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Degradation kinetics and disinfection by-product formation of chlorimuron-ethyl during aqueous chlorination
Accepted Manuscript Degradation kinetics and disinfection by-product formation of chlorimuronethyl during aqueous chlorination Chen-Yan Hu, Yi-Li Lin, Ai-Ping Li, Bin Xu PII: DOI: Reference:
S1383-5866(17)33207-0 https://doi.org/10.1016/j.seppur.2018.04.049 SEPPUR 14545
To appear in:
Separation and Purification Technology
Received Date: Revised Date: Accepted Date:
1 October 2017 14 February 2018 19 April 2018
Please cite this article as: C-Y. Hu, Y-L. Lin, A-P. Li, B. Xu, Degradation kinetics and disinfection by-product formation of chlorimuron-ethyl during aqueous chlorination, Separation and Purification Technology (2018), doi: https://doi.org/10.1016/j.seppur.2018.04.049
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Degradation kinetics and disinfection by-product formation of chlorimuron-ethyl during aqueous chlorination
Chen-Yan Hu1, Yi-Li Lin2*, Ai-Ping Li1, Bin Xu3
1. College of Environmental and Chemical Engineering, Shanghai University of Electric Power, Shanghai 200090; P.R. China. 2. Department of Safety, Health and Environmental Engineering, National Kaohsiung University of Science and Technology, Kaohsiung 824, Taiwan, R.O.C. 3. State Key Laboratory of Pollution Control and Resource Reuse, Institute of Disinfection By-product Control in Water Treatment, College of Environmental Science and Engineering, Tongji University, Shanghai 200092, P. R. China.
* Corresponding author: Yi-Li Lin Email: [email protected] Phone: +886-7-6011000#2328 Fax: +886-7-6011061
1
ABSTRACT: The degradation kinetics of chlorimuron-ethyl during chlorination was
analyzed, and the effects of the influencing factors of pH, bromide and ammonium concentrations and reaction temperature on degradation kinetics were evaluated. Regarding degradation kinetics, the chlorination of chlorimuron-ethyl followed second-order kinetics, first-order kinetics when the chlorine concentration was changed, and first-order kinetics when the chlorimuron-ethyl concentration was changed. The apparent second-order rate constant was calculated, which significantly decreased
increasing
pH.
The
rate constant
of the
reaction
between
chlorimuron-ethyl− and HOCl was calculated as 3.25 (± 1.14) × 103 M-1 min-1, and under the acid-catalyzed condition, the rate constant significantly increased to 6.41 (± 1.62) × 107 M−1 min−1. The degradation rate of chlorimuron-ethyl decreased with increasing bromide (at pH 5.5–7.0) and ammonium concentrations (≤ 6 μM). Chlorimuron-ethyl chlorination is an endothermic reaction, with an activation energy estimated as 22.46 kJ mol−1 using the Arrhenius equation. During chlorimuron-ethyl chlorination, chloroform was the major volatile degradation product during chlorimuron-ethyl chlorination, and its concentration increased with the increasing reaction time and pH level. The distribution of disinfection by-products (DBPs) formed at different solution pH levels was quite distinct. In the presence of bromide, the concentration and species of brominated DBPs increased with the increasing bromide concentration. Notably, carcinogenicity and mutagenicity of brominated DBPs were higher than their chlorinate analogs.
Keywords: Chlorimuron-ethyl; Chlorination; Kinetics; Disinfection by-products
(DBPs); Herbicide; Water treatment
2
1. Introduction To increase agricultural production, many different types of herbicides have been extensively applied worldwide to destroy pests and weeds in farmland [1]. Sulfonylurea herbicides are the most widely used herbicides [2]. With the modification of functional groups on the aromatic and heterocyclic rings of sulfonylurea, more than 30 types of sulfonylurea herbicides have been developed, which are widely used globally. In 1988, the US Geological Survey analyzed 214 samples of surface and underground water collected from Midwestern United States. In this survey, nicosulfuron and chlorimuron-methyl were detected in 53% and 31% of the total samples, respectively, and chlorosulfuron, chloropyrazosulfuron and other sulfonylurea herbicides were detected in 5% of the samples [3]. Struger et al. analyzed samoles of surface waters collected from central Canada for 3 consecutive years and reported high concentrations of sulfonylurea herbicides in the samples due to wash-off by rain [4]. The presence of sulfonylurea herbicides in the environment poses potential risks to human health due to their mutagenicity, toxicity and carcinogenicity
[5,
6].
Chlorimuron-ethyl
2-(((((4-chloro-6-methoxypyrimidin-2-yl)
amino)
(the
CAS
carbonyl)
name:
amino)
ethyl
sulfonyl)
benzoate) is a nitrogen-containing herbicide in the family of sulfonylurea herbicides and it is widely used in modern agriculture to control weeds in wheat, rice, and corn due to its high efficiency and the low dosage required for application (2–75 g ha–1) [7, 8]. A concentration of chlorimuron-ethyl as high as 0.3 μg L–1 has been detected in surface water, and a median concentration of < 0.01 μg L–1 has been detected in surface and ground water [9]. In Heilongjiang Province in China, 400 ton of chlorimuron-ethyl is used annually in soybean fields for weed control [10]. Chlorimuron-ethyl is a long-term residual herbicide with a low Kow value of 2.5 [11] 3
and a high water solubility of 1.2 g L–1 at pH 7 [12]; therefore, it can easily be leached from the soil into the surface or groundwater [13, 14] that can be used as a source of drinking water, thus posing a threat to human health. Therefore, identifying the ultimate fate of chlorimuron-ethyl in drinking water treatment process is necessary. Chlorination is an efficient and convenient method of killing waterborne pathogenic microorganisms to guarantee the safety of drinking water. Thus, chlorine is used as a disinfection agent in water treatment plants worldwide [15, 16]. However, chlorine is a strong oxidant and can react with organic matter to form harmful disinfection by-products (DBPs), including trihalomethanes (THMs) and haloacetic acids (HAAs), which are reported to be carcinogenic and mutagenic to animals and humans [17-19]. Several studies have focused on the chlorination of herbicides such as diuron [20], chlortoluron [21] and dinoseb [22], but not chlorimuron-ethyl. Few studies have reported the degradation of chlorimuron-ethyl by microorganisms [23] and through Fenton, photo-Fenton, or ozonation processes [24]. It is essential to determine the degradation kinetics and DBP formation of chlorimuron-ethyl during chlorination, and it is necessary to evaluate the effects of water parameters such as pH, bromide and ammonium concentrations, and temperature on degradation kinetics, which have been demonstrated to have a crucial influence on the degradation of herbicides during chlorination [21, 25]. The results can provide valuable information for the control of chlorimuron-ethyl and corresponding DBP formation in the water treatment process. . The objectives of this study were (1) to investigate the degradation kinetics of chlorimuron-ethyl during chlorination, (2) to study the effect of the influencing factors of pH, bromide and ammonium concentrations, and temperature on chlorimuron-ethyl degradation, and (3) to assess the yield of volatile DBPs during 4
chlorimuron-ethyl chlorination as a function of the reaction time and the pH of the solution.
2. Materials and Methods 2.1. Chemicals and reagents Chlorimuron-ethyl (>99%) and DBP standards (EPA 551A and 551B) including chloroform (CF), bromodichloromethane (BDCM), dibromochloromethand (DBCM), bromoform (TBM), trichloroacetonitrile (TCAN), dichloroacetonitrile (DCAN), bromochloroacetonitrile trichloronitromethane
(BCAN), (TCNM),
dibromoacetonitrile
1,1-dichloro-2-propanone
(DBAN),
(1,1-DCP)
and
1,1,1-trichloroacetone (1,1,1-TCP) were obtained from Sigma-Aldrich (St. Louis, MO, USA) and used as received. Analytical grade of NaOCl (available chlorine 4.00 4.99%), Na2S2O3, NaBr, NH4Cl, phosphate, carbonate, acetate, NaOH, and HCl were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). High performance liquid chromatography (HPLC) grade methanol and methyl tert-butyl ether (MTBE) were purchased from J.T. Baker (USA). All solutions were prepared with Milli-Q water.
2.2. Experimental procedures All experiments were performed in a 200-mL batch reactor equipped with a dispenser and were controlled at 25 ± 1°C in a water bath. The experiments were initiated by adding an aliquot of concentrated chlorine solution to the reactor containing 7.23 μM of chlorimuron-ethyl solution (3 mg L−1) buffered using 10 mM acetate for a pH of 4–5, 10 mM phosphate for a pH of 5–8, and 10 mM carbonate for a pH of 8–10. Small volumes of H2SO4 and NaOH were used to adjust the pH of the 5
solution to the desired initial values. At different reaction times, 1 mL of the solution was rapidly transferred into an HPLC vial containing 50 μLof NH4Cl solution (1 M) to quench the reaction, and the solution was then analyzed using HPLC as soon as possible. The kinetic experiments were conducted over a pH range of 4–10 at the molar ratio of chlorine to chlorimuron-ethyl of 7.4–25.0:1. The chlorimuron-ethyl concentration was monitored until more than 50% of chlorimuron-ethyl was degraded. The effects of bromide and ammonium concentrations (0–3.62 μM and 0–6 μM, respectively) and the reaction temperature (20°C, 25°C, 30°C, 35°C and 40°C) on chlorimuron-ethyl chlorination were evaluated at a chlorimuron-ethyl concentration of 7.23 μM, pH 7 (maintained using phosphate buffer) and a chlorine concentration of 108.6 μM (as Cl2). For conducting kinetic and DBP formation studies, the chlorine concentration should be at least 10 times the concentration of the target substance [20, 21, 26]. Accordingly, in this study, the chlorine concentration was set at 15 times the chlorimuron-ethyl concentration to analyze the effect of bromide and ammonium concentrations, and temperature on chlorimuron-ethyl chlorination; a similar procedure was followed for the DBP formation experiments. Additionally, a blank experiment was conducted. During the experimental period, no chlorimuron-ethyl degradation was observed without the addition of chlorine (data not shown). The
experiments
for
the
determination
of
DBP
formation
during
chlorimuron-ethyl chlorination were conducted in duplicate under headspace-free conditions in 45-mL glass of screw-cap vials with polytetrafluoroethylene-lined septa. A typical experimental condition was 7.23 μM chlorimuron-ethyl solutions dosed with 108.6 μM chlorine dosage at a pH range of 5–10 using 10 mM acetate, phosphate, or carbonate buffers. At designated reaction times, the reaction was 6
quenched using NH4Cl, and water samples then were extracted using MTBE to analyze DBP formation.
2.3 Analytical methods The chlorimuron-ethyl concentration was analyzed using an Agilent 1260 infinity HPLC system equipped with an autosampler and a Waters X Terra C18 column (250 mm × 4.6 mm, 5 μm, Waters, USA), and the concentration was then detected using a UV detector at a wavelength of 254 nm. The mobile phase consisted of 80%/20% (v/v) methanol and Milli-Q water (the pH of 4.5 was maintained by the addition of phosphoric acid). The flow rate was 1 mL min–1, and the injection volume was
5
μL.
The
chlorine
concentration
was
analyzed
using
the
N, N-diethyl-p-phenylenediamine (DPD) colorimetric method [27]. The pH of the solution pH was measured using a regularly calibrated pH meter (Sartorius, Germany). DBP formation was quantified following the method 551.1 of the US Environmental Protection Agency (US EPA) [28] on a gas-chromatograph (GC-2010 Plus, Shimadzu, Japan) equipped with an electron capture detector and a RTX-5 fused silica capillary column (30 m × 0.25 mm id, 0.25-μm film thickness). Helium was used as the carrier gas at a flow-rate of 0.94 mL min–1. The temperature program was as follows: injector temperature of 200°C, detector temperature of 290°C, a column temperature of 37°C (hold for 10 min) that was increased to 50°C (with a rate of 5°C per min), then increased to 260°C (with a rate of 30°C per min) and hold for 10 mins. The detection limits of CF, BDCM, DBCM, TBM, TCAN, DCAN, BCAN, DBAN, TCNM, 1,1-DCP and 1,1,1-TCP were 2, 1, 1, 2, 0.5, 1, 1, 1, 5, 1 and 0.5 μg L-1, respectively. 7
3. Results and Discussion 3.1 Reaction kinetics The chlorination of chlorimuron-ethyl was performed at a chlorimuron-ethyl concentration of 7.23 μM, chlorine concentration of 53.8–180.9 μM, pH 7 and reaction temperature of 25°C. The results are provided in Fig. 1, in which [Chlorimuron-ethyl]T,0
is
the
initial
chlorimuron-ethyl
concentration
and
[Chlorimuron-ethyl]T,t is the chlorimuron-ethyl concentration at different reaction times. A good linear fit (R2 > 0.97) in all experimental data verified the pseudo-first-order reaction kinetics of chlorimuron-ethyl degradation, and the slopes revealed the pseudo-first-order rate constants (kobs). The upper-right insert figure in Fig.1 reveals that kobs is linearly related to the initial total chlorine concentration ([HOCl]T, including HOCl and OCl−) (R2 = 0.999), which indicates that the chlorination of chlorimuron-ethyl is a first-order reaction when the chlorine concentration is increased. Accordingly, the chlorination of chlorimuron-ethyl follows the second-order kinetics, first-order kinetics when the chlorine concentration is changed, and first-order kinetics when the chlorimuron-ethyl concentration is changed; these findings have also been reported for other herbicides [21, 26, 29]. Additional experiments at different initial concentrations of chlorimuron-ethyl were conducted, and the results are provided in Fig. S1 in the Supplementary Information. With the addition of a high concentration of chlorine, the degradation of chlorimuron-ethyl at different initial concentrations followed first-order kinetics. Therefore, the kinetics of chlorimuron-ethyl chlorination are described in Eq. (1), where kapp is the apparent second-order rate constant at specific pH levels, which can be calculated from the slope of the plot of kobs versus [HOCl]T. 8
d[Chlorimu ron - ethyl]T = kapp [HOCl]T [Chlorimuron-ethyl]T dt = kobs [Chlorimuron-ethyl]T
(1)
3.2 Effect of pH on the reaction rates of elementary reactions The reaction rates of reactions between chlorine and organic compounds are highly affected by the pH of the solution [30]. Therefore, the effect of pH on chlorimuron-ethyl chlorination was investigated at initial chlorimuron-ethyl and chlorine concentrations of 7.23 and 108.6 μM, respectively, and the results are depicted in Fig. 2. The pKa of chlorimuron-ethyl is 4.20 [12],and that of NaClO is 7.5. At a pH of > 4.2, most chlorimuron-ethyl molecules will dissociate to chlorimuron-ethyl− so that the total amount of chlorimuron-ethyl species will be represented by the ionized one ([Chlorimuron-ethyl]T≈ [Chlorimuron-ethyl−]). At a pH of ≤ 7.5, the major chlorine species in solution will be HOCl. Therefore, the reactions due to chlorimuron-ethyl degradation during chlorination are as follows: Ka1 = 2.88×10−8
HOCl ↔OCl-+H+ [HOCl]T = [HOCl] + [OCl-]
(2) (3)
Chlorimuron-ethyl ↔ Chlorimuron-ethyl- + H+
Ka2 = 6.31×10−5
(4)
[Chlorimuron-ethyl]T = [Chlorimuron-ethyl] + [Chlorimuron-ethyl−] ≈[Chlorimuron-ethyl−]
(5)
Chlorimuron-ethyl− + HOCl+ H+→ products
k1
(6)
Chlorimuron-ethyl− + HOCl→ products
k2
(7)
Because OCl− has weaker oxidation power than HOCl, and the calculated reaction rate constant of reactions between OCl− and chlorimuron-ethyl was very close to zero (data not shown), the OCl− oxidizing chlorimuron-ethyl− reaction was negligible. Therefore, the degradation of chlorimuron-ethyl can be expressed as: 9
d[Chlorimu ron - ethyl]T = kapp [HOCl]T [Chlorimuron-ethyl]T dt = k1 [Chlorimuron-ethyl−] [HOCl] [H+] + k2 [Chlorimuron-ethyl−] [HOCl]
(8)
After replacing [HOCl] with [HOCl]T/(1 + α) and [Chlorimuron-ethyl−] with [Chlorimuron-ethyl]T, where α is the ionization percentage of chlorine, the degradation of chlorimuron-ethyl can be further expressed as: 2 d[Chlorimu ron - ethyl]T = k1 ([H] ) k 2 [H] [HOCl]T [Chlorimuron-ethyl]T (9) dt Ka1 [H]
Comparing Eqs. (1) and (9), the apparent rate constant kapp can be expressed as: kapp =
k1 ([H] ) 2 k 2 [H] Ka1 [H]
(10)
The rate constants k1 and k2 were calculated as 6.41 (±1.62) × 107 M−1 min−1 and 3.25 (±1.14) × 103 M−1 min−1, respectively, which were compared with the rate constants of several herbicides listed in Table 1. The acid-catalyzed reaction rate constant k1 of chlorimuron-ethyl was 104 times higher than k2, which was similar to the rate constants of chlortoluron, dinoseb and metribuzin chlorination (Table 1). The k2 value of chlorimuron-ethyl was much higher than those of chlortoluron, dinoseb, and bensulfuron-methyl listed in Table 1, which indicates that chlorimuron-ethyl can react with chlorine and consume HOCl much faster than the aforementioned herbicides during chlorination.
3.3 Effect of bromide concentration on chlorimuron-ethyl chlorination The bromide concentrations detected in the Yellow River and Danjiangkou Reservoir in China were 0.75 and 0.48 μM, respectively [31]. During the chlorination of organic compounds, bromide can react with HOCl to form HOBr, leading to the formation of more toxic brominated DBPs than the chlorinated DBPs [17-19]. 10
Therefore, the effect of the bromide concentration on chlorimuron-ethyl chlorination at pH levels of 5.5–8.5 was investigated, and the results are provided in Fig. 3. Chlorimuron-ethyl chlorination at different bromide concentrations over the pH range of 5.5–8.5 followed pseudo-first-order kinetics well. By contrast, bromide had the opposite effect on chlorimuron-ethyl degradation over the pH range of 5.5–7.0 (Fig. 3(a)&(b)) and the pH level of 8.5 (Fig. 3(c)). Bromide reacts with HOCl quickly in an aqueous solution and forms the stronger oxidizing agent, HOBr (Eq. (11)), under basic conditions, which increases the degradation of organic matter [26, 32]. This phenomenon explains the increasing rate of chlorimuron-ethyl degradation when the bromide concentration increased from 0 to 6.0 μM at pH 8.5. However, at a pH of 5.5–7.0, the presence of bromide slightly inhibited chlorimuron-ethyl degradation under the conditions of the increasing bromide concentration and decreasing pH, which may have been due to a series of complex reactions that transferred HOCl and HOBr to less oxidative agents [Eqs. (2) and (11)-(13)] [33, 34]. Further research on the role of bromide in chlorimuron-ethyl chlorination is required. HOCl + Br− → HOBr + Cl−
k3 = 1.55×103 M−1 s−1
(11)
HOBr ↔ OBr− + H+
Ka3 = 1.26 × 10−9
(12)
HOCl + HOBr → BrO3− + Cl- + 3H+
k4
(13)
3.4 Effect of ammonium concentration on chlorimuron-ethyl chlorination Ammonium is widely present in water and can react with chlorine to form chloramines, including monochloramine (NH2Cl), dichloramine (NHCl2) and trichloramine (NCl3) [35]. The formation of trichloramine is negligible at a pH level of > 4.5 [36]. Thus, reactions of chloramine formation only and the corresponding rate constants are expressed in Eqs (14)–(16) [16]. 11
NH4+ ↔ NH3 + H+
Ka4 = 5×10−10
(14)
HOCl + NH3 → NH2Cl + H2O
k6 = 4.2×106 M−1s−1
(15)
HOCl + NH2Cl → NHCl2 + H2O
k7 = 1.1×103 M−1s−1
(16)
The effect of ammonium concentration on the degradation of chlorimuron-ethyl during chlorination was investigated at a pH of 7, and the results are presented in Fig. 4.
The
experimental
data
indicated
that
the
degradation
rate
of
chlorimuron-ethyl during chlorination decreased with the increasing ammonium concentration due to the competitive consumption of HOCl with ammonium (k 2 in Eq. (7) << k6 in Eq. (15). Because monochloramine and dichloramine have weaker oxidation power than HOCl [36], the degradation of chlorimuron-ethyl by chloramines is negligible.
3.5 Effect of temperature on chlorimuron-ethyl chlorination Temperature plays a crucial role in the rate constant, product yield, and species distribution in chemical reactions. To determine the effect of temperature on chlorimuron-ethyl degradation during chlorination, a set of experiments were conducted at temperatures between 20°C and 40°C. The correlation of ln(kapp) and the reciprocal of temperature (T–1) are provided in Fig. 5. As illustrated in Fig. 5, an increasing temperature accelerated the degradation rate of chlorimuron-ethyl, with a good linear fit between ln(kapp) and T–1 (R2 = 0.99). Using the Arrhenius equation, the activation energy of chlorimuron-ethyl chlorination was estimated to be 22.46 kJ mol−1, which was higher than that of bensulfuron-methly chlorination [26]. The results validate that a more efficient reaction occurs between chlorimuron-ethyl and chlorine, as discussed in Section 3.2.
12
3.6 DBP formation during chlorimuron-ethyl chlorination The formation of carbonaceous and nitrogenous DBPs during chlorimuron-ethly chlorination was evaluated, and the results are presented in Fig. 6. Four DBPs were detected: CF, DCAN, TCNM and 1,1,1-TCP. CF was the major volatile degradation product, and the CF concentration increased with the reaction time, reaching a maximum concentration of 84.2 μg L−1 on day 7 due to excess chlorine in the solution. However, the concentrations of DCAN, TCNM, and 1,1,1-TCP were one-order lower than that of CF. TCNM formation increased quickly on day 1 and remained relatively stable until day 7. A similar phenomenon was also observed during the chlorination of metribuzin and bensulfuron-methyl [26, 32], which implied that TCNM is a stable DBP end-product that does not undergo further hydrolysis or oxidation. As for DCAN and 1,1,1-TCP, The maximum concentrations of DCAN and 1,1,1-TCP (4.8 and 5.3 μg L−1, respectively) were reached on day 1, and the concentrations decreased gradually as the reaction time passed, which can be explained by the hydrolysis and oxidation of the DBPs formed by chlorine [37]. According to Yang et al. [38], TCP can be hydrolyzed to CF, and DCAN tends to hydrolyze at a pH of > 7 to produce dichloroacetic acid; this hydrolysis reaction is accelerated in the presence of free chlorine [39]. This may also explain why DCP and TCAN were not detected in this study. Different pH levels exert considerable influence on the dissociation of HOCl, the degree of halogenation, and the hydrolysis reactions during organic compound chlorination, which results in the formation of DBPs. Therefore, the effect of the pH of the solution on DBP formation during chlorimuron-ethyl chlorination for 7 days was investigated, and the results are provided in Fig. 7. CF was the dominant DBP, and its concentration increased significantly from 31.9 to 137.4 μg L−1 as the pH level 13
increased from 5 to 9. This phenomenon can be explained by the OH− requirement in the haloform formation process as well as the hydrolysis of haloacetonitriles (HANs) or haloketones under basic conditions [40-42]. The concentrations of DCAN and 1,1,1-TCP decreased continuously as the pH level increased from 5 to 9, which can be explained by base-catalyzed hydrolysis decomposition, and the hydrolysis rate can be accelerated by increasing the pH level [38]. TCNM formation could be due to the cleavage of the aromatic ring and side chains containing carbonyl and dimethylamine groups during chlorimuron-ethyl chlorination, which changed slightly as the pH increased from 5 to 9 [37]. The negligible effect of pH on TCNM formation has also been reported by other researchers during chlorination of other nitrogenous precursors [43, 44]. Bromide participation in chlorimuron-ethyl chlorination may result in the formation of more toxic brominated DBPs than chlorinated analogs. Therefore, the effect of the bromide concentration on DBP formation during chlorimuron-ethyl chlorination at a pH of 7 for 7 days was investigated, and the results are provided in Fig. 8. The total concentrations of THMs and HANs increased as the molar ratio of bromide to chlorine increased from 0 to 0.05. The THM concentration reached 90.81 μg L-1, which exceeded the standard concentration of 80 μg L-1 set in the National Primary Drinking Water Regulations of the US EPEA [45] at the molar ratio of bromide to chlorine of 0.01. The presence of bromide during chlorimuron-ethyl chlorination resulted in the formation of five more brominated DBPs: BDCM, DBCM, TBM, BCAN and DBAN, which may be due to the reaction between HOBr and the benzene and heterocyclic rings of chlorimuron-ethyl. The reaction rate between HOBr and organic matter were dozens of times higher than that between HOCl and organic matter [46]. Thus, with the increase of bromide concentration, more HOBr will be 14
produced and subsequently reacteded with chlorimuron-ethly. This reaction will consume the precursors of chlorinated DBPs. Therefore, at a higher molar ratio of bromide to chlorine, the formation of brominated DBPs increased, which is a notable finding because brominated DBPs present higher mutagenicity and carcinogenicity than their chlorinated analogs [47]. Considering the chemical structure of chlorimuron-ethly, other DBPs including nitrosamines (such as nitrosodimethylamine) and HAAs, may also been formed. This should be further investigated because their carcinogenicity is higher than that of chlorinated and regulated DBPs [20, 48].
4. Conclusions (1) The chlorination of chlorimuron-ethyl followed second-order kinetics, first-order kinetics when the chlorine concentration changed, and first-order kinetics when the chlorimuron-ethyl concentration changed. The reaction rate constants of the acid-catalyzed reaction of chlorimuron-ethyl with HOCl and the reaction of chlorimuron-ethyl with HOCl were calculated as 6.41 (±1.62) × 107 M−2 min−1 and 3.25 (±1.14) × 103 M−1min−1, respectively. (2) The degradation rate of chlorimuron-ethyl during chlorination decreased with increasing bromide (at pH 5.5–7.0) and ammonium concentrations (≤ 6 μM), whereas the degradation rate of chlorimuron-ethyl increased with the increasing bromide concentration (at pH 8.5). (3) Chlorimuron-ethyl chlorination is an endothermic reaction. Using the Arrhenius equation, its activation energy was estimated to be 22.46 kJ mol−1. (4) Four volatile DBPs were detected during chlorimuron-ethyl chlorination: CF, DCAN, TCNM and 1,1,1-TCP. CF was the major volatile degradation product, 15
and its concentration increased with increasing pH. Increasing the molar ratio of bromide to chlorine resulted in a shift from the chlorinated DBP species to the brominated species, which may cause higher carcinogenic risks to public health. Therefore, adept strategies should be applied to control DBP formation in water treatment processes.
Acknowledgments This study was supported in part by the Natural Science Foundation of China (No. 51678354 and 51778444), State Key Laboratory of Pollution Control and Resource Reuse Foundation (No. PCRRK16005) and the Ministry of Science and Technology in Taiwan (MOST-104-2221-E-327-001-MY3).
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Environmental
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Technology
2002;36:1980-87. [20] Chen WH, Young TM. Influence of nitrogen source on NDMA formation during chlorination of diuron. Water research 2009;43:3047-56. [21] Xu B, Tian FX, Hu CY, Lin YL, Xia SJ, Rong R, Li DP. Chlorination of chlortoluron: kinetics, pathways and chloroform formation. Chemosphere 2011;83:909-16. [22] Zhang TY, Xu B, Hu CY, Li M, Xia SJ, Tian FX, Gao NY. Degradation kinetics and chloropicrin formation during aqueous chlorination of dinoseb. Chemosphere 2013;93:2662-8. [23] Zang H, Yu Q, Lv T, Cheng Y, Feng L, Cheng X, Li C. Insights into the 18
degradation of chlorimuron-ethyl by Stenotrophomonas maltophilia D310-3. Chemosphere 2016;144:176-84. [24] Gozzi F, Machulek A, Ferreira VS, Osugi ME, Santos APF, Nogueira JA, Dantas RF, Esplugas S, de Oliveira SC. Investigation of chlorimuron-ethyl degradation by Fenton, photo-Fenton and ozonation processes. Chemical Engineering Journal 2012;210:444-50. [25] Hua G, Reckhow DA, Kim J. Effect of bromide and iodide ions on the formation and speciation of disinfection byproducts during chlorination. Environmental Science & Technology 2006;40:3050-56. [26] Hu C-Y, Cheng M, Lin Y-L. Chlorination of bensulfuron-methyl: Kinetics, reaction factors and disinfection by-product formation. Journal of the Taiwan Institute of Chemical Engineers 2015;53:46-51. [27] Clesceri LS, Greenberg AE, Eaton AD. Standard methods for the eamination of water and wastewater. Washington, DC, American Public Health Association, American Water Works Association, Water Environment Federation; 1999. [28] USEPA. Method 551.1. Cincinnati, Ohio 45268, Office of Research and Development, U.S. Environmental Protection Agency; 1995. [29] Xu B, Gao NY, Cheng H, Hu CY, Xia SJ, Sun XF, Wang X, Yang S. Ametryn degradation by aqueous chlorine: kinetics and reaction influences. Journal of hazardous materials 2009;169:586-92. [30] Acero JL, Real FJ, Benitez FJ, Gonzalez M. Kinetics of reactions between chlorine or bromine and the herbicides diuron and isoproturon. Journal of Chemical Technology & Biotechnology 2007;82:214-22. [31] Hu C-Y, Zhu H-Z, Lin Y-L, Zhang T-Y, Zhang F, Xu B. Dissolved organic matter fractions and disinfection by-product formation potential from major raw waters 19
in the water-receiving areas of south-to-north water diversion project, China. Desalination and Water Treatment 2014;56:1689-97. [32] Hu C-Y, Li A-P, Lin Y-L, Ling X, Cheng M. Degradation kinetics and DBP formation during chlorination of metribuzin. Journal of the Taiwan Institute of Chemical Engineers 2017;80:255-61. [33] Liu C, von Gunten U, Croue JP. Enhanced bromate formation during chlorination of
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research 2007;41:1193-200. [40] Chang EE, Chiang PC, Chao SH, Lin YL. Relationship between chlorine consumption and chlorination by-products formation for model compounds. Chemosphere 2006;64:1196-203. [41] Peters RJB, de Leer EWB, de Galan L. Chlorination of cyanoethanoic acid in aqueous medium. Environmental Science & Technology 24:81-86. [42] Dore M, Merlet N, Loat JD. Reactivity of halogens with aqueous micropollutants: a mechanism for the formation of trihalomethanes. J. Am. Water Works Assoc. 1982;74:103-07. [43] Hu J, Song H, Karanfil T. Comparative Analysis of Halonitromethane and Trihalomethane Formation and Speciation in Drinking Water: The Effects of Disinfectants, pH, Bromide, and Nitrite. Environmental Science & Technology 2010;44:794-99. [44] Shan J, Hu J, Kaplan-Bekaroglu SS, Song H, Karanfil T. The effects of pH, bromide and nitrite on halonitromethane and trihalomethane formation from amino acids and amino sugars. Chemosphere 2012;86:323-8. [45] USPEA. National primary drinking water regulations, Stage 2 Disinfectants and disinfection byproducts rule, 71(2), Federal Register; 2006. [46] Westerhoff P, Chao P, Mash H. Reactivity of natural organic matter with aqueous chlorine and bromine. Water research 2004;38:1502-13. [47] Uyak V, Toroz I. Investigation of bromide ion effects on disinfection by-products formation and speciation in an Istanbul water supply. Journal of hazardous materials 2007;149:445-51. [48] Lin Y-L. Effects of physicochemical properties of nanofiltration membranes on the rejection of small organic DBP precursors. Journal of Environmental 21
Engineering 2013;139:127-36.
22
Figure Captions Fig. 1 Peseudo-first-order kinetics plot of chlorimuron-ethyl chlorination at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23 μM, and five different chlorine dosages. Fig. 2 Apparent second-order rate constant of chlorimuron-ethyl degradation during chlorination at different pH levels. Fig. 3 Effect of bromide concentration on the degradation of chlorimuron-ethyl during chlorination at pH 6 (a), pH 7 (b) and pH 9 (c) at 25°C, [chlorimuron-ethyl]0 = 7.23 μM, and [HOCl]T = 108.5 μM. Fig. 4 Effect of ammonium concentration on the degradation of chlorimuron-ethyl during chlorination at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23μM, and [HOCl]T = 108.5 μM. Fig. 5 Effect of temperature concentration on the degradation of chlorimuron-ethyl during chlorination at pH 7, [chlorimuron-ethyl]0 = 7.23 μM, and [HOCl]T = 108.5 μM. Fig. 6 DBP formation as a function of reaction time during chlorination of chlorimuron-ethyl at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23 μM, and [HOCl]T = 108.5 μ. Fig. 7 Chloroform and other DBP formation as a function of pH during chlorination of chlorimuron-ethyl at 25°C, [chlorimuron-ethyl]0 = 7.23 μM, [HOCl]T = 108.5 μM, and reaction time = 7 d. Fig. 8 Effect of bromide concentration on DBP formation during chlorination of chlorimuron-ethyl at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23 μM, [HOCl]T = 108.5 μM, and reaction time = 7 d.
23
[HOCl]T = 72.3 μM, R2=0.9765
3.0
[HOCl]T = 108.5 μM, R2=0.9988 [HOCl]T = 144.6 μM, R2=0.9995
2.5
[HOCl]T = 180.8 μM, R2=0.9991
2.0 0.12
1.5
R2=0.9997
0.10
kobs(min-1)
ln([Chlorimuron-ethyl]/[Chlorimuron-ethyl]0)
[HOCl]T = 53.8 μM, R2=0.9973
1.0
0.08
0.06
0.04
0.5
0.02 60
80
100
120
140
160
180
[HOCl]T (μM)
0.0 0
10
20
30
40
50
Time (min)
Fig. 1
24
60
70
80
90
100
200
Fig. 2
25
ln ([Chlorimuron-ethyl]t/ [Chlorimuron-ethyl]0)
(a) 3.0 2.5 2.0 1.5
[Br-]:[Cl2]=0, R2=0.9984 [Br-]:[Cl2]=0.05, R2=0.9941 [Br-]:[Cl2]=0.1, R2=0.9986 [Br-]:[Cl2]=0.2, R2=0.9933 [Br-]:[Cl2]=0.3, R2=0.9714 [Br-]:[Cl2]=0.5, R2=0.9757
1.0 0.5 0.0 0
5
10
15
20
25
30
Time (min)
ln ([Chlorimuron-ethyl]t/ [Chlorimuron-ethyl]0)
(b)
[Br-]:[Cl2]=0, R2=0.9967 [Br-]:[Cl2]=0.05, R2=0.9991 [Br-]:[Cl2]=0.1, R2=0.9975 [Br-]:[Cl2]=0.2, R2=0.9985 [Br-]:[Cl2]=0.3, R2=0.9984 [Br-]:[Cl2]=0.5, R2=0.9988
3.0 2.5 2.0 1.5 1.0 0.5 0.0 0
10
20
30
40
Time (min)
ln ([Chlorimuron-ethyl]t/ [Chlorimuron-ethyl]0)
(c) [Br-]:[Cl2]=0, R2=0.9977 [Br-]:[Cl2]=0.05, R2=0.9744 [Br-]:[Cl2]=0.1, R2=0.9934 [Br-]:[Cl2]=0.2, R2=0.9955 [Br-]:[Cl2]=0.3, R2=0.9953 [Br-]:[Cl2]=0.5, R2=0.9968
2.0
1.5
1.0
0.5
0.0 0
10
20
30
40
Time (min)
Fig. 3
26
50
60
Fig. 4
27
6.7
Experimental data Fitting for the Arrhrnius equation
6.6
ln(kapp) (M-1min-1)
6.5 6.4 6.3 6.2 6.1 6.0 3.15
3.20
3.25
3.30 -3
3.35 -1
1/T (10 K )
Fig. 5
28
3.40
3.45
100
8 CF TCNM
CF concentration (μg/L)
1,1,1-TCP
6
60 4 40
2 20
0
0 0
1
2
3
4
Time (day)
Fig. 6
29
5
6
7
8
Other DBP concentration (μg/L)
DCAN
80
CF
140
12
DCAN
CF concentration (μg/L)
1,1,1,-TCP
10
100 8 80 6 60 4 40 2
20 0
0 5
6
7
8
pH
Fig. 7
30
9
Other DBP concentration (μg/L)
TCNM
120
100
THM concentration (μg/L)
5
CF BDCM DBCM TBM DCAN BCAN DBAN
80
4
3 60 2 40
HAN concentration (μg/L)
120
1
20
0
0
0.01
0
0.02
0.05
n([Br2]/[Cl2])
Fig. 8
Table 1 Rate constants for the chlorination of different herbicides. Herbicides
Reaction rate constants
T (℃)
References
k1 (M-2 min-1)
k2 (M-1 min-1)
Chlorimuron-ethyl
6.41 (±1.62) × 107
3.25 (±1.14)× 103
25
This study
Ametryn
-
9.48 (±2.04)× 104
25
[29]
Bensulfuron-methyl
-
1.72× 101
25
[26]
Chlortoluron
5.20 (±0.17) × 105
5.18 (±0.65)
25
[35]
Dinoseb
1.20 (±0.48) × 106
1.98 (±0.36) × 102
25
[22]
Metribuzin
7.72 (±0.90) × 108
8.22(±4.00)× 103
25
[32]
31
Highlights
Chlorimuron-ethyl chlorination follows second-order kinetics.
Bromide enhanced the chlorination of chlorimuron-ethyl at pH 8.5, but inhibited it at pH 5.5-7.0.
Four volatile DBPs were identified including CF, DCAN, TCNM and 1,1,1-TCP.
Five brominated DBPs with high carcinogenic risks to public health were identified, including BDCM, DBCM, TBM, BCAN and DBAN.
32
S1383-5866(17)33207-0 https://doi.org/10.1016/j.seppur.2018.04.049 SEPPUR 14545
To appear in:
Separation and Purification Technology
Received Date: Revised Date: Accepted Date:
1 October 2017 14 February 2018 19 April 2018
Please cite this article as: C-Y. Hu, Y-L. Lin, A-P. Li, B. Xu, Degradation kinetics and disinfection by-product formation of chlorimuron-ethyl during aqueous chlorination, Separation and Purification Technology (2018), doi: https://doi.org/10.1016/j.seppur.2018.04.049
This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
Degradation kinetics and disinfection by-product formation of chlorimuron-ethyl during aqueous chlorination
Chen-Yan Hu1, Yi-Li Lin2*, Ai-Ping Li1, Bin Xu3
1. College of Environmental and Chemical Engineering, Shanghai University of Electric Power, Shanghai 200090; P.R. China. 2. Department of Safety, Health and Environmental Engineering, National Kaohsiung University of Science and Technology, Kaohsiung 824, Taiwan, R.O.C. 3. State Key Laboratory of Pollution Control and Resource Reuse, Institute of Disinfection By-product Control in Water Treatment, College of Environmental Science and Engineering, Tongji University, Shanghai 200092, P. R. China.
* Corresponding author: Yi-Li Lin Email: [email protected] Phone: +886-7-6011000#2328 Fax: +886-7-6011061
1
ABSTRACT: The degradation kinetics of chlorimuron-ethyl during chlorination was
analyzed, and the effects of the influencing factors of pH, bromide and ammonium concentrations and reaction temperature on degradation kinetics were evaluated. Regarding degradation kinetics, the chlorination of chlorimuron-ethyl followed second-order kinetics, first-order kinetics when the chlorine concentration was changed, and first-order kinetics when the chlorimuron-ethyl concentration was changed. The apparent second-order rate constant was calculated, which significantly decreased
increasing
pH.
The
rate constant
of the
reaction
between
chlorimuron-ethyl− and HOCl was calculated as 3.25 (± 1.14) × 103 M-1 min-1, and under the acid-catalyzed condition, the rate constant significantly increased to 6.41 (± 1.62) × 107 M−1 min−1. The degradation rate of chlorimuron-ethyl decreased with increasing bromide (at pH 5.5–7.0) and ammonium concentrations (≤ 6 μM). Chlorimuron-ethyl chlorination is an endothermic reaction, with an activation energy estimated as 22.46 kJ mol−1 using the Arrhenius equation. During chlorimuron-ethyl chlorination, chloroform was the major volatile degradation product during chlorimuron-ethyl chlorination, and its concentration increased with the increasing reaction time and pH level. The distribution of disinfection by-products (DBPs) formed at different solution pH levels was quite distinct. In the presence of bromide, the concentration and species of brominated DBPs increased with the increasing bromide concentration. Notably, carcinogenicity and mutagenicity of brominated DBPs were higher than their chlorinate analogs.
Keywords: Chlorimuron-ethyl; Chlorination; Kinetics; Disinfection by-products
(DBPs); Herbicide; Water treatment
2
1. Introduction To increase agricultural production, many different types of herbicides have been extensively applied worldwide to destroy pests and weeds in farmland [1]. Sulfonylurea herbicides are the most widely used herbicides [2]. With the modification of functional groups on the aromatic and heterocyclic rings of sulfonylurea, more than 30 types of sulfonylurea herbicides have been developed, which are widely used globally. In 1988, the US Geological Survey analyzed 214 samples of surface and underground water collected from Midwestern United States. In this survey, nicosulfuron and chlorimuron-methyl were detected in 53% and 31% of the total samples, respectively, and chlorosulfuron, chloropyrazosulfuron and other sulfonylurea herbicides were detected in 5% of the samples [3]. Struger et al. analyzed samoles of surface waters collected from central Canada for 3 consecutive years and reported high concentrations of sulfonylurea herbicides in the samples due to wash-off by rain [4]. The presence of sulfonylurea herbicides in the environment poses potential risks to human health due to their mutagenicity, toxicity and carcinogenicity
[5,
6].
Chlorimuron-ethyl
2-(((((4-chloro-6-methoxypyrimidin-2-yl)
amino)
(the
CAS
carbonyl)
name:
amino)
ethyl
sulfonyl)
benzoate) is a nitrogen-containing herbicide in the family of sulfonylurea herbicides and it is widely used in modern agriculture to control weeds in wheat, rice, and corn due to its high efficiency and the low dosage required for application (2–75 g ha–1) [7, 8]. A concentration of chlorimuron-ethyl as high as 0.3 μg L–1 has been detected in surface water, and a median concentration of < 0.01 μg L–1 has been detected in surface and ground water [9]. In Heilongjiang Province in China, 400 ton of chlorimuron-ethyl is used annually in soybean fields for weed control [10]. Chlorimuron-ethyl is a long-term residual herbicide with a low Kow value of 2.5 [11] 3
and a high water solubility of 1.2 g L–1 at pH 7 [12]; therefore, it can easily be leached from the soil into the surface or groundwater [13, 14] that can be used as a source of drinking water, thus posing a threat to human health. Therefore, identifying the ultimate fate of chlorimuron-ethyl in drinking water treatment process is necessary. Chlorination is an efficient and convenient method of killing waterborne pathogenic microorganisms to guarantee the safety of drinking water. Thus, chlorine is used as a disinfection agent in water treatment plants worldwide [15, 16]. However, chlorine is a strong oxidant and can react with organic matter to form harmful disinfection by-products (DBPs), including trihalomethanes (THMs) and haloacetic acids (HAAs), which are reported to be carcinogenic and mutagenic to animals and humans [17-19]. Several studies have focused on the chlorination of herbicides such as diuron [20], chlortoluron [21] and dinoseb [22], but not chlorimuron-ethyl. Few studies have reported the degradation of chlorimuron-ethyl by microorganisms [23] and through Fenton, photo-Fenton, or ozonation processes [24]. It is essential to determine the degradation kinetics and DBP formation of chlorimuron-ethyl during chlorination, and it is necessary to evaluate the effects of water parameters such as pH, bromide and ammonium concentrations, and temperature on degradation kinetics, which have been demonstrated to have a crucial influence on the degradation of herbicides during chlorination [21, 25]. The results can provide valuable information for the control of chlorimuron-ethyl and corresponding DBP formation in the water treatment process. . The objectives of this study were (1) to investigate the degradation kinetics of chlorimuron-ethyl during chlorination, (2) to study the effect of the influencing factors of pH, bromide and ammonium concentrations, and temperature on chlorimuron-ethyl degradation, and (3) to assess the yield of volatile DBPs during 4
chlorimuron-ethyl chlorination as a function of the reaction time and the pH of the solution.
2. Materials and Methods 2.1. Chemicals and reagents Chlorimuron-ethyl (>99%) and DBP standards (EPA 551A and 551B) including chloroform (CF), bromodichloromethane (BDCM), dibromochloromethand (DBCM), bromoform (TBM), trichloroacetonitrile (TCAN), dichloroacetonitrile (DCAN), bromochloroacetonitrile trichloronitromethane
(BCAN), (TCNM),
dibromoacetonitrile
1,1-dichloro-2-propanone
(DBAN),
(1,1-DCP)
and
1,1,1-trichloroacetone (1,1,1-TCP) were obtained from Sigma-Aldrich (St. Louis, MO, USA) and used as received. Analytical grade of NaOCl (available chlorine 4.00 4.99%), Na2S2O3, NaBr, NH4Cl, phosphate, carbonate, acetate, NaOH, and HCl were purchased from Sinopharm Chemical Reagent Co., Ltd. (Shanghai, China). High performance liquid chromatography (HPLC) grade methanol and methyl tert-butyl ether (MTBE) were purchased from J.T. Baker (USA). All solutions were prepared with Milli-Q water.
2.2. Experimental procedures All experiments were performed in a 200-mL batch reactor equipped with a dispenser and were controlled at 25 ± 1°C in a water bath. The experiments were initiated by adding an aliquot of concentrated chlorine solution to the reactor containing 7.23 μM of chlorimuron-ethyl solution (3 mg L−1) buffered using 10 mM acetate for a pH of 4–5, 10 mM phosphate for a pH of 5–8, and 10 mM carbonate for a pH of 8–10. Small volumes of H2SO4 and NaOH were used to adjust the pH of the 5
solution to the desired initial values. At different reaction times, 1 mL of the solution was rapidly transferred into an HPLC vial containing 50 μLof NH4Cl solution (1 M) to quench the reaction, and the solution was then analyzed using HPLC as soon as possible. The kinetic experiments were conducted over a pH range of 4–10 at the molar ratio of chlorine to chlorimuron-ethyl of 7.4–25.0:1. The chlorimuron-ethyl concentration was monitored until more than 50% of chlorimuron-ethyl was degraded. The effects of bromide and ammonium concentrations (0–3.62 μM and 0–6 μM, respectively) and the reaction temperature (20°C, 25°C, 30°C, 35°C and 40°C) on chlorimuron-ethyl chlorination were evaluated at a chlorimuron-ethyl concentration of 7.23 μM, pH 7 (maintained using phosphate buffer) and a chlorine concentration of 108.6 μM (as Cl2). For conducting kinetic and DBP formation studies, the chlorine concentration should be at least 10 times the concentration of the target substance [20, 21, 26]. Accordingly, in this study, the chlorine concentration was set at 15 times the chlorimuron-ethyl concentration to analyze the effect of bromide and ammonium concentrations, and temperature on chlorimuron-ethyl chlorination; a similar procedure was followed for the DBP formation experiments. Additionally, a blank experiment was conducted. During the experimental period, no chlorimuron-ethyl degradation was observed without the addition of chlorine (data not shown). The
experiments
for
the
determination
of
DBP
formation
during
chlorimuron-ethyl chlorination were conducted in duplicate under headspace-free conditions in 45-mL glass of screw-cap vials with polytetrafluoroethylene-lined septa. A typical experimental condition was 7.23 μM chlorimuron-ethyl solutions dosed with 108.6 μM chlorine dosage at a pH range of 5–10 using 10 mM acetate, phosphate, or carbonate buffers. At designated reaction times, the reaction was 6
quenched using NH4Cl, and water samples then were extracted using MTBE to analyze DBP formation.
2.3 Analytical methods The chlorimuron-ethyl concentration was analyzed using an Agilent 1260 infinity HPLC system equipped with an autosampler and a Waters X Terra C18 column (250 mm × 4.6 mm, 5 μm, Waters, USA), and the concentration was then detected using a UV detector at a wavelength of 254 nm. The mobile phase consisted of 80%/20% (v/v) methanol and Milli-Q water (the pH of 4.5 was maintained by the addition of phosphoric acid). The flow rate was 1 mL min–1, and the injection volume was
5
μL.
The
chlorine
concentration
was
analyzed
using
the
N, N-diethyl-p-phenylenediamine (DPD) colorimetric method [27]. The pH of the solution pH was measured using a regularly calibrated pH meter (Sartorius, Germany). DBP formation was quantified following the method 551.1 of the US Environmental Protection Agency (US EPA) [28] on a gas-chromatograph (GC-2010 Plus, Shimadzu, Japan) equipped with an electron capture detector and a RTX-5 fused silica capillary column (30 m × 0.25 mm id, 0.25-μm film thickness). Helium was used as the carrier gas at a flow-rate of 0.94 mL min–1. The temperature program was as follows: injector temperature of 200°C, detector temperature of 290°C, a column temperature of 37°C (hold for 10 min) that was increased to 50°C (with a rate of 5°C per min), then increased to 260°C (with a rate of 30°C per min) and hold for 10 mins. The detection limits of CF, BDCM, DBCM, TBM, TCAN, DCAN, BCAN, DBAN, TCNM, 1,1-DCP and 1,1,1-TCP were 2, 1, 1, 2, 0.5, 1, 1, 1, 5, 1 and 0.5 μg L-1, respectively. 7
3. Results and Discussion 3.1 Reaction kinetics The chlorination of chlorimuron-ethyl was performed at a chlorimuron-ethyl concentration of 7.23 μM, chlorine concentration of 53.8–180.9 μM, pH 7 and reaction temperature of 25°C. The results are provided in Fig. 1, in which [Chlorimuron-ethyl]T,0
is
the
initial
chlorimuron-ethyl
concentration
and
[Chlorimuron-ethyl]T,t is the chlorimuron-ethyl concentration at different reaction times. A good linear fit (R2 > 0.97) in all experimental data verified the pseudo-first-order reaction kinetics of chlorimuron-ethyl degradation, and the slopes revealed the pseudo-first-order rate constants (kobs). The upper-right insert figure in Fig.1 reveals that kobs is linearly related to the initial total chlorine concentration ([HOCl]T, including HOCl and OCl−) (R2 = 0.999), which indicates that the chlorination of chlorimuron-ethyl is a first-order reaction when the chlorine concentration is increased. Accordingly, the chlorination of chlorimuron-ethyl follows the second-order kinetics, first-order kinetics when the chlorine concentration is changed, and first-order kinetics when the chlorimuron-ethyl concentration is changed; these findings have also been reported for other herbicides [21, 26, 29]. Additional experiments at different initial concentrations of chlorimuron-ethyl were conducted, and the results are provided in Fig. S1 in the Supplementary Information. With the addition of a high concentration of chlorine, the degradation of chlorimuron-ethyl at different initial concentrations followed first-order kinetics. Therefore, the kinetics of chlorimuron-ethyl chlorination are described in Eq. (1), where kapp is the apparent second-order rate constant at specific pH levels, which can be calculated from the slope of the plot of kobs versus [HOCl]T. 8
d[Chlorimu ron - ethyl]T = kapp [HOCl]T [Chlorimuron-ethyl]T dt = kobs [Chlorimuron-ethyl]T
(1)
3.2 Effect of pH on the reaction rates of elementary reactions The reaction rates of reactions between chlorine and organic compounds are highly affected by the pH of the solution [30]. Therefore, the effect of pH on chlorimuron-ethyl chlorination was investigated at initial chlorimuron-ethyl and chlorine concentrations of 7.23 and 108.6 μM, respectively, and the results are depicted in Fig. 2. The pKa of chlorimuron-ethyl is 4.20 [12],and that of NaClO is 7.5. At a pH of > 4.2, most chlorimuron-ethyl molecules will dissociate to chlorimuron-ethyl− so that the total amount of chlorimuron-ethyl species will be represented by the ionized one ([Chlorimuron-ethyl]T≈ [Chlorimuron-ethyl−]). At a pH of ≤ 7.5, the major chlorine species in solution will be HOCl. Therefore, the reactions due to chlorimuron-ethyl degradation during chlorination are as follows: Ka1 = 2.88×10−8
HOCl ↔OCl-+H+ [HOCl]T = [HOCl] + [OCl-]
(2) (3)
Chlorimuron-ethyl ↔ Chlorimuron-ethyl- + H+
Ka2 = 6.31×10−5
(4)
[Chlorimuron-ethyl]T = [Chlorimuron-ethyl] + [Chlorimuron-ethyl−] ≈[Chlorimuron-ethyl−]
(5)
Chlorimuron-ethyl− + HOCl+ H+→ products
k1
(6)
Chlorimuron-ethyl− + HOCl→ products
k2
(7)
Because OCl− has weaker oxidation power than HOCl, and the calculated reaction rate constant of reactions between OCl− and chlorimuron-ethyl was very close to zero (data not shown), the OCl− oxidizing chlorimuron-ethyl− reaction was negligible. Therefore, the degradation of chlorimuron-ethyl can be expressed as: 9
d[Chlorimu ron - ethyl]T = kapp [HOCl]T [Chlorimuron-ethyl]T dt = k1 [Chlorimuron-ethyl−] [HOCl] [H+] + k2 [Chlorimuron-ethyl−] [HOCl]
(8)
After replacing [HOCl] with [HOCl]T/(1 + α) and [Chlorimuron-ethyl−] with [Chlorimuron-ethyl]T, where α is the ionization percentage of chlorine, the degradation of chlorimuron-ethyl can be further expressed as: 2 d[Chlorimu ron - ethyl]T = k1 ([H] ) k 2 [H] [HOCl]T [Chlorimuron-ethyl]T (9) dt Ka1 [H]
Comparing Eqs. (1) and (9), the apparent rate constant kapp can be expressed as: kapp =
k1 ([H] ) 2 k 2 [H] Ka1 [H]
(10)
The rate constants k1 and k2 were calculated as 6.41 (±1.62) × 107 M−1 min−1 and 3.25 (±1.14) × 103 M−1 min−1, respectively, which were compared with the rate constants of several herbicides listed in Table 1. The acid-catalyzed reaction rate constant k1 of chlorimuron-ethyl was 104 times higher than k2, which was similar to the rate constants of chlortoluron, dinoseb and metribuzin chlorination (Table 1). The k2 value of chlorimuron-ethyl was much higher than those of chlortoluron, dinoseb, and bensulfuron-methyl listed in Table 1, which indicates that chlorimuron-ethyl can react with chlorine and consume HOCl much faster than the aforementioned herbicides during chlorination.
3.3 Effect of bromide concentration on chlorimuron-ethyl chlorination The bromide concentrations detected in the Yellow River and Danjiangkou Reservoir in China were 0.75 and 0.48 μM, respectively [31]. During the chlorination of organic compounds, bromide can react with HOCl to form HOBr, leading to the formation of more toxic brominated DBPs than the chlorinated DBPs [17-19]. 10
Therefore, the effect of the bromide concentration on chlorimuron-ethyl chlorination at pH levels of 5.5–8.5 was investigated, and the results are provided in Fig. 3. Chlorimuron-ethyl chlorination at different bromide concentrations over the pH range of 5.5–8.5 followed pseudo-first-order kinetics well. By contrast, bromide had the opposite effect on chlorimuron-ethyl degradation over the pH range of 5.5–7.0 (Fig. 3(a)&(b)) and the pH level of 8.5 (Fig. 3(c)). Bromide reacts with HOCl quickly in an aqueous solution and forms the stronger oxidizing agent, HOBr (Eq. (11)), under basic conditions, which increases the degradation of organic matter [26, 32]. This phenomenon explains the increasing rate of chlorimuron-ethyl degradation when the bromide concentration increased from 0 to 6.0 μM at pH 8.5. However, at a pH of 5.5–7.0, the presence of bromide slightly inhibited chlorimuron-ethyl degradation under the conditions of the increasing bromide concentration and decreasing pH, which may have been due to a series of complex reactions that transferred HOCl and HOBr to less oxidative agents [Eqs. (2) and (11)-(13)] [33, 34]. Further research on the role of bromide in chlorimuron-ethyl chlorination is required. HOCl + Br− → HOBr + Cl−
k3 = 1.55×103 M−1 s−1
(11)
HOBr ↔ OBr− + H+
Ka3 = 1.26 × 10−9
(12)
HOCl + HOBr → BrO3− + Cl- + 3H+
k4
(13)
3.4 Effect of ammonium concentration on chlorimuron-ethyl chlorination Ammonium is widely present in water and can react with chlorine to form chloramines, including monochloramine (NH2Cl), dichloramine (NHCl2) and trichloramine (NCl3) [35]. The formation of trichloramine is negligible at a pH level of > 4.5 [36]. Thus, reactions of chloramine formation only and the corresponding rate constants are expressed in Eqs (14)–(16) [16]. 11
NH4+ ↔ NH3 + H+
Ka4 = 5×10−10
(14)
HOCl + NH3 → NH2Cl + H2O
k6 = 4.2×106 M−1s−1
(15)
HOCl + NH2Cl → NHCl2 + H2O
k7 = 1.1×103 M−1s−1
(16)
The effect of ammonium concentration on the degradation of chlorimuron-ethyl during chlorination was investigated at a pH of 7, and the results are presented in Fig. 4.
The
experimental
data
indicated
that
the
degradation
rate
of
chlorimuron-ethyl during chlorination decreased with the increasing ammonium concentration due to the competitive consumption of HOCl with ammonium (k 2 in Eq. (7) << k6 in Eq. (15). Because monochloramine and dichloramine have weaker oxidation power than HOCl [36], the degradation of chlorimuron-ethyl by chloramines is negligible.
3.5 Effect of temperature on chlorimuron-ethyl chlorination Temperature plays a crucial role in the rate constant, product yield, and species distribution in chemical reactions. To determine the effect of temperature on chlorimuron-ethyl degradation during chlorination, a set of experiments were conducted at temperatures between 20°C and 40°C. The correlation of ln(kapp) and the reciprocal of temperature (T–1) are provided in Fig. 5. As illustrated in Fig. 5, an increasing temperature accelerated the degradation rate of chlorimuron-ethyl, with a good linear fit between ln(kapp) and T–1 (R2 = 0.99). Using the Arrhenius equation, the activation energy of chlorimuron-ethyl chlorination was estimated to be 22.46 kJ mol−1, which was higher than that of bensulfuron-methly chlorination [26]. The results validate that a more efficient reaction occurs between chlorimuron-ethyl and chlorine, as discussed in Section 3.2.
12
3.6 DBP formation during chlorimuron-ethyl chlorination The formation of carbonaceous and nitrogenous DBPs during chlorimuron-ethly chlorination was evaluated, and the results are presented in Fig. 6. Four DBPs were detected: CF, DCAN, TCNM and 1,1,1-TCP. CF was the major volatile degradation product, and the CF concentration increased with the reaction time, reaching a maximum concentration of 84.2 μg L−1 on day 7 due to excess chlorine in the solution. However, the concentrations of DCAN, TCNM, and 1,1,1-TCP were one-order lower than that of CF. TCNM formation increased quickly on day 1 and remained relatively stable until day 7. A similar phenomenon was also observed during the chlorination of metribuzin and bensulfuron-methyl [26, 32], which implied that TCNM is a stable DBP end-product that does not undergo further hydrolysis or oxidation. As for DCAN and 1,1,1-TCP, The maximum concentrations of DCAN and 1,1,1-TCP (4.8 and 5.3 μg L−1, respectively) were reached on day 1, and the concentrations decreased gradually as the reaction time passed, which can be explained by the hydrolysis and oxidation of the DBPs formed by chlorine [37]. According to Yang et al. [38], TCP can be hydrolyzed to CF, and DCAN tends to hydrolyze at a pH of > 7 to produce dichloroacetic acid; this hydrolysis reaction is accelerated in the presence of free chlorine [39]. This may also explain why DCP and TCAN were not detected in this study. Different pH levels exert considerable influence on the dissociation of HOCl, the degree of halogenation, and the hydrolysis reactions during organic compound chlorination, which results in the formation of DBPs. Therefore, the effect of the pH of the solution on DBP formation during chlorimuron-ethyl chlorination for 7 days was investigated, and the results are provided in Fig. 7. CF was the dominant DBP, and its concentration increased significantly from 31.9 to 137.4 μg L−1 as the pH level 13
increased from 5 to 9. This phenomenon can be explained by the OH− requirement in the haloform formation process as well as the hydrolysis of haloacetonitriles (HANs) or haloketones under basic conditions [40-42]. The concentrations of DCAN and 1,1,1-TCP decreased continuously as the pH level increased from 5 to 9, which can be explained by base-catalyzed hydrolysis decomposition, and the hydrolysis rate can be accelerated by increasing the pH level [38]. TCNM formation could be due to the cleavage of the aromatic ring and side chains containing carbonyl and dimethylamine groups during chlorimuron-ethyl chlorination, which changed slightly as the pH increased from 5 to 9 [37]. The negligible effect of pH on TCNM formation has also been reported by other researchers during chlorination of other nitrogenous precursors [43, 44]. Bromide participation in chlorimuron-ethyl chlorination may result in the formation of more toxic brominated DBPs than chlorinated analogs. Therefore, the effect of the bromide concentration on DBP formation during chlorimuron-ethyl chlorination at a pH of 7 for 7 days was investigated, and the results are provided in Fig. 8. The total concentrations of THMs and HANs increased as the molar ratio of bromide to chlorine increased from 0 to 0.05. The THM concentration reached 90.81 μg L-1, which exceeded the standard concentration of 80 μg L-1 set in the National Primary Drinking Water Regulations of the US EPEA [45] at the molar ratio of bromide to chlorine of 0.01. The presence of bromide during chlorimuron-ethyl chlorination resulted in the formation of five more brominated DBPs: BDCM, DBCM, TBM, BCAN and DBAN, which may be due to the reaction between HOBr and the benzene and heterocyclic rings of chlorimuron-ethyl. The reaction rate between HOBr and organic matter were dozens of times higher than that between HOCl and organic matter [46]. Thus, with the increase of bromide concentration, more HOBr will be 14
produced and subsequently reacteded with chlorimuron-ethly. This reaction will consume the precursors of chlorinated DBPs. Therefore, at a higher molar ratio of bromide to chlorine, the formation of brominated DBPs increased, which is a notable finding because brominated DBPs present higher mutagenicity and carcinogenicity than their chlorinated analogs [47]. Considering the chemical structure of chlorimuron-ethly, other DBPs including nitrosamines (such as nitrosodimethylamine) and HAAs, may also been formed. This should be further investigated because their carcinogenicity is higher than that of chlorinated and regulated DBPs [20, 48].
4. Conclusions (1) The chlorination of chlorimuron-ethyl followed second-order kinetics, first-order kinetics when the chlorine concentration changed, and first-order kinetics when the chlorimuron-ethyl concentration changed. The reaction rate constants of the acid-catalyzed reaction of chlorimuron-ethyl with HOCl and the reaction of chlorimuron-ethyl with HOCl were calculated as 6.41 (±1.62) × 107 M−2 min−1 and 3.25 (±1.14) × 103 M−1min−1, respectively. (2) The degradation rate of chlorimuron-ethyl during chlorination decreased with increasing bromide (at pH 5.5–7.0) and ammonium concentrations (≤ 6 μM), whereas the degradation rate of chlorimuron-ethyl increased with the increasing bromide concentration (at pH 8.5). (3) Chlorimuron-ethyl chlorination is an endothermic reaction. Using the Arrhenius equation, its activation energy was estimated to be 22.46 kJ mol−1. (4) Four volatile DBPs were detected during chlorimuron-ethyl chlorination: CF, DCAN, TCNM and 1,1,1-TCP. CF was the major volatile degradation product, 15
and its concentration increased with increasing pH. Increasing the molar ratio of bromide to chlorine resulted in a shift from the chlorinated DBP species to the brominated species, which may cause higher carcinogenic risks to public health. Therefore, adept strategies should be applied to control DBP formation in water treatment processes.
Acknowledgments This study was supported in part by the Natural Science Foundation of China (No. 51678354 and 51778444), State Key Laboratory of Pollution Control and Resource Reuse Foundation (No. PCRRK16005) and the Ministry of Science and Technology in Taiwan (MOST-104-2221-E-327-001-MY3).
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in the water-receiving areas of south-to-north water diversion project, China. Desalination and Water Treatment 2014;56:1689-97. [32] Hu C-Y, Li A-P, Lin Y-L, Ling X, Cheng M. Degradation kinetics and DBP formation during chlorination of metribuzin. Journal of the Taiwan Institute of Chemical Engineers 2017;80:255-61. [33] Liu C, von Gunten U, Croue JP. Enhanced bromate formation during chlorination of
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Engineering 2013;139:127-36.
22
Figure Captions Fig. 1 Peseudo-first-order kinetics plot of chlorimuron-ethyl chlorination at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23 μM, and five different chlorine dosages. Fig. 2 Apparent second-order rate constant of chlorimuron-ethyl degradation during chlorination at different pH levels. Fig. 3 Effect of bromide concentration on the degradation of chlorimuron-ethyl during chlorination at pH 6 (a), pH 7 (b) and pH 9 (c) at 25°C, [chlorimuron-ethyl]0 = 7.23 μM, and [HOCl]T = 108.5 μM. Fig. 4 Effect of ammonium concentration on the degradation of chlorimuron-ethyl during chlorination at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23μM, and [HOCl]T = 108.5 μM. Fig. 5 Effect of temperature concentration on the degradation of chlorimuron-ethyl during chlorination at pH 7, [chlorimuron-ethyl]0 = 7.23 μM, and [HOCl]T = 108.5 μM. Fig. 6 DBP formation as a function of reaction time during chlorination of chlorimuron-ethyl at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23 μM, and [HOCl]T = 108.5 μ. Fig. 7 Chloroform and other DBP formation as a function of pH during chlorination of chlorimuron-ethyl at 25°C, [chlorimuron-ethyl]0 = 7.23 μM, [HOCl]T = 108.5 μM, and reaction time = 7 d. Fig. 8 Effect of bromide concentration on DBP formation during chlorination of chlorimuron-ethyl at 25°C, pH 7, [chlorimuron-ethyl]0 = 7.23 μM, [HOCl]T = 108.5 μM, and reaction time = 7 d.
23
[HOCl]T = 72.3 μM, R2=0.9765
3.0
[HOCl]T = 108.5 μM, R2=0.9988 [HOCl]T = 144.6 μM, R2=0.9995
2.5
[HOCl]T = 180.8 μM, R2=0.9991
2.0 0.12
1.5
R2=0.9997
0.10
kobs(min-1)
ln([Chlorimuron-ethyl]/[Chlorimuron-ethyl]0)
[HOCl]T = 53.8 μM, R2=0.9973
1.0
0.08
0.06
0.04
0.5
0.02 60
80
100
120
140
160
180
[HOCl]T (μM)
0.0 0
10
20
30
40
50
Time (min)
Fig. 1
24
60
70
80
90
100
200
Fig. 2
25
ln ([Chlorimuron-ethyl]t/ [Chlorimuron-ethyl]0)
(a) 3.0 2.5 2.0 1.5
[Br-]:[Cl2]=0, R2=0.9984 [Br-]:[Cl2]=0.05, R2=0.9941 [Br-]:[Cl2]=0.1, R2=0.9986 [Br-]:[Cl2]=0.2, R2=0.9933 [Br-]:[Cl2]=0.3, R2=0.9714 [Br-]:[Cl2]=0.5, R2=0.9757
1.0 0.5 0.0 0
5
10
15
20
25
30
Time (min)
ln ([Chlorimuron-ethyl]t/ [Chlorimuron-ethyl]0)
(b)
[Br-]:[Cl2]=0, R2=0.9967 [Br-]:[Cl2]=0.05, R2=0.9991 [Br-]:[Cl2]=0.1, R2=0.9975 [Br-]:[Cl2]=0.2, R2=0.9985 [Br-]:[Cl2]=0.3, R2=0.9984 [Br-]:[Cl2]=0.5, R2=0.9988
3.0 2.5 2.0 1.5 1.0 0.5 0.0 0
10
20
30
40
Time (min)
ln ([Chlorimuron-ethyl]t/ [Chlorimuron-ethyl]0)
(c) [Br-]:[Cl2]=0, R2=0.9977 [Br-]:[Cl2]=0.05, R2=0.9744 [Br-]:[Cl2]=0.1, R2=0.9934 [Br-]:[Cl2]=0.2, R2=0.9955 [Br-]:[Cl2]=0.3, R2=0.9953 [Br-]:[Cl2]=0.5, R2=0.9968
2.0
1.5
1.0
0.5
0.0 0
10
20
30
40
Time (min)
Fig. 3
26
50
60
Fig. 4
27
6.7
Experimental data Fitting for the Arrhrnius equation
6.6
ln(kapp) (M-1min-1)
6.5 6.4 6.3 6.2 6.1 6.0 3.15
3.20
3.25
3.30 -3
3.35 -1
1/T (10 K )
Fig. 5
28
3.40
3.45
100
8 CF TCNM
CF concentration (μg/L)
1,1,1-TCP
6
60 4 40
2 20
0
0 0
1
2
3
4
Time (day)
Fig. 6
29
5
6
7
8
Other DBP concentration (μg/L)
DCAN
80
CF
140
12
DCAN
CF concentration (μg/L)
1,1,1,-TCP
10
100 8 80 6 60 4 40 2
20 0
0 5
6
7
8
pH
Fig. 7
30
9
Other DBP concentration (μg/L)
TCNM
120
100
THM concentration (μg/L)
5
CF BDCM DBCM TBM DCAN BCAN DBAN
80
4
3 60 2 40
HAN concentration (μg/L)
120
1
20
0
0
0.01
0
0.02
0.05
n([Br2]/[Cl2])
Fig. 8
Table 1 Rate constants for the chlorination of different herbicides. Herbicides
Reaction rate constants
T (℃)
References
k1 (M-2 min-1)
k2 (M-1 min-1)
Chlorimuron-ethyl
6.41 (±1.62) × 107
3.25 (±1.14)× 103
25
This study
Ametryn
-
9.48 (±2.04)× 104
25
[29]
Bensulfuron-methyl
-
1.72× 101
25
[26]
Chlortoluron
5.20 (±0.17) × 105
5.18 (±0.65)
25
[35]
Dinoseb
1.20 (±0.48) × 106
1.98 (±0.36) × 102
25
[22]
Metribuzin
7.72 (±0.90) × 108
8.22(±4.00)× 103
25
[32]
31
Highlights
Chlorimuron-ethyl chlorination follows second-order kinetics.
Bromide enhanced the chlorination of chlorimuron-ethyl at pH 8.5, but inhibited it at pH 5.5-7.0.
Four volatile DBPs were identified including CF, DCAN, TCNM and 1,1,1-TCP.
Five brominated DBPs with high carcinogenic risks to public health were identified, including BDCM, DBCM, TBM, BCAN and DBAN.
32