Degradation of sulfadiazine in water by a UV/O₃ process: performance and degradation pathway

Abstract

In this study, the performance of a combined UV/O₃ process for aquatic sulfadiazine (SDZ) removal was investigated. By comparing with UV irradiation or ozonation, the UV/O₃ process showed excellent performance for SDZ removal, particularly on the mineralization of SDZ. The degradation rate of SDZ by UV/O₃ increased with the increment of O₃ gas concentration and UV intensity. Meanwhile, initial solution pH also played an important role in the UV/O₃ process. The pH increment of solutions (3.0–9.0) could promote the SDZ degradation rate which was mainly ascribed to the generation of ˙OH by O₃ self-decomposition and the dissociation of SDZ. But this phenomenon was seriously reversed for the production of more free radical scavengers like CO₃²⁻ as initial pH increased from 9.0 to 11.0. Two different degradation pathways of SDZ by UV/O₃ were proposed based on the combination of theoretical calculations and experimental intermediate identification. Pathway I was initiated by hydroxyl radicals which involved oxidative cleavage of the S–N bond and pathway II was induced by direct UV irradiation which involved the opening of the pyrimidine ring. According to the results obtained in this work, UV/O₃ was recommended as an effective method to remove SDZ.

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Introduction

Antibiotics have been widely used in humans and animals for the prevention and treatment of infection and diseases as medicines, and also used as a growth promoter in the livestock industry.¹ Most recently, Zhang et al.² have reported that about 92 700 tons of 36 typical antibiotics were utilized in China. Besides, the estimated annual dosages of antibiotics for humans and animals have reached 2000 tons in Germany.³

Generally, used antibiotics are excreted as metabolites, conjugates, or in their original parent form and released to the wastewater system.⁴ Moreover, most antibiotics cannot be removed by conventional wastewater treatment processes and are ultimately discharged into receiving water.⁵ Thus, more and more researchers have reported that various antibiotics were detected in water matrices at trace levels.^6–8 The sulfa drugs or sulfonamides, as the most ancient preventative antibiotics for the treatment of infections in humans and veterinary purposes,⁹ were repeatedly detected in aquaculture. Even though the detective concentration of sulfonamides in water environment is at the ng L⁻¹ to μg L⁻¹ level, the residual sulfonamides can induce significant changes in the biosphere,¹ including acute or chronic toxic effects on organisms, development of antibiotic-resistant bacteria and alteration of aquatic flora.¹⁰ Therefore, it is necessary to adopt alternative process to remove sulfonamides from water.

Recent decades, ozonation was proposed to be a suitable process for the removal of various contaminates including antibiotics,^11,12 dyes,¹³ and textile wastewater,¹⁴ due to the strong oxidation ability of O₃ molecule (E₀ = 2.07 V). Garoma et al.¹⁵ have reported that O₃ can completely remove sulfonamides from aqueous solution. However, low O₃ utilization rate and low mineralization in O₃ process have been two barriers limiting the application of O₃ process. To overcome these disadvantages, various strategies have been used in combination with ozone to enhance the oxidation process, particularly UV radiation. UV/O₃ process has shown to be an available option as it has the advantages of oxidizing and destroying organic pollutants in water by combining the beneficial effects of ozonation (direct O₃ reactions) with the generation of hydroxyl radicals (˙OH).^16,17 The generation of ˙OH involves several chemical processes.¹⁸ The decomposition of aqueous O₃ in pure water has been firstly initiated by hydroxide ion (eqn (1)) and this reaction leads to the production of radicals that accelerates decomposition process by chains of radical reactions and produce ˙OH (eqn (2)–(4)).^19,20 Furthermore, hydrogen peroxide is produced in these processes, thus UV/O₃ oxidizing system also involves the behaviour of UV/H₂O₂ and O₃/H₂O₂. Although UV/O₃ has already applied to degrade many refractory containminents,^21–23 the performance of combined UV/O₃ process on sulfonamides degradation is also obscure and the mechanism or pathway of sulfonamides should be sufficiently confirmed as well.

O₃ + OH⁻ → HO₂⁻ + O₂ (1)

O₃ + H₂O₂ → HO₂⁻ + ˙OH + O₂ (2)

The aim of our study is to investigate the removal efficiency of a typical sulfonamide (sulfadiazine, SDZ) by UV/O₃ process under different operational conditions in aqueous solution. In addition, degradation pathways of SDZ in UV/O₃ process were also proposed simultaneously based on the theoretical calculation and experimental detection.

Experiment and calculation methods

Materials and reagents

Sulfadiazine (C₁₀H₁₀N₄O₂S, 99% purity) was purchased from TCI technology CO. Ltd. and used as received without further purification. O₃ was generated from pure oxygen gas (99.9%) by an O₃ generator (DHX-SS-1G, Harbin Jiu Jiu Electrochemistry Engineering Ltd., Harbin, China). Low-pressure mercury lamps (Haoyan, China) emitting primarily at 254 nm were used as UV-radiation source and placed at the center of the reactor. All organic solvents used for analysis were HPLC grade. Other chemical reagents, such as phosphoric acid and sodium hydroxide were analytical grade (Tianjin chemical Works Co., Tianjin, China). All solutions were prepared using deionized water.

Experimental approach

Batch experiments were carried out in a 1.0 L capacity cylindrical glass reactor with the dimensions of 10 cm in diameter, 16 cm height. The reactor was covered by aluminium foil to avoid the release of radiation. The initial pH value was adjusted to 7.0 by phosphoric acid and sodium hydroxide, and not corrected in the experiment. Before the experiment, UV lamp was turned on to warm up at least 20 minutes to make sure stable output. The solution was stirred continuously using a magnetic stirrer. For UV/O₃ experiments, appropriate numbers of lamps were inserted into the reactor. At time zero, lamps were turned on and O₃ was continuously bubbled into the reactor at the rate of 0.4 L min⁻¹ through a diffuser that was placed at the bottom of the reactor. The off gas was absorbed by potassium iodide solution. The ozone gas concentration could be adjusted by changing the voltage of the machine. Experiments were conducted at room temperature and the increase due to UV irradiation was less than 2 °C. For ozonation experiments, UV lamps were switched off and the experiments were conducted as described as above. For UV experiments, the procedure was the same as UV/O₃ process except without O₃ gas input. The initial concentration of SDZ was 25 mg L⁻¹ unless other specified. The range of gas O₃ concentration was 30.0–43.8 mg L⁻¹ and of UV intensity was 0.30–0.90 mW cm⁻². Aliquot samples were withdrawn periodically and analysed for residual concentration of SDZ, intermediates and total organic carbon (TOC). The samples involved in O₃ were quenched with 10 μL of sodium thiosulfate (1.0 mol L⁻¹).

Analytical methods

The average UV intensity was measured by the method of potassium ferrioxalate actinometry.²⁴ The O₃ gas concentration was measured by the potassium iodide method.²⁵ SDZ was analyzed by ultra performance liquid chromatography (UPLC) (Waters ACQUITY UPLC system) equipped with a C18 column (2.1 × 50 mm, 1.7 μm; Waters) interfaced with a TUV detector. The mobile phase was a mixture of acetonitrile (A) and 0.1% formic acid solution (B) and the ratio was 15 : 85. The detection wavelength was set at 265 nm. The injection volume of the sample was 2 μL and the flow rate was 0.1 mL min⁻¹. The degradation intermediate products of SDZ removal by UV/O₃ were separated by the Waters ACQUITY UPLC but interfaced with a tandem quadrupole mass detector (Xevo TQD) (Waters ACQUITY UPLC-MS-MS, Milford MA, United States). Mass spectral analysis was conducted in positive mode electrospray ionization ((+)ESI) over a mass range of 50–500 m/z. The cone voltage used was 30 V conducted in auto full scan mode (MS2). Details of the UPLC method description were in accord with the conditions of SDZ quantification mentioned above.

Computational details

Molecular calculations were performed to facilitate the understanding of SDZ degradation mechanisms by UV/O₃ process. The reaction model used in the computational part of this study was the reaction between the SDZ molecule and the ˙OH radical. Hartree–Fock (HF) methods suffer from spin contamination and it is not suitable for the reaction system involved in the ˙OH radical. In contrast to the HF methods, density functional theory (DFT) methods used the exact electron density instead of the wave function to calculate molecular properties and energies. They suffer from spin contamination less than HF methods and this feature makes them suitable for calculations involved ˙OH radical system. Therefore, geometry optimizations of the reactants were performed with the DFT method.²⁶ The DFT calculations were carried out as implemented in Gaussian 09 code,²⁷ using the exchange–correlation functional B3LYP in combination with the 6-31 (g, p) basis set. The IEFPCM model was adopted to evaluate solvent effect of water. Vibrational frequencies were calculated for the determination of the structures as stationary points and true minima on the potential energy surfaces.

Results and discussion

Degradation of SDZ by UV/O₃

As shown in Fig. 1, comparative experiments in pure water were conducted to investigate the degradation efficiency and mineralization of SDZ by O₃ alone, UV alone, and UV/O₃ processes where the O₃ gas concentration was 38.4 mg L⁻¹, average UV intensity was 0.3 mW cm⁻², and the initial pH value was 7.0. Direct UV could only remove 20% SDZ after 10 min irradiation and little TOC removal rate after 30 min irradiation, and the result was consistent with previous study.²⁸ Compared with UV process, O₃ and UV/O₃ processes showed much higher SDZ removal rate. O₃ has strong oxidation ability and it could easily react with SDZ.¹⁵ In the UV/O₃ process, it is believed that UV irradiation could increase the O₃ usage rate and generate more ˙OH which enhanced the degradation of SDZ.^29,30

Fig. 1 Degradation of SDZ by UV, O₃ and UV/O₃ processes. The inset shows TOC mineralization of SDZ removal by UV, O₃ and UV/O₃ processes after 30 min. Experimental conditions: O₃ gas concentration = 38.3 mg L⁻¹, UV intensity = 0.3 mW cm⁻², initial SDZ concentration = 25 mg L⁻¹ and initial pH = 7.0.

Although SDZ could be eliminated by O₃ within a short time, TOC removal rate of SDZ degradative process was only 12% after 30 min ozonation indicating that SDZ could not be completely mineralized by O₃ and many degradation intermediates might accumulate in the system. However, for the mineralization of SDZ, O₃ and UV/O₃ make great differences as shown in the inset of Fig. 1. It was found that UV/O₃ process exhibited a better performance in TOC removal rate than individual O₃ process after 30 min treatment. The advantages of UV/O₃ in mineralization of SDZ were mainly ascribed to multiple chemical reactions involved as previously described and the generation of ˙OH which owned a high oxidation potential and was less discriminating in the types of attacking functional groups.^31,32 Thus, the UV/O₃ process is a feasible way to eliminate and mineralize SDZ and intermediates from water and it would be meaningful to clarify the factors which significantly influence the performance of UV/O₃ process.

Key factors affecting SDZ degradation in the UV/O₃ system

Effect of influent O₃ gas concentration. The effect of influent O₃ gas concentration on SDZ removal in the UV/O₃ system is presented in Fig. 2. The removal rate of SDZ was enhanced with the increase of O₃ gas concentration from 30.0 to 43.8 mg L⁻¹ as expected. Obviously, dissolved O₃ concentration in water transferred from gas-phase O₃ significantly increased as influent O₃ gas concentration increased. SDZ could be directly destroyed by dissolved O₃. Meanwhile, more hydroxyl radicals generated by the self-decomposition of dissolved O₃ at pH 7.0 (eqn (1)) and the reaction under UV irradiation (eqn (2)) could also contribute to the degradation of SDZ.

Fig. 2 Removal of SDZ by UV/O₃ process at different O₃ concentrations. Experimental conditions: UV intensity = 0.3 mW cm⁻²; initial SDZ concentration = 25 mg L⁻¹; initial pH = 7.0.

Effect of UV intensity. As shown in Fig. 3, the effect of UV intensity on SDZ removal in the UV/O₃ system is investigated. According to the results, the removal rate of SDZ was slightly improved with the increase of UV intensity. Excess UV radiation might accelerate the self-decomposition of dissolved O₃ generating more hydroxyl radicals in solution. However, the promotion of ozone self-decomposition by UV radiation was also limited with the continuous increase of UV intensity. In order to achieve both degradation efficiency and energy saving, the applied UV intensity was set at 0.30 mW cm⁻² in this study.

Fig. 3 Removal of SDZ by UV/O₃ process at different UV intensities. Experimental conditions: O₃ concentration = 38.4 mg L⁻¹; initial SDZ concentration = 25 mg L⁻¹; initial pH value = 7.0.

Effect of initial solution pH. In general, pH played an important role in O₃-based processes as pH decided ozonation types consisted of direct oxidation by molecular O₃ under acidic condition and radical oxidation by ˙OH under alkaline condition.^33–35 In order to understand the influence of initial pH on the UV/O₃ system performance, SDZ removal in the range of pH values from 3.0 to 11.0 are investigated in Fig. 4. It was found that the degradation rate of SDZ was increased with the increment of pH ranged from 3.0 to 9.0 but decreased at pH 11.0. In O₃-based processes, OH⁻ could promote the self-decomposition of O₃ and the generation of ˙OH possessing the stronger oxidation ability than molecular O₃. However, higher pH in the solution leaded to the creation of more free radical scavengers (i.e. CO₃²⁻, HCO₃⁻) derived from mineralization of organic matters, resulting in a decrease of ˙OH in turn^36,37 to restrain the oxidation process.

Fig. 4 Removal of SDZ by UV/O₃ process at different pH values. Experimental conditions: O₃ concentration was 38.4 mg L⁻¹; UV intensity = 0.30 mW cm⁻², initial SDZ concentration = 25 mg L⁻¹.

Besides, pH of the solution also decided existing form of SDZ considering the chemical nature of the SDZ molecule. SDZ is characterized by having two acid equilibrium constant values (pK_a1 = 2.49 and pK_a2 = 6.48) for the presence of amidogen and hydroxyl. Hence, deprotonated amidogen as nucleophilic group gradually dominated over the pH range of 3.0 to 11.0; moreover, dissociated hydroxyl as electrophilic group could be dominated as pH increased to 6.48. It is known that deprotonated amidogen and dissociated hydroxyl which both possessed more electron could be more easily attacked by ˙OH and molecular O₃. Thus, the acid–base species transformation of SDZ also explained the increment of SDZ degradation rate with the increasing of pH in the UV/O₃ system. In addition, SDZ degradation rate was slightly promoted with the increase of pH ranged from 7.0 to 9.0, but neutral solution was much convenient for direct disposal of wastewater. So, pH 7.0 was considered as the most appropriate value for SDZ oxidation by UV/O₃.

Analysis of degradation pathway of SDZ by UV/O₃ process

Molecular calculations. Natural charge distribution calculation was employed to predict the vulnerable sites of SDZ with respect to ˙OH radical attack and to determine the intermediates.³⁸ The optimized SDZ molecule is shown in Fig. 5 and the calculated values are listed in Table 1. The lengths of bonds in optimized molecule were shown in Table S1.† According to the results, 7(N), 9(O), 10(O) and 17(N) had relatively bigger negative values. Correspondingly, the connected bonds of these atoms were easily broken under UV/O₃ process. This conclusion was double checked with the basis 6-311+ g(d) and the results was consistent with this. The charges of atoms under basis 6-311+ g(d) were shown in Table S2.†

Fig. 5 Optimized SDZ molecule structure.

Table 1 Charges (C_i) of atoms in SDZ molecule

Atoms	Ci	Atoms	Ci	Atoms	Ci	Atoms	Ci
1C	0.07	8S	2.364	15C	−0.291	22H	0.268
2C	−0.343	9O	−0.958	16C	−0.191	23H	0.256
3C	0.068	10O	−0.978	17N	−0.828	24H	0.255
4N	−0.525	11C	−0.407	18H	0.245	25H	0.273
5C	0.592	12C	−0.199	19H	0.267	26H	0.428
6N	−0.531	13C	−0.289	20H	0.245	27H	0.428
7N	−0.892	14C	0.203	21H	0.469

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Identification of intermediates and possible reaction pathway

To elucidate the degradation pathways of SDZ oxidation by UV/O₃, the reaction intermediates were detected by UPLC-MS/MS. Six compounds ([M + H]⁺) were identified with peaks of m/z 251, m/z 187, m/z 118, m/z 114, m/z 96 and m/z 85. The peak of m/z 251 represented parent SDZ. Product 1 (P1) was identified as 2-pyrimidinamine with m/z 96 and product 2 (P2) was identified as hydroxylated 2-pyrimidinamine. Besides, product 3 (P3) was identified as 4-amino-N-methylbenzenosulfonamide with m/z 187 and product 4 (P4) was identified as 3-imino-propen-1-one oxime with m/z 85. Product 5 (P5) was identified 2-hydroxy-3-hydroxyimino-acrylimidic acid with m/z 118.

According to analytical intermediates above, degradation pathways of SDZ oxidation by UV/O₃ are illustrated in Fig. 6. The degradation of SDZ by UV/O₃ involved two possible pathways (I and II). In pathway I, as predicted by molecular calculations, the generation of P1 and I_a (m/z 174) were ascribed to the cleavage of S–N bond by ˙OH attack. This process could also be found in SDZ removal by other oxidation systems such as, gamma irradiation process and Fenton process.^38,39 Then ˙OH added to pyrimidine ring of P1 resulting in the formation of P2. In pathway II, the first step for the degradation of SDZ in the UV/O₃ system involved the open of pyrimidine ring generating P3 and P4. Then ˙OH took place of hydrogen atoms of P4 to form P5. However, according to the calculation results, pyrimidine ring was not the prior target which ˙OH radicals could attack.

Fig. 6 Possible degradation pathways of SDZ by UV/O₃.

In order to figure out the mechanism, the reaction intermediates of SDZ oxidation by individual UV radiation and O₃ were detected. The peak of m/z 187 was not detected in O₃ process, meanwhile, only two intermediates with peak of m/z 187 and m/z 114 were found in individual UV radiation. Thus, we concluded that the pathway II might be initiated by direct UV radiation not by ˙OH. Moreover, SDZ oxidation by UV/O₃ producing more abundant intermediates than by alone UV or O₃ indicated that the UV/O₃ system possessed advantages in mineralization of contaminants.

It should be mentioned that although I_a had only been deduced as possible intermediate of the SDZ degradation (not directly detected), the proposed mechanism can be substantiated by considering that they were quickly eliminated by the ˙OH, thereby preventing accumulation in the system.

Conclusions

SDZ removal using combined UV/O₃ process was investigated and its possible degradation pathway was induced based on both theoretical calculation and experimental detection. Results showed that the UV/O₃ process was an effective method to degrade SDZ from water. In addition, it performed better considering the degradation efficiency and mineralization compared with UV and O₃ alone processes under same conditions. Effects of O₃ gas concentration, UV intensity, and initial solution pH value on SDZ removal by the UV/O₃ process were investigated. Solution pH played an important role in UV/O₃ process as it not only affected the decomposition of aqueous O₃ and the ˙OH production, but also decided the existing forms of SDZ in solution.

Natural population analysis was used to predict the possible degradation pathway of SDZ in the UV/O₃ process. And calculation results showed that 7(N), 9(O), 10(O) and 17(N) had relatively bigger negative values and the connected bonds of these atoms were easily broken under UV/O₃ process. By combining the analytical results of detected intermediates by UPLC-MS, two degradation pathways were deduced. In pathway I, degradation of SDZ was initiated by the broken of S–N bond under the attack of ˙OH radical. And, pathway II might be initiated by the direct UV radiation based on similar intermediates detected in UV alone irradiation but not in ozonation processes. The UV/O₃ system could produce more easy-to-degrade intermediates than UV or ozonation alone in SDZ degradation which implied that the UV/O₃ system be more suitable for the mineralization of such organic pollutants.

Acknowledgements

This work was financially supported by National science and technology plan of China (2014BAD02B03), and Fund for young top-notch talent teachers by Harbin Institute of Technology (AUGA5710052514). The authors also gratefully acknowledge the financial support by State Key Laboratory of Urban Water Resource and Environment (ES201602).

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Footnote

† Electronic supplementary information (ESI) available. See DOI: 10.1039/c6ra09078h

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