Decomposition of β-N-methylamino-L -alanine (BMAA) and 2,4-diaminobutyric acid (DAB) during chlorination and consequent disinfection byproducts formation

Water Research 159 (2019) 365e374

Contents lists available at ScienceDirect

Water Research journal homepage: www.elsevier.com/locate/watres

Decomposition of b-N-methylamino-L-alanine (BMAA) and 2,4-diaminobutyric acid (DAB) during chlorination and consequent disinfection byproducts formation Yu Cao a, 1, Shaoyang Hu a, 1, Tingting Gong a, *, Qiming Xian a, Bin Xu b a

State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, 210023, China State Key Laboratory of Pollution Control and Resource Reuse, College of Environmental Science and Engineering, Tongji University, Shanghai, 200092, China b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 12 August 2018 Received in revised form 9 April 2019 Accepted 3 May 2019 Available online 6 May 2019

b-N-methylamino-L-alanine (BMAA) and 2,4-diaminobutyric acid (DAB) are two newly identified algal toxins, and they may react with chlorine to undergo decomposition and generate disinfection byproducts (DBPs) during pre-chlorination as well as chlorine disinfection. In this study, the decomposition of BMAA and DAB during chlorination and the consequent DBPs formation were investigated. The BMAA and DAB concentrations in source waters were determined, the decomposition kinetics of BMAA and DAB and the formation of DBPs during chlorination were studied, the formation pathways of DBPs from BMAA and DAB were explored, and the factors which may affect the decomposition and DBPs formation were examined. The results revealed that BMAA and DAB were commonly detected in source waters from Taihu Lake, and the highest level of BMAA reached 230.8 ng/L, while the concentrations of DAB were generally around 2.0 ng/L. The decomposition of BMAA and DAB during chlorination both followed pseudo-first-order decay while the decomposition rate constant of DAB was significantly higher than that of BMAA. Trihalomethanes (THMs), haloacetic acids (HAAs), and haloacetonitriles (HANs) were all generated during the chlorination of BMAA and DAB with relatively high yields. Notably, the THMs, HAAs, and HANs yields of each carbon atom from BMAA and DAB were significantly higher than that from other organic precursors, and the formation of HANs from DAB was significantly higher than that from BMAA. The formation pathways of DBPs from BMAA and DAB were tentatively proposed and verified through theoretical calculations. Of note, the proposed formation pathways of THMs and HAAs from BMAA were similar to that from DAB, while the proposed formation pathways of HANs from BMAA and DAB showed some differences. Chlorine dose, pH and temperature all affected the decomposition of BMAA and DAB and DBPs formation during chlorination. © 2019 Elsevier Ltd. All rights reserved.

Keywords: Algal toxins BMAA DAB Chlorination Decomposition DBPs formation

1. Introduction Excessive nutrient loading results in eutrophication of surface waters, which may further lead to cyanobacterial (blue-green algae) blooms (Merel et al., 2013). Cyanobacteria are able to release harmful algal toxins, which comprise more than 100 compounds with varying chemical structures and toxicological properties (Lawton et al., 1994; Merel et al., 2013). Due to the negative effects to human health (Briand et al., 2003; Griffiths and Saker, 2003;

* Corresponding author. E-mail address: [email protected] (T. Gong). 1 Both contributed equally to this manuscript. https://doi.org/10.1016/j.watres.2019.05.007 0043-1354/© 2019 Elsevier Ltd. All rights reserved.

Kuiper-Goodman et al., 1999; Pouria et al., 1998), algal toxins have become a serious public concern recently. Algal toxins mainly include hepatotoxins acting on the liver, neurotoxins acting on the nervous system and dermatotoxins causing skin irritation (Merel et al., 2013). Among them, microcystins (MCs, a class of hepatotoxins) are a main family of algal toxins which have been most frequently studied due to their wide occurrence (Fristachi and Sinclair, 2008). A guideline limit of 1 mg/L for MC-LR, one of the most common MCs, has been proposed by the World Health Organization in drinking water (WHO, 1998). Recently, a neurotoxin b-N-methylamino-L-alanine (BMAA) has been newly identified in different countries all over the world (Brand et al., 2010; Esterhuizen and Downing, 2008; Johnson et al., 2008; Li et al., 2010;

366

Y. Cao et al. / Water Research 159 (2019) 365e374

Metcalf et al., 2008). It has been reported that this algal toxin may be produced by all known categories of cyanobacteria (Cox et al., 2005), and thus it may be widely present in surface waters with cyanobacteria. However, as a novel algal toxin, BMAA has not been extensively studied yet. BMAA is a non-protein amino acid with a molecular weight of 118 Da as shown in Fig. 1a. It mainly acts on motor neurons and could also lead to intraneuronal protein misfolding (Banack et al., 2010). BMAA is assumed to be related to different neurodegenerative diseases (Banack et al., 2010; Murch et al., 2004; Pablo et al., 2009). Nevertheless, owing to the lack of toxicological data, no guideline for BMAA has been proposed in drinking water. Besides, 2,4-diaminobutyric acid (DAB) (Fig. 1b), an isomer of BMAA, was also detected in cyanobacteria samples
n and Hellena €s, 2008). DAB is neurotoxic (Krüger et al., 2010; Rose through inhibition of the absorption of the neurotransmitter gaminobutyric acid in neurons and glia cells (Iversen and Kelly, 1975). As two novel algal toxins, BMAA and DAB call for further investigation due to their ubiquity in source waters and diverse adverse effects. Pre-oxidation is widely adopted to oxidize constituents in source waters to facilitate their removal during water treatment, and especially it is widely applied for the pretreatment of algaecontaining source waters (Henderson et al., 2008; Hoko and Makado, 2011; Plummer and Edzwald, 2002). Pre-chlorination is one of the most commonly used pre-oxidation methods. The algal toxins in source waters may undergo decomposition during prechlorination (Acero et al., 2005), and some disinfection byproducts (DBPs) may also be generated from the algal toxins during this process (Chu et al., 2017). Besides, the conventional water treatment system is capable of removing algal cells and the majority of intracellular algal toxins effectively, but it poorly removes extracellular algal toxins, which are released by cyanobacteria to water. Thus the remaining algal toxins in water may also further react with chlorine to form DBPs during chlorine disinfection. Previous studies have reported the formation of various DBPs, including trihalomethanes (THMs), haloaldehydes, and haloacetonitriles (HANs), from MC-LR during chlorination (Chu et al., 2017). However, up to now few studies have reported the decomposition of the novel algal toxins BMAA and DAB during water chlorination and the consequent DBPs formation, which may be a critical issue that should be concerned. Generally the organic DBPs in disinfected waters can be categorized to carbonaceous DBPs (C-DBPs) and nitrogenous DBPs (NDBPs) (Richardson et al., 2007). THMs and haloacetic acids (HAAs) are the two classes of C-DBPs regulated by U.S. EPA, and thus they have received the most public concern (Richardson et al., 2007). Dihalo and trihalo species of HAAs were the most commonly detected in disinfected waters (Richardson et al., 2007). N-DBPs are generally more toxic than C-DBPs (Muellner et al., 2007; Richardson et al., 2007), and thus they have attracted more and more attention. HANs are an important class of N-DBPs, which were widely detected with relatively high concentrations in disinfected waters (Krasner et al., 1989; Richardson et al., 2007; Williams et al., 1997). Dihalo and trihalo species of HANs were the most commonly

detected (Richardson et al., 2007). Since algal toxins contain nitrogen, the formation of N-DBPs from BMAA and DAB should not be ignored. Thus, the formation of the three important classes of CDBPs and N-DBPs, THMs, HAAs, and HANs, from BMAA and DAB during chlorination should be concerned. Thus, the purposes of this study were to determine the concentrations of BMAA and DAB in source waters, to investigate the decomposition kinetics of BMAA and DAB and the formation of THMs, HAAs, and HANs from BMAA and DAB during chlorination, to explore the formation pathways of THMs, HAAs, and HANs from BMAA and DAB during chlorination, and to examine the effects of different factors on the decomposition of BMAA and DAB and the formation of DBPs during chlorination. 2. Materials and methods 2.1. Chemicals and reagents BMAA (
97%), DAB (
95%), H2SO4 (95.0e98.0%), NaOH (
98%), sodium sulfate (
99%), and NH4Cl (
99%) were purchased from Sigma Aldrich. The standard solutions (2000 mg/L) of THMs mixture, HAAs mixture, and HANs mixture in methyl tert-butyl ether (MtBE) were provided by Supelco (USA). The standard solution of simetone (100 mg/L) was obtained from AccuStandard (USA). A sodium hypochlorite stock solution (4e4.99%) was purchased from J&K Scientific, diluted to around 2000 mg/L as Cl2, and standardized using the N,N-diethyl-p-phenylene diamine (DPD) ferrous titrimetric method (APHA et al., 1995) every month. Methanol (HPLC grade), acetonitrile (HPLC grade), and NH3$H2O (25%, HPLC grade) were obtained from Merck. Formic acid (HPLC grade) was purchased from ROE Scientific. MtBE (99.9%) was provided by Tedia. Ultrapure water (18.2 MU$cm) was supplied by a Simplicity UV ultrapure water system (Merck Millipore). 2.2. Collection and pretreatment of source water samples from Taihu Lake The source water samples were collected in October 2015 (during a cyanobacterial bloom) at 10 different locations from Taihu Lake. The distribution of the sampling points is presented in Fig. S1 and Table S1 in the supporting information (SI). The collected source water samples were filtered with 0.45 mm membrane and then stored at 4  C to minimize changes in constituents. Prior to pretreatment, the samples were brought back to room temperature. Oasis MCX cartridges (500 mg, 6 cc) were used for preconcentration of BMAA and DAB in the collected source water samples. Firstly, the solid phase extraction (SPE) cartridges were conditioned prior to use with 5 mL of methanol followed by 5 mL of ultrapure water. Then, 1 L of a sample was loaded onto the cartridges at a flow rate of 20 mL/min. After adsorption, elution with 5 mL of methanol containing 0.1% formic acid followed by 5 mL of methanol containing 5% NH3$H2O was performed. Finally, the eluent was mixed and evaporated to 0.3 mL, diluted with acetonitrile/water (v/v ¼ 1:1, containing 0.1% formic acid) to 1 mL, spiked with simetone as the internal standard, and then filtered with a 0.22 mm membrane before HPLC-MS/MS analysis. Triplicate samples were prepared. 2.3. Preparation of chlorinated samples

Fig. 1. Chemical structures of (a) BMAA, and (b) DAB.

To investigate the decomposition kinetics of BMAA or DAB during chlorination, a set of chlorinated BMAA or DAB samples with different contact times were prepared. Fourteen aliquots (100 mL) of a BMAA or DAB solution (100 mg/L) were prepared. For aliquot 1, no NaOCl was dosed. For aliquots 2e14, NaOCl was dosed at 5 mg/L

Y. Cao et al. / Water Research 159 (2019) 365e374

as Cl2 (in high excess compared with BMAA or DAB), and then the pH was adjusted to 7 with NaOH and H2SO4 solutions. Chlorination was conducted in headspace-free amber glass bottles at 20  C. The contact times of aliquots 2e14 were 2 min, 5 min, 10 min, 20 min, 30 min, 50 min, 1.5 h, 3 h, 6 h, 12 h, 24 h, 36 h, and 48 h, respectively. Notably, to better determine the decomposition kinetics of BMAA or DAB during chlorination, a relatively high concentration of BMAA or DAB (100 mg/L) and a relatively long reaction time range (0e48 h) were adopted in this study. To study the formation of DBPs during chlorination, another set of chlorinated BMAA or DAB samples with different contact times were prepared. Fourteen aliquots (100 mL) of a BMAA or DAB solution (100 mg/L) were prepared. For aliquot 1, no NaOCl was dosed. For aliquots 2e14, NaOCl was dosed at 120 mg/L as Cl2 (Cl2/N molar ratio ¼ 1:1), and then the pH was adjusted to 7 with NaOH and H2SO4 solutions. Chlorination was conducted in headspace-free amber glass bottles at 20  C. The contact times of aliquots 2e14 were 10 min, 20 min, 30 min, 45 min, 1 h, 2 h, 4 h, 6 h, 8 h, 12 h, 24 h, 36 h, and 48 h, respectively. For the two sets of samples, after the contact times, the chlorine residual in each aliquot was measured by the DPD ferrous titrimetric method (APHA et al., 1995) and then dechlorinated with 105% of the requisite stoichiometric amount of NH4Cl. Triplicate samples were prepared. To examine the factors affecting the decomposition of BMAA or DAB during chlorination, three sets of chlorinated BMAA or DAB samples were prepared. To examine the effect of chlorine dose, six aliquots (100 mL) of a BMAA or DAB solution (100 mg/L) were prepared. For the six aliquots, NaOCl was dosed at 15, 30, 60, 120, 240, and 600 mg/L as Cl2 (Cl2/N molar ratio ¼ 0.125, 0.25, 0.5, 1, 2, and 5), respectively, and then the pH was adjusted to 7 with NaOH and H2SO4 solutions. Chlorination was conducted in headspace-free amber glass bottles at 20  C. The contact times were 2 h for the chlorinated BMAA samples and 1 h for the chlorinated DAB samples. After the contact times, the chlorine residual in each aliquot was measured by the DPD ferrous titrimetric method (APHA et al., 1995) and then dechlorinated with 105% of the requisite stoichiometric amount of NH4Cl. Notably, the Cl2/N molar ratios adopted in this study were selected based on previous studies. It has been reported that the average concentration of dissolved organic nitrogen in source waters was around 0.19 mg/L as N (i.e., 13.6 mM) (Lee et al., 2006), and the free chlorine dose in drinking water treatment plants was generally below 5 mg/L as Cl2 (i.e., 70.4 mM), mostly below 3 mg/L as Cl2 (i.e., 42.3 mM) (Connell et al., 2000). Thus, the Cl2/N molar ratios in drinking water treatment plants were generally below 5:1, mostly below 3:1. In this study, we selected the Cl2/N molar ratios of 0.125:1e5:1 for investigation, which were generally in accordance with the Cl2/N molar ratios in drinking water treatment plants. To examine the effect of pH, six aliquots (100 mL) of a BMAA or DAB solution (100 mg/L) were prepared. For all the aliquots, NaOCl was dosed at 120 mg/L as Cl2 (Cl2/N molar ratio ¼ 1:1). The pH of the six aliquots was adjusted to 4, 6, 7, 8, 9, and 10, respectively with NaOH and H2SO4 solutions. Chlorination was conducted in headspace-free amber glass bottles at 20  C. The contact times were 2 h for the chlorinated BMAA samples and 1 h for the chlorinated DAB samples. After the contact times, the chlorine residual in each aliquot was measured by the DPD ferrous titrimetric method (APHA et al., 1995) and then dechlorinated with 105% of the requisite stoichiometric amount of NH4Cl. To examine the effect of temperature, three aliquots (100 mL) of a BMAA or DAB solution (100 mg/L) were prepared. For all the aliquots, NaOCl was dosed at 120 mg/L as Cl2 (Cl2/N molar ratio ¼ 1:1), and then the pH was adjusted to 7 with NaOH and H2SO4 solutions. Chlorination was conducted in headspace-free amber glass bottles at 10, 20, and 30  C for the three aliquots, respectively. The contact times were 2 h for the chlorinated BMAA

367

samples and 1 h for the chlorinated DAB samples. After the contact times, the chlorine residual in each aliquot was measured by the DPD ferrous titrimetric method (APHA et al., 1995) and then dechlorinated with 105% of the requisite stoichiometric amount of NH4Cl. Similarly, to examine the factors affecting the formation of DBPs during chlorination of BMAA or DAB, three sets of chlorinated BMAA or DAB samples were prepared to investigate the effect of chlorine dose, pH, and temperature, respectively. The sample preparation was the same as before except that the contact time for these samples was 24 h. Triplicate samples were prepared. 2.4. Pretreatment of chlorinated samples The samples for BMAA or DAB analysis were pretreated with SPE. Oasis MCX cartridges (60 mg, 3 cc) were used for eliminating the matrices in the chlorinated samples. Firstly, the SPE cartridges were conditioned prior to use with 3 mL of methanol followed by 3 mL of ultrapure water. Then, 1 mL of a sample was loaded onto the cartridges at a flow rate of 10 mL/min. After adsorption, elution with 2 mL of methanol containing 0.1% formic acid followed by 2 mL of methanol containing 5% NH3$H2O was performed. Finally, the eluent was mixed and evaporated to 0.3 mL, diluted with acetonitrile/water (v/v ¼ 1:1, containing 0.1% formic acid) to 1 mL, spiked with simetone as the internal standard, and then filtered with a 0.22 mm membrane before HPLC-MS/MS analysis. The samples for DBPs analysis were pretreated following the U.S. EPA Method 551.1 and 552.3 (U.S. EPA, 1995, 2003). 2.5. Sample analysis BMAA and DAB were determined with an AB Sciex ESI-tqMS (AB SCIEX API4000) coupled by an Agilent HPLC system (Agilent Technologies G1316A-1260 TCC). A ZIC-HILIC column (3.5 mm, 150  2.1 mm, Merck) was used for chromatographic separation. BMAA and DAB were analyzed by multiple reaction monitoring (MRM) mode. THMs, HAAs, and HANs were determined with a gas chromatograph (Agilent 7890A) equipped with an electron capture detector (GC-ECD). A DB-1 capillary column (30 m  0.25 mm, 1 mm film thickness) was used for separation. The instrument parameters are detailed in the SI. 3. Results and discussion 3.1. Determination of BMAA and DAB in source water samples As listed in SI Table S3, for the analysis of BMAA and DAB in this study, the detection limits were 0.2 and 1.4 ng/L, respectively, the recoveries were in the range of 90e110%, and the precisions were in the range of 1.2e8.7%. Besides, BMAA and DAB were not detected in the solvent blank sample as well as the SPE blank sample (SI Figs. S2eS3), indicating that the detection and pretreatment processes did not introduce any contamination. The ten source water samples were collected from Taihu Lake at different locations during a cyanobacterial bloom. Both BMAA and DAB were detected in all the samples and their concentrations are shown in SI Fig. S4. The concentrations of BMAA in the ten samples were in the range of 0.89e230.8 ng/L, while that of DAB were ranged from 1.83 to 2.09 ng/L. The samples from Y3, W1, and W4, which contained relatively high levels of BMAA, were in the zones with the most severe cyanobacterial bloom, which was in accordance with a previous study reporting the spatial distribution of MCs in Taihu Lake (Shi et al., 2015). Notably, BMAA and DAB were widely present in the source water of Taihu Lake, and they may undergo decomposition during pre-chlorination of water treatment and also be organic precursors to form DBPs during pre-chlorination as well as

368

Y. Cao et al. / Water Research 159 (2019) 365e374

chlorine disinfection. 3.2. Decomposition kinetics of BMAA and DAB and formation of DBPs during chlorination 3.2.1. Decomposition kinetics of BMAA and DAB during chlorination The concentrations of BMAA and DAB with contact time during chlorination are shown in Fig. 2a, and the corresponding decomposition percentages are presented in Fig. 2b. The initial concentration of both BMAA and DAB was 100 mg/L. For BMAA, the concentration decreased continuously to 0.7 mg/L within 6 h, and the decomposition percentage reached 99.3% at 6 h. Afterwards, the decomposition percentage generally kept stable until 48 h, indicating that the decomposition of BMAA was complete within 6 h. For DAB, the concentration decreased continuously to 3.3 mg/L within 1.5 h, and the decomposition percentage reached 96.7% at 1.5 h. The decomposition percentage generally kept stable from 1.5 h to 48 h, indicating that the decomposition of DAB was complete within 1.5 h. Based on the concentrations of BMAA and DAB at different contact times, the decomposition kinetics of BMAA and DAB was explored and the results are shown in Fig. 2c and d and Table 1. According to the regression analysis, the decomposition of both BMAA and DAB followed pseudo-first-order decay (higher R2 values). A previous study reported that the decomposition of another algal toxin MC-LR also followed pseudo-first-order decay (Acero et al., 2005). Furthermore, the apparent rate constant of DAB (k ¼ 2.04 h1) was significantly higher than that of BMAA (k ¼ 0.988 h1), indicating that the decomposition of DAB was significantly faster than that of BMAA, which might be due to the differences in their chemical structures. Previous studies have reported that the reaction between amino acids and chlorine started with the electrophilic attack of HOCl to the amino groups (Armesto et al., 1993; Na and Olson, 2007). For BMAA, one amino group is primary and the other one is secondary, which are at a and b sites, while for DAB, both of the amino groups are primary, which are at a and g sites. The amino groups in DAB might be more reactive towards the electrophilic attack of HOCl, resulting in the higher reactivity of DAB compared with that of BMAA.

3.2.2. Formation of DBPs from BMAA and DAB during chlorination The formation of the five DBPs, including trichloromethane (TCM), dichloroacetic acid (DCAA), trichloroacetic acid (TCAA), dichloroacetonitrile (DCAN), and trichloroacetonitrile (TCAN), from BMAA and DAB during chlorination was studied. The linear ranges, detection limits, recoveries, and precisions for the analysis of the DBPs are shown in SI Table S4. The concentrations of TCM in chlorinated BMAA and DAB samples with contact time are presented in Fig. 3a. During the chlorination of BMAA, the concentration of TCM kept increasing from 0 to 27.0 mg/L within 10 h, and then was relatively stable until 24 h. While during the chlorination of DAB, it kept increasing from 0 to 28.0 mg/L within 10 h, and then was relatively stable until 24 h. The formation of TCM from DAB was slightly higher than that from BMAA at different contact times. The concentrations of DCAA and TCAA in chlorinated BMAA and DAB samples with contact time are presented in Fig. 3b. During the chlorination of BMAA, the concentration of DCAA increased continuously from 0 to 26.0 mg/L within 12 h, which kept stable until 24 h, while that of TCAA kept increasing from 0 to 11.0 mg/L until 24 h. During the chlorination of DAB, the concentration of DCAA also increased continuously from 0 to 26.0 mg/L within 12 h, which kept relatively stable until 24 h, while that of TCAA kept increasing from 0 to 11.5 mg/L until 24 h. The formation of DCAA and TCAA from DAB was slightly higher than that from BMAA at different contact times, and the formation of DCAA was significantly higher than that of TCAA from both BMAA and DAB. The concentrations of DCAN and TCAN in chlorinated BMAA and DAB samples with contact time are presented in Fig. 3c. During the chlorination of BMAA, the concentration of DCAN increased from 0 to 35.0 mg/L within 4 h, and then decreased continuously to 23.0 mg/L until 24 h, while that of TCAN increased from 0 to 15.0 mg/ L within 2 h, and then decreased continuously to 3.0 mg/L until 24 h. During the chlorination of DAB, the concentration of DCAN kept increasing from 0 to 40.0 mg/L within 6 h, and then decreased continuously to 33.0 mg/L until 24 h, while that of TCAN kept increasing from 0 to 27.0 mg/L within 4 h, and then decreased continuously to 11.0 mg/L until 24 h. The results indicated that the formation of DCAN and TCAN decreased with relatively long

Fig. 2. (a) Concentrations, and (b) decomposition percentages of BMAA and DAB with contact time during chlorination (The initial concentration of BMAA or DAB was 100 mg/L, the chlorine dose was 5 mg/L as Cl2, the pH was 7, and the temperature was 20  C). Regression of the decomposition kinetics of (c) BMAA, and (d) DAB.

Y. Cao et al. / Water Research 159 (2019) 365e374

369

Table 1 Decomposition kinetic parameters of BMAA and DAB during chlorination. Kinetic model

Parameters

BMAA

DAB

Pseudo-first-order c ¼ c0*exp (ekt)

k (h1) C0 (M) R2 k (M1$h1) C0 (M) R2

0.988 (±0.070) 7.83  107 (±2.2  108) 0.982 2.17  106 (±3.5  105) 8.89  107 (±3.2  108) 0.959

2.04 (±0.27) 8.15  107 (±3.8  108) 0.968 4.18  106 (±1.0  106) 8.40  107 (±7.0  108) 0.919

Pseudo-second-order c ¼ 1/(1/c0 þ kt)

contact times, which might be owing to the decomposition of DCAN and TCAN via hydrolysis and chlorination reactions to form other DBPs, such as haloacetamides, HAAs, and THMs (Glezer et al., 1999; Li et al., 2009), which was in agreement with the proposed formation pathways (Section 3.3, Fig. 4c). The formation of DCAN was significantly higher than that of TCAN from both BMAA and DAB. Previous studies also reported that another algal toxin MC-LR generated significant higher levels of DCAN than TCAN during chlorination (Chu et al., 2017). Different from the C-DBPs (TCM, DCAA, and TCAA), the formation of DCAN and TCAN from DAB was significantly higher than that from BMAA at different contact times, which might be owing to the differences in formation pathways of HANs from BMAA and DAB (Section 3.3). To compare the formation of TCM, DCAA, TCAA, DCAN, and TCAN from BMAA and DAB with that from other organic precursors, the DBP (TCM, DCAA, TCAA, DCAN, or TCAN) yield of each carbon atom from organic precursors was calculated according to the following equation:

DBP yield of each carbon atom ¼

groups (Armesto et al., 1993; Na and Olson, 2007), and thus BMAA might first undergo chlorine substitution on the a-amino group to form the dichloro species. Based on the reactions in previous studies, the formed dichloro species was unstable, and might further undergo decarboxylation, chlorine substitution, hydrolysis reactions and rearrangement to form chloral (Blatchley et al., 2003; Boyce, 1983; Stanbro and Smith, 1979). According to the formation pathways of TCM in previous studies, chloral might decompose to generate TCM (Blatchley et al., 2003; Li et al., 2009). Similarly, BMAA might first undergo a series of chlorine substitution, decarboxylation, hydrolysis reactions and rearrangement to form 2-Nmethylamino-2,2-dichloroacetic acid, which might further decompose to generate DCAA (Blatchley et al., 2003; Boyce, 1983; Stanbro and Smith, 1979). DCAA might continue to react with HOCl to form TCAA. Besides, BMAA might also undergo a series of chlorine substitution, decarboxylation, hydrolysis reactions and rearrangement to form 2-N-methylamino-2,2-dichloroacetonitrile,

Molar concentration of a DBP Initial molar concentration of precursor*number of carbon atoms in precursor

The TCM, DCAA, TCAA, DCAN, and TCAN yields of each carbon atom from different organic precursors (including amino acids, pharmaceuticals, and MCs) are summarized in SI Tables S5eS7. Notably, it was found that the TCM yields of each carbon atom from BMAA (6.69%) and DAB (6.96%) were significantly higher than that from other organic precursors (0.02e0.87%), especially higher than MC-LR (0.87%), which is a typical algal toxin and has been previously reported to show relatively high TCM yield of each carbon atom among various organic precursors (Chu et al., 2012, 2016; 2017; Hureiki et al., 1994). Similarly, the DCAA, TCAA, DCAN, and TCAN yields of each carbon atom from BMAA (0.61e6.16%) and DAB (2.08e8.85%) were also found to be much higher than that from other organic precursors (below 0.67%) (Hong et al., 2009; Jia et al., 2016). Therefore, BMAA and DAB might be two important organic precursors of THMs, HAAs and HANs in concern owing to their high DBPs yields compared with other typical DBPs precursors. 3.3. Formation pathways of DBPs from BMAA and DAB during chlorination 3.3.1. Proposed formation pathways of THMs, HAAs, and HANs from BMAA during chlorination The formation pathways of THMs, HAAs, and HANs from BMAA during chlorination were tentatively proposed based on the reactions in previous studies and are presented in Fig. 4a. Previous studies reported that the reaction between amino acids and chlorine started with the electrophilic attack of HOCl to the amino

which might further decompose to form DCAN (Stanbro and Smith, 1979). DCAN might continue to react with HOCl to form TCAN. In addition, according to previous studies, TCM, DCAA, and TCAA might also be generated from DCAN during chlorination (Glezer et al., 1999; Li et al., 2009). As shown in Fig. 4c, DCAN might undergo a series of chlorine substitution, hydrolysis reactions and rearrangement to form chlorine-substituted acetamide, which might further undergo hydrolysis, chlorine substitution, decarboxylation reactions and rearrangement to generate DCAA, TCAA, and TCM (Chu et al., 2009; Glezer et al., 1999). 3.3.2. Proposed formation pathways of THMs, HAAs, and HANs from DAB during chlorination The proposed formation pathways of THMs, HAAs, and HANs from DAB during chlorination are shown in Fig. 4b. The proposed formation pathways of THMs and HAAs from DAB during chlorination were generally the same as that from BMAA, but the formation pathways of HANs from DAB were partially different from that of BMAA due to the differences of the amino groups in BMAA and DAB. For BMAA, only the a-amino group might be involved in the formation of DCAN and TCAN, while for DAB, both the a- and gamino groups might be involved. Besides the a-amino group, the gamino group in DAB might also contribute to the nitrogen in DCAN and TCAN. As shown in Fig. 4b, DAB might undergo a series of chlorine substitution, decarboxylation, hydrolysis reactions and rearrangement to form DCAN (Blatchley et al., 2003; Boyce, 1983; Stanbro and Smith, 1979). DCAN might further react with HOCl to form TCAN. The proposed formation pathways were in accordance

370

Y. Cao et al. / Water Research 159 (2019) 365e374

with the results demonstrating that the formation of DCAN and TCAN from DAB was significantly higher than that from BMAA (Fig. 3c), while the formation of TCM, DCAA, and TCAA from DAB was just slightly higher than that from BMAA (Fig. 3a and b). 3.3.3. Verification of the proposed formation pathways Notably, the formation pathways in this study were generally proposed based on the reactions in previous studies. To verify the proposed formation pathways of THMs, HAAs, and HANs from BMAA and DAB during chlorination, theoretical calculations were performed using Gaussian 09 software at the B3LYP/3-21G* level. Firstly, the 2FED2HOMO values of the atoms in BMAA and DAB were calculated. Based on the frontier molecular orbital theory, a larger 2FED2HOMO value indicates higher susceptibility to obtain electrophilic attacks. As shown in Table 2, 5N (Fig. 1) in both BMAA (0.65) and DAB (0.78) had the highest 2FED2HOMO values. Thus, the site of 5N should be most easily attacked by electrophiles, which supported the proposed formation pathways. Secondly, the entropy

40

Concentration (μg/L)

35

(a)

BMAA DAB

30 25 20 15 10 5 0 0

Concentration (μg/L)

45

5

10 15 Time (h)

(b)

20

25

BMAA-DCAA DAB-DCAA BMAA-TCAA DAB-TCAA

36 27 18 9 0 0

Concentration (μg/L)

60

5

10 15 Time (h)

(c)

20

25

BMAA-DCAN DAB-DCAN BMAA-TCAN DAB-TCAN

50 40 30 20 10 0 0

5

10 15 Time (h)

20

25

Fig. 3. Concentrations of (a) TCM, (b) DCAA and TCAA, and (c) DCAN and TCAN with contact time during chlorination of BMAA and DAB (The initial concentration of BMAA or DAB was 100 mg/L, the chlorine dose was 120 mg/L as Cl2, the pH was 7, and the temperature was 20  C).

(DS), enthalpy (DH), and Gibbs free energy (DG) of all the reactions in the proposed formation pathways were calculated. Notably, since the reactions a9 and b9 contain more than one single reaction, the DS, DH, and DG of them are not available. As shown in SI Tables S8eS10, DG values of all the reactions at 20  C were below 0 kJ$mol‒1, suggesting that all the reactions in the proposed formation pathways were favorable, and thus the proposed formation pathways should be reasonable. 3.4. Factors affecting the decomposition of BMAA and DAB and formation of DBPs during chlorination 3.4.1. Chlorine dose To examine the effects of different factors on the decomposition of BMAA or DAB during chlorination, the contact time at which the decomposition of BMAA or DAB was moderate (neither too low nor too high) was favored. It was found that with the contact time of 2 h for BMAA or 1 h for DAB, the decomposition percentage of BMAA or DAB was neither too low nor too high. Thus, the contact times of 2 h for BMAA and 1 h for DAB were selected for investigation. The effect of chlorine dose on the decomposition of BMAA and DAB and the formation of DBPs during chlorination is presented in Fig. 5. As shown in Fig. 5a, with the increasing Cl2/N molar ratio from 0.125 to 5, the decomposition percentage of BMAA increased from 60.0% to 94.0%, while that of DAB increased from 35.0% to 90.0%, indicating that increasing the chlorine dose accelerated the decomposition of BMAA and DAB. The results were in agreement with a previous study demonstrating that the decomposition percentages of various amino acids increased with the increasing Cl2/N molar ratio during chlorination (Na and Olson, 2007). In the previous study, the decomposition percentages of various amino acids approached 100% with the Cl2/N molar ratio higher than 1.5 due to the reaction of Cl2 with the amino groups (Na and Olson, 2007). In this study, the decomposition percentages of BMAA and DAB also reached 85.0% and 90.0% respectively with the Cl2/N molar ratio of 2. For the formation of DBPs, as shown in Fig. 5b, during the chlorination of BMAA, the concentration of TCM increased from 9.0 to 26.0 mg/L when the Cl2/N molar ratio increased from 0.125 to 1, but decreased continuously to 15.0 mg/L when the Cl2/N molar ratio further increased until 5. The reason might be that with higher chlorine doses, some other DBPs were more easily generated from BMAA, decreasing the precursors for the formation of TCM. But during the chlorination of DAB, the concentration of TCM kept increasing from 6.0 to 80.0 mg/L when the Cl2/N molar ratio increased from 0.125 to 5. DAB might be more active than BMAA in forming TCM with high doses of chlorine. The formation of DCAA and TCAA from BMAA and DAB with different Cl2/N molar ratios during chlorination is shown in Fig. 5c During the chlorination of BMAA, the concentration of DCAA increased from 10.0 to 34.0 mg/L with the increasing Cl2/N molar ratio from 0.125 to 5, while that of TCAA increased from 5.0 to 15.0 mg/L, suggesting that the formation of both DCAA and TCAA increased with the increasing Cl2/N molar ratio. Similar results were obtained for DAB, with the concentration of DCAA increasing from 11.0 to 40.0 mg/L, and that of TCAA increasing from 4.0 to 16.0 mg/L. The formation of HAAs increased with the increasing chlorine dose during chlorination of both BMAA and DAB, indicating that high chlorine doses favored the formation of HAAs from the two organic precursors during chlorination. Previous studies also demonstrated that increasing the chlorine dose enhanced the formation of HAAs during chlorination of raw waters (Hua and Reckhow, 2008). The formation of DCAN and TCAN from BMAA and DAB with different Cl2/N molar ratios during chlorination is shown in Fig. 5d, which was similar to that of DCAA and TCAA. During the chlorination of BMAA, the concentration of DCAN increased from 9.0 to 30.0 mg/L, and that of TCAN increased from 1.0

Y. Cao et al. / Water Research 159 (2019) 365e374

371

Fig. 4. Proposed formation pathways of TCM, DCAA, TCAA, DCAN, and TCAN from (a) BMAA, and (b) DAB. Proposed formation pathways of TCM, DCAA, and TCAA from (c) DCAN.

Table 2 2FED2HOMO values of the atoms in BMAA and DAB. BMAA Atom 1C 2N 3C 4C 5N 6C 7O 8O

DAB 2FED2HOMO 0.03 0.55 0.03 0.24 0.65 0.03 0.01 0.00

Atom 1N 2C 3C 4C 5N 6C 7O 8O

2FED2HOMO 0.29 0.03 0.10 0.05 0.78 0.13 0.05 0.16

to 10.0 mg/L with the increasing Cl2/N molar ratio from 0.125 to 5, while during the chlorination of DAB, the concentration of DCAN increased from 18.0 to 40.0 mg/L, and that of TCAN increased from 4.0 to 22.0 mg/L. The results indicated that higher chlorine doses favored the formation of HANs from the two organic precursors during chlorination. A previous study also reported that the formation of DCAN increased with the increasing chlorine/MC-LR molar ratio (Chu et al., 2017). Notably, the formation of HANs from DAB was significantly higher than that from BMAA with various chlorine doses, which supported the proposed formation pathways of HANs. As elaborated above, generally higher Cl2/N

molar ratios favored the formation of THMs, HAAs, and HANs from BMAA and DAB during chlorination. Thus, on the basis of disinfection efficiency fulfillment, lower chlorine doses might be recommended in drinking water treatment plants. 3.4.2. pH The effect of pH on the decomposition of BMAA and DAB and the formation of DBPs during chlorination is shown in SI Fig. S5. As shown in Fig. S5a, the decomposition percentage of BMAA decreased from 100% to 73.0% with the increasing pH from 4 to 10, while that of DAB decreased from 95.0% to 70.0%, indicating that acidic conditions favored the decomposition of BMAA and DAB during chlorination. This might be mainly attributed to the speciation of BMAA/DAB and HOCl at different pH values (Na and Olson, 2007). A previous study also reported that the reaction rate constant of MC-LR with chlorine decreased continuously with the increasing pH from 4 to 9 (Acero et al., 2005). For the formation of DBPs, as shown in Fig. S5b, during the chlorination of BMAA, the concentration of TCM increased from 16.0 to 31.0 mg/L when the pH increased from 4 to 9, and kept relatively stable when the pH further increased to 10, while during the chlorination of DAB, the concentration of TCM increased from 21.0 to 35.0 mg/L when the pH increased from 4 to 9, and kept relatively stable when the pH further increased to 10. It has been reported that alkaline

372

Y. Cao et al. / Water Research 159 (2019) 365e374

(a)

BMAA DAB

Concentration (μg/L)

100 80 60 40 20 0 60

0.125

0.25

(c)

Concentration (μg/L)

0.5 1 Cl2/N

2.5

40

(b)

60 40 20

70

0.125

(d)

60

30 20 10

BMAA DAB

80

0

5

BMAA-TCAA BMAA-DCAA DAB-TCAA DAB-DCAA

50

0

100

Concentration (μg/L)

Decomposition percentage (%)

120

50

0.25

0.5 1 Cl2/N

2

5

BMAA-TCAN BMAA-DCAN DAB-TCAN DAB-DCAN

40 30 20 10

0.125

0.25

0.5 1 Cl2/N

2

5

0

0.125

0.25

0.5 1 Cl2/N

2

5

Fig. 5. (a) Decomposition percentages of BMAA and DAB with different Cl2/N molar ratios (The initial concentration of BMAA and DAB was 100 mg/L, the pH was 7, and the temperature was 20  C. The contact time of BMAA was 2 h, and the contact time of DAB was 1 h). Concentrations of (b) TCM, (c) DCAA and TCAA, and (d) DCAN and TCAN with different Cl2/N molar ratios during chlorination of BMAA and DAB (The initial concentration of BMAA and DAB was 100 mg/L, the pH was 7, the temperature was 20  C, and the contact time was 24 h).

conditions favored the formation of THMs (Hua and Reckhow, 2008, 2012), which was owing to the base-catalyzed reactions in THMs formation (Peters et al., 1980; Reckhow et al., 1990). A previous study also demonstrated that the formation of THMs increased with the increasing pH during the chlorination of MC-LR (Chu et al., 2017). The formation of DCAA and TCAA from BMAA and DAB with different pHs during chlorination is shown in Fig. S5c. During the chlorination of BMAA, the concentration of DCAA decreased continuously from 44.0 to 18.0 mg/L with the increasing pH from 4 to 10, while that of TCAA decreased from 21.0 to 4.0 mg/L. Similar results were obtained for DAB, with the concentration of DCAA decreasing from 34.0 to 13.0 mg/L, and that of TCAA decreasing from 20.0 to 4.0 mg/L. The formation of HAAs decreased with the increasing pH during the chlorination of both BMAA and DAB, suggesting that lower pH favored the formation of HAAs, which was in agreement with previous studies (Hua and Reckhow, 2008). The formation of DCAN and TCAN from BMAA and DAB with different pHs during chlorination is shown in Fig. S5d, which was similar to that of DCAA and TCAA. During the chlorination of BMAA, the concentration of DCAN decreased from 33.0 to 16.0 mg/L, and that of TCAN decreased from 9.0 to 4.0 mg/L with the increasing pH, while during the chlorination of DAB, the concentration of DCAN decreased from 45.0 to 16.0 mg/L, and that of TCAN decreased from 20.0 to 4.0 mg/L. The results indicated that lower pH favored the formation of HANs, which might be owing to the enhanced decomposition of HANs with higher pH (Glezer et al., 1999). 3.4.3. Temperature The effect of temperature on the decomposition of BMAA and DAB and the formation of DBPs during chlorination is shown in SI Fig. S6. As shown in Fig. S6a, the decomposition percentage of BMAA increased from 63.0% to 84.0% when the temperature increased from 10 to 20  C, but did not show a significant change when the temperature further increased to 30  C. Similarly, the

decomposition percentage of DAB increased from 61.0% to 85.0% when the temperature increased from 10 to 20  C, but did not show a significant change when the temperature further increased to 30  C. For the formation of DBPs, as shown in Fig. S6b, during the chlorination of BMAA, the concentration of TCM increased from 11.0 to 51.0 mg/L when the temperature increased from 10 to 30  C, while during the chlorination of DAB, the concentration of TCM increased from 12.0 to 32.0 mg/L when the temperature increased from 10 to 30  C, suggesting that increasing the temperature enhanced the formation of TCM. Previous studies also reported that higher temperature favored the formation of THMs (Hua and Reckhow, 2008). The formation of DCAA and TCAA from BMAA and DAB with different temperatures during chlorination is shown in Fig. S6c. During the chlorination of BMAA, the concentration of DCAA increased continuously from 18.0 to 39.0 mg/L with the increasing temperature from 10 to 30  C, while that of TCAA increased from 8.0 to 16.0 mg/L. Similar results were obtained for DAB, with the concentration of DCAA increasing from 16.0 to 33.0 mg/L, and that of TCAA increasing from 9.0 to 15.0 mg/L. The formation of HAAs increased with the increasing temperature during the chlorination of both BMAA and DAB, suggesting that higher temperature favored the formation of HAAs, which was in accordance with previous studies demonstrating that increasing the temperature enhanced the formation of HAAs (Hua and Reckhow, 2008). The formation of DCAN and TCAN from BMAA and DAB with different temperatures during chlorination is shown in Fig. S6d. During the chlorination of BMAA, the concentration of DCAN decreased from 29.0 to 20.0 mg/L, and that of TCAN decreased from 7.0 to 4.0 mg/L with the increasing temperature, while during the chlorination of DAB, the concentration of DCAN decreased from 38.0 to 32.0 mg/L, and that of TCAN decreased from 19.0 to 13.0 mg/L. The results indicated that increasing the temperature decreased the formation of HANs, which might be owing to the enhanced hydrolysis of HANs with higher temperatures (Glezer et al., 1999).

Y. Cao et al. / Water Research 159 (2019) 365e374

Notably, the effect of temperature on the formation of HANs was not so significant as that of pH. 4. Conclusions In this study, the decomposition of BMAA and DAB during chlorination and the consequent DBPs formation were investigated. It was found that BMAA and DAB were commonly detected in source waters from Taihu Lake in China, and the highest level of BMAA reached 230.8 ng/L, while the concentrations of DAB were generally around 2.0 ng/L. The decomposition of BMAA and DAB both followed pseudo-first-order decay while the decomposition rate constant of DAB was significant higher than that of BMAA. THMs, HAAs, and HANs were all generated during the chlorination of BMAA and DAB with relatively high yields. Notably, the THMs, HAAs, and HANs yields of each carbon atom from BMAA and DAB were significantly higher than that from other organic precursors, and the formation of HANs from DAB was significantly higher than that from BMAA. The formation pathways of DBPs from BMAA and DAB were tentatively proposed and verified through theoretical calculations. Of note, the proposed formation pathways of THMs and HAAs from BMAA were similar to that from DAB, while the proposed formation pathways of HANs from BMAA and DAB showed some differences. Chlorine dose, pH and temperature all affected the decomposition of BMAA and DAB and DBPs formation during chlorination. Declaration of interests The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgement This work was supported by the National Natural Science Foundation of China (Grants 51878331, 21876078, 51508264), the Natural Science Foundation of Jiangsu Province, China (Grant BK20150582), the Key Research and Development Program of Jiangsu Province, China (Grant BE2017711), and the Fundamental Research Funds for the Central Universities. Appendix B. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.watres.2019.05.007. References Acero, J.L., Rodriguez, E., Meriluoto, J., 2005. Kinetics of reactions between chlorine and the cyanobacterial toxins microcystins. Water Res. 39, 1628e1638. APHA, AWWA, WEF, 1995. Standard Methods for the Examination of Water and Wastewater, nineteenth ed. Washington, DC. Armesto, X.L., Canle, M.L., Santaballa, J.A., 1993. a-Amino acids chlorination in aqueous media. Tetrahedron 49 (1), 275e284. Banack, S.A., Caller, T.A., Stommel, E.W., 2010. The cyanobacteria derived toxin b-Nmethylaminol-alanine and amyotrophic lateral sclerosis. Toxins 2, 2837e2850. Blatchley III, E.R., Margetas, D., Duggirala, R., 2003. Copper catalysis in chloroform formation during water chlorination. Water Res. 37, 4385e4394. Boyce, S.D., 1983. Reaction pathways of trihalomethane formation from the halogenation of dihydroxyaromatic model compounds for humic acid. Environ. Sci. Technol. 17 (4), 202e211. Brand, L.E., Pablo, J., Compton, A., Hammerschlag, N., Mash, D.C., 2010. Cyanobacterial blooms and the occurrence of the neurotoxin, beta-N-methylamino-Lalanine (BMAA), in south Florida aquatic food webs. Harmful Algae 9, 620e635. Briand, J.F., Jacquet, S., Bernard, C., Humbert, J.F., 2003. Health hazards for terrestrial vertebrates from toxic cyanobacteria in surface water ecosystems. Vet. Res. 34 (4), 361e377. Chu, W., Gao, N., Zhao, S., Li, Q., 2009. Factors affecting formation of THMs during dissolved organic nitrogen acetamide chlorination in drinking water. Environ.

373

Sci. 30 (5), 1376e1380. Chu, W., Gao, N., Krasner, S.W., Templeton, M.R., Yin, D., 2012. Formation of halogenated C-, N-DBPs from chlor(am)ination and UV irradiation of tyrosine in drinking water. Environ. Pollut. 161, 8e14. Chu, W., Chu, T., Bond, T., Du, E., Guo, Y., Gao, N., 2016. Impact of persulfateand ultraviolet light activated persulfate pre-oxidation on the formation of trihalomethanes, haloacetonitriles and halonitromethanes from thechlor(am) ination of three antibiotic chloramphenicols. Water Res. 93, 48e55. Chu, W., Yao, D., Deng, Y., Sui, M., Gao, N., 2017. Production of trihalomethanes, haloacetaldehydes and haloacetonitriles during chlorination of microcystin-LR and impacts of pre-oxidation on their formation. J. Hazard Mater. 327, 153e160. Connell, G.F., Routt, J.C., Macler, B., Andrews, R.C., Chen, J.M., Chowdhury, Z.K., Crozes, G.F., Finch, G.B., Hoehn, R.C., Jacangelo, J.G., Penkal, A., Schaeffer, G.R., Schulz, C.R., Uza, M.P., 2000. Committee report: disinfection at large and medium-size systems. J. Am. Water Work. Assoc. 92, 32e43. Cox, P.A., Banack, S.A., Murch, S.J., Rasmussen, U., Tien, G., Bidigare, R.R., Metcalf, J.S., Morrison, L.F., Codd, G.A., Bergman, B., 2005. Diverse taxa of cyanobacteria produce b-N-methylamino-L-alanine, a neurotoxic amino acid. Proc. Natl. Acad. Sci. U. S. A 102 (14), 5074e5078. Esterhuizen, M., Downing, T.G., 2008. b-N-methylamino-L-alanine (BMAA) in novel South African cyanobacterial isolates. Ecotoxicol. Environ. Saf. 71 (2), 309e313. Fristachi, A., Sinclair, J.L., 2008. Occurrence of cyanobacterial harmful algal blooms workgroup report. In: Hudnell, K.H. (Ed.), Cyanobacterial Harmful Algal Blooms: State of the Science and Research Needs. Springer, New York, USA, pp. 45e103. Glezer, V., Harris, B., Tal, N., Iosefzon, B., Lev, O., 1999. Hydrolysis of haloacetonitriles: linear free energy relationship, kinetics and products. Water Res. 33 (8), 1938e1948. Griffiths, D.J., Saker, M.L., 2003. The Palm Island mystery disease 20 years on: a review of research on the cyanotoxin cylindrospermopsin. Environ. Toxicol. 18 (2), 78e93. Henderson, R., Chips, M., Cornwell, N., Hitchins, P., Holden, B., Hurley, S., Parsons, S.A., Wetherill, A., Jefferson, B., 2008. Experiences of algae in UK waters: a treatment perspective. Water Environ. J. 22 (3), 184e192. Hoko, Z., Makado, P.K., 2011. Optimization of algal removal process at Morton Jaffray water works, Harare, Zimbabwe. Phys. Chem. Earth 36 (14e15), 1141e1150. Hong, C.H., Wong, M.H., Liang, Y., 2009. Amino acids as precursors of trihalomethane and haloacetic acid formation during chlorination. Arch. Environ. Contam. Toxicol. 56, 638e645. Hua, G., Reckhow, D.A., 2008. DBP formation during chlorination and chloramination: effect of reaction time, pH, dosage, and temperature. J. Am. Water Work. Assoc. 100 (8), 82e95. Hua, G., Reckhow, D.A., 2012. Effect of alkaline pH on the stability of halogenated DBPs. J. Am. Water Work. Assoc. 104 (2), 49e50.
, J.P., Legube, B., 1994. Chlorination studies of free and combined Hureiki, L., Croue amino acids. Water Res. 28, 2521e2531. Iversen, L.L., Kelly, J.S., 1975. Uptake and metabolism of g-aminobutyric acid by neurones and glial cells. Biochem. Pharmacol. 24, 933e938. Jia, A., Wu, C., Duan, Y., 2016. Precursors and factors affecting formation of haloacetonitriles and chloropicrin during chlor(am)ination of nitrogenous organic compounds in drinking water. J. Hazard Mater. 208, 411e418. Johnson, H.E., King, S.R., Banack, S.A., Webster, C., Callanaupa, W.J., Cox, P.A., 2008. Cyanobacteria (Nostoc commune) used as a dietary item in the Peruvian highlands produce the neurotoxic amino acid BMAA. J. Ethnopharmacol. 118 (1), 159e165. Krasner, S.W., McGuire, M.J., Jacangelo, J.G., Patania, N.L., Reagan, K.M., Aieta, E.M., 1989. The occurrence of disinfection byproducts in United States drinking water. J. Am. Water Work. Assoc. 81, 41e53. €nch, B., Oppenha €user, S., Luckas, B., 2010. LCeMS/MS determination of Krüger, T., Mo the isomeric neurotoxins BMAA (b-N-methylamino-L-alanine) and DAB (2,4diaminobutyric acid) in cyanobacteria and seeds of Cycas revoluta and Lathyrus latifolius. Toxicon 55, 547e557. Kuiper-Goodman, T., Falconer, I., Fitzgerald, J., 1999. Human health aspects. In: Chorus, I., Bartram, J. (Eds.), Toxic Cyanobacteria in Water: a Guide to Their Public Health Consequences, Monitoring and Management. E & FN Spon, London, UK, pp. 113e153. Lawton, L.A., Edwards, C., Codd, G.A., 1994. Extraction and high-performance liquid chromatographic method for the determination of microcystins in raw and treated waters. Analyst 119, 1525e1530. Lee, W., Westerhoff, P., Esparza-Soto, M., 2006. Occurrence and removal of dissolved organic nitrogen in US water treatment plants. J. Am. Water Work. Assoc. 98, 102e110. Li, A., Chen, Z., Shen, J., Zhai, X., Yang, L., Zhao, S., 2009. THMs and HAAs formation by tryptophan during chlorination disinfection. In: International Conference on Energy and Environment Technology Proceedings, vol. 2, pp. 888e891. Li, A., Tian, Z., Li, J., Yu, R., Banack, S.A., Wang, Z., 2010. Detection of the neurotoxin BMAA within cyanobacteria isolated from freshwater in China. Toxicon 55 (5), 947e953. s, E., Thomas, O., 2013. State of Merel, S., Walker, D., Chicana, R., Snyder, S., Baure knowledge and concerns on cyanobacterial blooms and cyanotoxins. Environ. Int. 59, 303e327. Metcalf, J.S., Banack, S.A., Lindsay, J., Morrison, L.F., Cox, P.A., Codd, G.A., 2008. Cooccurrence of b-N-methylamino-L-alanine, a neurotoxic amino acid with other cyanobacterial toxins in British waterbodies, 1990e2004. Environ. Microbiol. 10 (3), 702e708. Muellner, M.G., Wagner, E.D., Mccalla, K., Richardson, S.D., Woo, Y.T., Plewa, M.J.,

374

Y. Cao et al. / Water Research 159 (2019) 365e374

2007. Haloacetonitriles vs. regulated haloacetic acids: are nitrogen-containing DBPs more toxic? Environ. Sci. Technol. 41, 645e651. Murch, S.J., Cox, P.A., Banack, S.A., Steele, J.C., Sacks, O.W., 2004. Occurrence of bmethylamino-L-alanine (BMAA) in ALS/PDC patients from Guam. Acta Neurol. Scand. 110 (4), 267e269. Na, C., Olson, T.M., 2007. Relative reactivity of amino acids with chlorine in mixtures. Environ. Sci. Technol. 41, 3220e3225. Pablo, J., Banack, S.A., Cox, P.A., Johnson, T.E., Papapetropoulos, S., Bradley, W.G., Buck, A., Mash, D.C., 2009. Cyanobacterial neurotoxin BMAA in ALS and Alzheimer's disease. Acta Neurol. Scand. 120 (4), 216e225. Peters, C.J., Young, R.J., Perry, R., 1980. Factors influencing formation of haloforms in the chlorination of humic materials. Environ. Sci. Technol. 14 (11), 1391e1395. Plummer, J.D., Edzwald, J.K., 2002. Effects of chlorine and ozone on algal cell properties and removal of algae by coagulation. J. Water Supply Res. Technol. Aqua 51 (6), 307e318. Pouria, S., de Andrade, A., Barbosa, J., Cavalcanti, R.L., Barreto, V.T.S., Ward, C.J., Preiser, W., Poon, G.K., Neild, G.H., Codd, G.A., 1998. Fatal microcystin intoxication in haemodialysis unit in Caruaru, Brazil. Lancet 352 (9121), 21e26. Reckhow, D.A., Singer, P.C., Malcolm, R.L., 1990. Chlorination of humic materials: byproduct formation and chemical interpretations. Environ. Sci. Technol. 24 (11), 1655e1664. Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., DeMarini, D.M., 2007. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection by-products in drinking water: a review and roadmap for research.

Mutat. Res. 636 (1e3), 178e242.
n, J., Hellena €s, K., 2008. Determination of the neurotoxin BMAA (b-NmethyRose lamino-L-alanine) in cycad seed and cyanobacteria by LCeMS/MS (liquid chromatography tandem mass spectrometry). Analyst 133, 1785e1789. Shi, K., Zhang, Y., Xu, H., Zhu, G., Qin, B., Huang, C., Liu, X., Zhou, Y., Lv, H., 2015. Long-Term satellite observations of microcystin concentrations in Lake Taihu during cyanobacterial bloom Periods. Environ. Sci. Technol. 49, 6448e6456. Stanbro, W.D., Smith, W.D., 1979. Kinetics and mechanism of the decomposition of N-Chloroalanine in aqueous solution. Environ. Sci. Technol. 13 (4), 446e451. U.S. EPA, 1995. Method 551.1: Determination of Chlorinated Disinfection Byproducts, Chlorinated Solvents, and Halogenated Pesticides/herbicides in Drinking Water by Liquid Liquid Extraction and Gas Chromatography with Electron Capture detection. Revision 1.0. National Exposure Research Laboratory, Office of Research and Development, Cincinnati, OHIO. U.S. EPA, 2003. Method 552.3: Determination of Haloacetic Acids and Dalapon in Drinking Water by Liquid-Liquid Microextraction, Derivatization, and Gas Chromatography with Electron Capture Detection. Revision 1.0. Technical Support Center, Office of Ground Water and Drinking Water, Cincinnati, OHIO. WHO, 1998. In: Guidelines for Drinking-Water Quality. Addendum to Volume 2, Health Criteria and Other Supporting Information, second ed. World Health Organization, Geneva. Williams, D.T., LeBel, G.L., Benoit, F.M., 1997. Disinfection byproducts in Canadian drinking water. Chemosphere 34 (2), 299e316.