, 109–120. Li, C., Li, X., Graham, N., 2005. A study of the preparation and reactivity of potassium ferrate. Chemosphere 61, 537–543. Li, X.L., Zou, S.C., Luan, T.G., 2010. Distribution of the three phenolic endocrine disrupting chemicals in the water of the Pearl River Estuary, Guang dong. J. Safety Environ. 10, 96–101 (in Chinese). Liu, Y., Zhao, S.L., Li, Q.B., 2006. Electrochemical oxidation of phenol in chloride-ion system. Environ. Sci. Technol. 11, 21–22. Ma, X.Y., Gao, N.Y., Li, Q.S., 2006. SPE-HPLC determination of trace endocrine disrupting chemicals in raw water. Water Wastewater Eng. 32, 6–10 (in Chinese). Nagel, S.C., vom Saal, F.S., Thayer, K.A., Dhar, M.G., Boechler, M., Welshons, W.V., 1997. Relative binding affinity serum modified access (RBA-SMA) assay predicts the relative in vivo bioactivity of the xenoestrogens bisphenol A and octylphenol. Environ. Health Persp. 105, 70–76. Peller, J.R., Mezyk, S.P., Cooper, W.J., 2009. Bisphenol A reactions with hydroxyl radicals: diverse pathways determined between water and tertiary treated wastewater solutions. Res. Che. Intermed. 35, 21–34. Rule, K.L., Ebbett, V.R., Vikesland, P.J., 2005. Formation of chloroform and chlorinated organics by free-chlorine-mediated oxidation of triclosan. Environ. Sci. Technol. 39, 3176–3185. Shao, X.L., Ma, J., Wen, G., 2008. Investigation of endocrine disrupting chemicals in a drinking water work located in Songhua River Basin. Environ. Sci. 29, 2723– 2728 (in Chinese). U.S. EPA. Bisphenol-a Action Plan. (accessed 24.10.12). Vikesland, P.J., Fiss, E.M., Wigginton, K.R., McNeil, K., Arnold, W.A., 2013. Halogenation of bisphenol-A, triclosan, and phenols in chlorinated water containing iodide. Environ. Sci. Technol. 47, 6764–6772. Wenk, J., von Gunten, U., Canonica, S., 2011. Effect of dissolved organic matter on the transformation of contaminants induced by excited triplet states and the hydroxyl radical. Environ. Sci. Technol. 45, 1334–1340. Zhang, H.F., Hu, J.Y., Chang, H., 2004. Determination of bisphenol A in surface and drinking water in Hangzhou by SPE-LC-MS method. Environ. Chem. 23, 584– 586 (in Chinese). Zhai, H., Zhang, X., 2011. Formation and decomposition of new and unknown polar brominated disinfection byproducts during chlorination. Environ. Sci. Technol. 45, 2194–2201. Zhang, S.L., Liu, D., 2010. Oxidation of phenol by sodium hypochlorite in ammonia nitrogen system. Environ. Sci. Technol. 11, 70–74. Zhao, Y.H., Guo, D.S., 2007. Analysis to the content of di—n—butyl phthalate. Nonylphenol, and bisphenol A in the water of Yellow River. Environ. Monitor. China 23, 19–21 (in Chinese).
