Yaoguang Guoa,
Zhiyuan Liuab,
Xiaoyi Lou*b,
Changling Fangb,
Pu Wanga,
Genying Wuc and
Jie Guan*a
aSchool of Environmental and Materials Engineering, Shanghai Polytechnic University, Shanghai 201209, China. E-mail: guanjie@sspu.edu.cn
bLaboratory of Quality Safety and Processing for Aquatic Product, East China Sea Fisheries Research Institute, Chinese Academy of Fishery Sciences, Shanghai 200090, China. E-mail: huoxingmayi@126.com
cLongquan Branch of Lishui Municipal Ecological Environment Bureau, Longquan 323700, China
First published on 20th April 2021
Antibiotic addition and chlorination are two common processes in fishery culture. Antibiotic residues not only pollute aquaculture water, but are also one of the potential precursors of disinfection by-products (DBPs) during chlorination. The degradation kinetics, products identification and reaction mechanism of sulfacetamide (SFA), a new sulfonamides antibiotics, and potential formation of haloacetic acids (HAAs) in chlorination were explored. The results showed that the degradation of SFA followed pseudo first-order kinetic model, and chlorinating agent dose, pH of water, water temperature, NH4+, HCO3− and humic acid (HA) had various effects on the degradation of SFA and the yields of HAAs. The presence of Br− accelerated both the degradation rate of SFA and more formation of Br-DBPs. Through the identification of intermediate products, we proposed the transformation pathway of SFA during the chlorination disinfection process. Namely, in this NaClO disinfection system, the C–S bond between the sulfonyl group and benzene ring, and S–N bond between sulfonyl and acylamino of SFA were broken, and then the primary formed groups were further oxidized to produce intermediates, such as chloroanilines and chlorophenols. And then chlorophenols were subsequently chlorinated to form toxic HAAs. The present study might be of significance for the evaluation of effective degradation of SFA and potential production of halogenate-DBPs (H-DBPs) during the chlorination disinfection process in aquaculture water.
Chlorination is the most common process in aquaculture water, with the disinfectants of chlorine compounds.9 Residual antibiotics are degraded during the chlorination process; however, studies have found that antibiotics undergo side reactions with chlorine disinfectants to generate halogenated disinfection by-products (H-DBPs) such as trihalomethanes (THMs) and haloacetic acids (HAAs).10,11 These H-DBPs have attracted extensive attention in the field of scientific research owing to the latent health risks.12 In seawater, hypobromous acid (HOBr) can be generated by the chlorine oxidation to be dominant oxidant species, leading to preferential formation of various brominated by-products.13,14 Rong et al.15 reported that norfloxacin (NOR) in seawater was substituted with HOBr to form two brominated DBPs (Br-DBPs) during chlorination, and the toxicity of Br-DBPs was far higher than that of chlorinated DBPs (Cl-DBPs). Marsà et al.16 reported that bromoacetic acid was 89.8 times higher than that of chloroacetic acid in human urothelial cells. THMs exposure resulted in adverse effect for upgrowth and significant tail shortening in zebrafish, and tribromoacetic acid (TBAA) and dichloroacetic acid (DCAA) can significantly increase the deformity rate of zebrafish embryos.17,18 In brief, possible derivative H-DBPs from antibiotics during chlorination disinfection process in aquaculture water might threaten health to aquatic animals, and subsequently intimidate the fishery economy. Therefore, the transformation fate of H-DBPs in aquaculture water during disinfection is significant for effective assessment of fishery health risk.
Sulfonamides (SAs) are broad-spectrum antibiotics used in aquaculture waters, and the most related researches focus on the degradation kinetics during chlorination disinfection process.19 In contrast, the transformation mechanism and the potential generation of H-DBPs in the chlorination process still need to be further studied to explore the threat to fisheries aquaculture. Consequently, in the present study, we selected a new SA antibiotic (sulfacetamide, SFA) as the model substrate (the detailed information was shown in Table S1 in the ESI†). We studied the relationship between reaction kinetics of SFA in chlorination and the parameters, such as available chlorine concentration, pH, temperature, and other common components (i.e. NH4+, HCO3− and HA) in aquaculture water. Based on the above analysis, the potential formation of H-DBPs represented by HAAs and transformation mechanism of SFA in the chlorination process were explored. The present research might provide technical support for the effective removal of SFA and the control of the production of HAAs during the chlorination process in aquaculture water.
To examine the potential formation of HAAs, 40 mL of chlorinated water sample for 24 h was transferred to a 100 mL centrifuge tube, and then 2 mL of concentrated sulfuric acid was added to make the pH of the solution lower than 0.5, and subsequently 9 g of baked sodium chloride was quickly added to dissolve with shaking. The water samples were then extracted twice with 4 mL MTBE, and each extraction was shaken for 5 min and left to stand for 5 min, ultimately the extracts were combined. Afterwards, 3 mL of the extract was transferred into a 50 mL centrifuge tube, and after adding 3 mL of a newly prepared 10% of sulfuric acid methanol solution, mixed well and placed it into 50 °C water bath for 2 h for derivatization. After the derivative water sample was cooled, 7 mL of 250 g L−1 sodium chloride solution was added, quickly followed shaken up. After leaving the water sample quietly for 5 min, a pipette was used to remove the water phase, and 1 mL of saturated sodium bicarbonate solution was slowly added to the organic layer to exhaust the gas. Finally, 1 mL of the upper organic phase was placed in a 1.5 mL brown sample bottle for GC-ECD analysis.
The analytical method to determine HAAs refers to ‘water quality-determination of haloacetic acids-gas chromatography’ (HJ 758-2015), a national environmental protection standard of the People's Republic of China. An Agilent 7890A gas chromatography spectrometer (GC) equipped with a DB-5 column (30 m × 0.32 mm×0.25 μm) and an electron capture detector (ECD) was used to determine the concentrations of HAAs in NaClO disinfection process. The injections were made in the splitless mode using an injection temperature of 210 °C and the injection volume was 1 μL. Nitrogen was used as carrier gas and make-up gas, the flow rate of carrier gas was 2.5 mL min−1. The temperature of ECD detector was 260 °C. The GC column was operated in a temperature programmed mode with an initial temperature of 40 °C held for 5 min, ramp first to 65 °C with a 5 °C min−1 rate, then to 75 °C with 1 °C min−1 rate, then to 135 °C with 5 °C min−1 rate, and then to 280 °C with 20 °C min−1 rate and held at that temperature for 5 min. The external standard method was used for quantitative analyses of halogenated methyl acetate derivatives of the nine HAAs mixed standards after methyl esterification pretreatment. The organic intermediates produced during the SFA degradation were identified by gas chromatography-mass spectrometry (GC-MS, Shimadzu GCMS-QP2010 plus) and the details are shown in Text S1.†
[Cl2] + SFA → products | (1) |
The rate equation of reaction (1) should be:
![]() | (2) |
The degradation of SFA at a concentration of 5 μM was examined at initial [Cl2] concentrations of 50, 75, 100, and 125 μM. Fig. S2† and 1 show that the degradation effect of SFA increased with the increase of the concentration of [Cl2], and the degradation of SFA satisfied the pseudo first-order kinetics with R2 > 0.998 (Fig. 1a). In the actual disinfection of aquaculture water, the concentration of disinfectant is usually much more than that of antibiotic.21 In the present study, the initial mole ratio of SFA to [Cl2] was set at lower than 1:
10. Therefore, the concentration of [Cl2] was far excessive and can be approximated as constant compared to SFA in the whole chlorination process. Therefore, the reaction kinetics between disinfectants [Cl2] and antibiotics SFA are recognized as pseudo first-order model. As such, eqn (2) can be simplified to eqn (3).22
![]() | (3) |
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Fig. 1 (a) Effect of [Cl2] dose on the degradation of SFA, and (b) the linear relationship between kobs and [Cl2] dose. Experimental conditions: [SFA]0 = 5 μM, pHini = 5, T = 20 °C. |
We can also observe that at pH 5 and temperature of 20 °C, as the concentration of chlorine disinfectant increased, the reaction rate constant kobs were 0.01044, 0.02298, 0.03783, and 0.05596 s−1, respectively, and there was a linear relationship between kobs and [Cl2] dose (R2 = 0.993) (Fig. 1b). Hence, it can be noted that the concentration of [Cl2] was an important factor for accelerating reaction kinetics.
The degradation of SFA in NaClO disinfection system at pH 5–9 was presented in Fig. S3† and 2, and the results of the experiment showed that the reaction rate of SFA was the fastest under neutral condition, and slowed down under weak acid or weak base condition. The observed pseudo first-order kinetic constants of SFA degradation first increased from 0.01044 to 0.03003 s−1 with the pH extended from 5 to 7, and then decreased to 0.00273 s−1 when pH further reached to 9. The macroscopic Ka1 and Ka2 constants of SFA (Fig. S4†) are related to amino and amide groups, respectively. The pKa1 and pKa2 of SFA were 1.76 and 5.22 (Table S1†),23 and the pKa of HOCl was 7.54,24 indicating both SFA and some available chlorine species existed in molecular state (i.e. SFA and HOCl) at pH 5, respectively. Although HOCl had strong oxidizability under acidic conditions, the molecular state of SFA had relatively weaker ability to provide electrons for molecular HOCl. When the pH value was in the range from 5.22 to 7.49, SFA was in the form of deprotonated state, and HOCl existed in the molecular state, thus, the deprotonated amide group of SFA can provide electrons more easily to HOCl with higher oxidizability. Furthermore, trace Cl2O, potent chlorinating agent at pH ≲ 8 dependent on the existence of HOCl (eqn (4)), is more potent electrophiles than HOCl/ClO−.25 As such, the electrophilic action to SFA by slight Cl2O can explain the effective removal of SFA and the increased reaction rate under neutral conditions. In alkaline conditions (pH > 7.49), SFA existed in deprotonic state and the available chlorine species mainly existed in the form of ClO−, however, the oxidative capbility and eletrophilicity of ClO− was much weaker than that of HOCl and Cl2O,24,25 which led to the slower degradation of SFA.
![]() | (4) |
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Fig. 2 (a) Oxidation of SFA by [Cl2], and (b) the variation of kobs at different pH values. Experimental conditions: [SFA]0 = 5 μM, [Cl2]0 = 50 μM, T = 20 °C. |
The temperature had a great influence on the degradation rate of SFA, and the degradation of SFA at high temperature was faster than that at low temperature (Fig. S5†). Fig. 3a shows the pseudo first order kinetic constants of SFA increased from 0.00507 to 0.01994 s−1 with the temperature elevated from 10 to 40 °C. The effect of temperature on chemical reaction rate can also be expressed by Arrhenius equation (eqn (5)).26
![]() | (5) |
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Fig. 3 (a) Effect of temperature, and (b) Arrhenius plot. Experimental conditions: [SFA]0 = 5 μM, [Cl2]0 = 50 μM, pHini = 5. |
According to Fig. 3b, the calculated observed activation energy was 35.14 kJ mol−1. Zhang et al.15 reported that the observed activation energies of the reactions of NOR and roxithromycin (ROX) with [Cl2] were 105.00 and 79.34 kJ mol−1, respectively. In contrast, the obtained observed activation energy of [Cl2] to oxidize SFA is much lower, indicating the reaction rate is faster under the same experimental conditions, and the reaction between SFA and [Cl2] is easier to occur from the perspective of activation energy as well.
NH4+ and HCO3− are common cation and anion in aquaculture water.27 The presence of NH4+ can inhibit the oxidation of SFA by [Cl2] (Fig. S6a† and 4a). With the concentration of NH4+ increased from 0 to 5 mg L−1, the kobs decreased from 0.01044 to 0.00076 s−1. When the concentration of NH4+ reached 5 mg L−1, only less than 40% of SFA was degraded within 10 min (Fig. S6a†). These phenomena might be attributed that HOCl could react with NH4+ to form NH2Cl, NHCl2 and NCl3 (eqn (6)–(9)),28,29 whose oxidation abilities were weaker than that of HOCl,30 leading to the decreased oxidation of SFA (Fig. S6a†).
NH4+ + H2O → NH3 + H3O+ | (6) |
HOCl + NH3 → NH2Cl + H2O | (7) |
HOCl + NH2Cl → NHCl2 + H2O | (8) |
HOCl + NHCl2 → NCl3 + H2O | (9) |
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Fig. 4 Effects of concentrations of (a) NH4+, (b) HCO3−, and (c) HA on the oxidation of SFA by [Cl2]. Experimental conditions: [SFA]0 = 5 μM, [Cl2]0 = 50 μM, pHini = 5, T = 20 °C. |
Fig. S6b† and 4b showed the HCO3− had little effect on SFA degradation. With the concentration of HCO3− increased from 0 to 100 mg L−1, kobs varied much more slightly and negligibly, i.e. only from 0.01044 to 0.01403 s−1. This phenomenon indicated that HCO3− had neglected influence on the transformation of SFA.
Humic acid (HA) is the main component of natural organic matter, which accounts for 50–90%.31,32 HA is commonly found in natural water bodies, and the concentration in water bodies ranges from 0 to 30 mg L−1.33 Therefore, the effects of HA on the degradation of SFA by [Cl2] were also evaluated. Fig. S6c† and 4c showed the negligible effect of altered concentrations of HA on the oxidation of SFA by [Cl2]. When the concentration of HA increased from 0 to 15 mg L−1, the changes of kobs were ignorable (Fig. 4c). At the same time, we measured the concentration changes of HA during the reaction. The results showed that HA hardly degraded in 420 s. [Cl2] with relatively low redox potential (e.g. EHOCl = 740 mV), has limited oxidation capacity to degrade HA in the examined 420 s.34
HOCl + Br− → HOBr + Cl− | (10) |
HOBr + SFA → products | (11) |
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Fig. 5 Effects of (a) different water bodies and (b) Br− concentration on SFA degradation. Experimental conditions: [SFA]0 = 5 μM, [Cl2]0 = 50 μM, pHini = 5, T = 20 °C. |
As can be seen from Fig. 6b, the formation of DCAA and TCAA increased with the pH rise, the reason for which might be alkaline pH promoted the reaction equilibrium movement towards the direction for formation of chloroacetic salts, further improving the generation of HAAs. In addition, the temperature also has a positive influence on the yield of HAAs (Fig. 6c). During the chlorination process, the concentrations of DCAA and TCAA gradually increased, and the total concentration of HAAs enlarged from 4.63 to 18.92 μg L−1, as the temperature increased from 10 to 40 °C. The degradation kinetics of SFA indicated the obtained lower observed activation energy encouraged the easily occurred reaction between SFA and [Cl2], possibly supporting the susceptive formation of HAAs.
Fig. 6d showed the effect of different types of aquaculture waters on the generation potential of HAAs. The production of DCAA and TCAA in marine culture water and seawater were much more than that in freshwater. Furthermore, the Br-HAAs, such as BCAA, BDCAA, DBAA, CDBAA and TBAA, can be detected in both marine culture water and seawater (Fig. 6d and S8†), which might attribute to high concentration Br− participating in the generation of higher oxidative HOBr (eqn (9)), further accelerating the formation of Br-HAAs (Fig. S8†).40
The transformation process of SFA in NaClO disinfection system can be divided into three steps: (1) with the oxidation of [Cl2], the C–S bond between sulfonyl group and benzene ring and S–N bond between sulfonyl and acylamino groups of SFA were scissored, and then the intermediate sulfonyl group formed p1(chloromethanesulfonyl chloride), and the rest intermediate aminobenzene were oxidized to produce p2 (4-hydroxyaniline), p3 (4-chlorophenol) and p4 (2-hydroxyaniline).20,43 (2) Due to the chlorination mechanism in NaClO system, the intermediate products p2 and p4 were substituted by chlorine to produce p5 (4-chloroaniline), p6 (2,4-dichloroaniline) and p7 (2-amino-4-chlorophenol). The intermediate product p3 was substituted by chlorine to form polychlorophenols (p8–p10).44 (3) The chlorophenols could be cleaved under the action of [Cl2], and then fully combined with Cl in the subsequent chlorination process to produce HAAs and other products.12,41,42
Footnote |
† Electronic supplementary information (ESI) available. See DOI: 10.1039/d1ra01605a |
This journal is © The Royal Society of Chemistry 2021 |