- Email: [email protected]
K2S2O8: Efficiency, mechanism and byproducts cytotoxicity
Journal of Environmental Management 225 (2018) 224–231
Contents lists available at ScienceDirect
Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman
Research article
Sulfonamides degradation assisted by UV, UV/H2O2 and UV/K2S2O8: Efficiency, mechanism and byproducts cytotoxicity
T
A. Acosta-Rangela,b,∗, M. Sánchez-Poloa, A.M.S. Poloa, J. Rivera-Utrillaa, M.S. Berber-Mendozab a
Department of Inorganic Chemistry, Faculty of Science, University of Granada, 18071 Granada, Spain Center of Postgraduate Research and Studies, Faculty of Engineering, University Autonomous of San Luis Potosí, Av. Dr. M. Nava No. 8, San Luis Potosí, S.L.P., 78290, Mexico b
A R T I C LE I N FO
A B S T R A C T
Keywords: Sulfonamides Intermediates Ultraviolet radiation Cytotoxicity
The objective of this study was to analyze the effectiveness of UVC, UVC/H2O2 and UVC/K2S2O8 on the degradation of SAs. Rate constant values increased in the order SMZ < SDZ < SML and showed the higher photodegradation of sulfonamides with a penta-heterocycle. Quantum yields were 1.72 × 10−5 mol E−1, 3.02 × 10−5 mol E−1, and 6.32 × 10−5 mol E−1 for SMZ, SDZ and SML, respectively, at 60 min of treatment. R254 values show that the dose habitually utilized for water disinfection is inadequate to remove this type of antibiotic. The initial sulfonamide concentration has a major impact on the degradation rate. The degradation rates were higher at pH 12 for SMZ and SML. SMZ and SML photodegradation kλ values are higher in tap versus distilled water. The presence of radical promoters generates a greater increase in the degradation rate, UVC/ K2S2O8 cost less energy, a mechanism was proposed, and the degradation by-products are less toxic than the original product.
1. Introduction
determination parameter quantic (direct photolysis) or compare the efficacy of the methods (indirect photolysis). These methods are characterized by generation of oxidizing species such as OH% or SO4%- radicals, which have a redox potential of 1.8–2.7 V and 2.5–3.1 V, respectively, with the latter being a more selective oxidant (Yang et al., 2017), have been successfully applied to destruct organic micropollutants in different water matrices (Souza et al., 2016). The major of studies using the Vibrio fischeri bacteria test to determine the toxicity of SAs (Białk-Bielińska et al., 2011; Sági et al., 2018). Knowledge of their byproducts cytotoxicity is still basic and restricted to just a few SAs (Cizmas et al., 2015). HEK 293 cells lack the expression of hormone receptors an important characteristic for studying basal toxicity associated with drugs (Pomati et al., 2006, 2007). The studies that using UV photolysis and its combination with PDS to eliminated sulfonamides still deficient in terms of mechanism, by-products, toxicity and quantification of radicals generated. Therefore, in this study, we evaluate the effectiveness of UVC photolysis, UVC/H2O2 and UVC/PDS in the oxidation of three SAs; sulfamethazine (SMZ), sulfadiazine (SDZ) and sulfamethizole (SML). The objectives of our work were to: i) A kinetic study was conducted to determine the quantum yield (Φλ ) of the UVC photolysis, ii) The influence of different operational variables (initial SAs concentration, oxidant dosage, solution pH, and water matrix) was
Ultraviolet radiation (UVC-254nm) is frequently applied to disinfect water intended for human consumption and wastewaters because it is effective to remove organic pollutants due to occupies a small space, and is readily managed and maintained (Lian et al., 2015; OcampoPérez et al., 2010; Prados-Joya et al., 2011). Sulfonamides (SAs) are widely used as antibiotics in human and veterinary medicine due to their strong antimicrobial activity, stable chemical properties, and low cost (Magureanu et al., 2015; Wang and Wang, 2016). SAs have been detected in surface waters at concentrations of 148–2978 ng L−1 (Iglesias et al., 2012). Prolonged exposure to low concentrations of antibiotics that accumulate in waters can be toxic for cells or lead to resistance in bacterial strains, posing a major public health problem (Baran et al., 2006). Previous studies have found that the direct photolysis based in the fluence quantification can be photodegraded only part of SAs during the UVC disinfection of water, SAs with penta-heterocycle are degraded faster (Li et al., 2017; Yang et al., 2017). By UVCactivated hydrogen peroxide (H2O2), peroximonosulfate (PMS) and peroxydisulfate (PDS) the SAs oxidation have been studied (Babić et al., 2015; Cui et al., 2016; Zhang et al., 2016b; Zhu et al., 2016). However, these efforts have focused in developed and optimized methods to
∗ Corresponding author. Center of Postgraduate Research and Studies, Faculty of Engineering, University Autonomous of San Luis Potosí, Av. Dr. M. Nava No. 8, San Luis Potosí, S.L.P., 78290, Mexico. E-mail address: [email protected] (A. Acosta-Rangel).
https://doi.org/10.1016/j.jenvman.2018.06.097 Received 2 November 2017; Received in revised form 15 February 2018; Accepted 30 June 2018 0301-4797/ © 2018 Elsevier Ltd. All rights reserved.
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Table 1 Molecular structure and physicochemical properties of the studied SAs. Molecular weight (g mol−1)
LogKow
Solubility (mg/L)
pKa
Sulfamethazine C12H14N4O2S (SMN)
278.33
0.19
1500 (29 °C)
pKa1: 2.00 pKa2: 6.99
Sulfadiazine C10H10N4O2S (SDZ)
250.278
−0.09
77 (25 °C)
pKa1: 2.01 pKa2: 6.99
Sulfamethizole C9H10N4O2S2 (SML)
270.33
0.54
1050 (37 °C)
pKa1: 1.95 pKa2: 6.71
Sulfonamide (SA)
Molecular structure
evaluated, iii) The quantification of OH% and SO4%- radicals generated was determined, iv) Degradation by-products were identified, and v) by-product cytotoxicity in cell HEK-293 was measured.
(2) where kob is the mean value of the slopes,
2. Material and methods
, are the reaction rate constants, and are the SMZ reaction rate constants with each radical. The photodegradation rate constant was determined by applying a pseudo-first-order model and considering 1 h as reaction follow-up time, the quantum yield (Φλ ) of each SA, the apparent photodegradation rate constant normalized by the lamp energy (k ′E ) (constant independent of fluctuations), the molar absorption capacity (ελ ) and R254 parameter representing the elimination percentages of SAs in water treatment plants to 400 J m−2, the radiant energy emitted by the lamp was 8.49 × 10-4 E m−2 to 254 nm determined by actinometry (Canonica et al., 1995; Prados-Joya et al., 2011; Sharpless and Linden, 2003).
2.1. Chemical substances All chemical reagents used were high-purity analytical grade reagents supplied by Sigma-Aldrich. All solutions were prepared with ultrapure water obtained using the Milli-Q® system (Millipore). Table 1 lists the molecular structures and physicochemical properties of the sulfonamides under study. 2.2. SAs degradation by UVC-254nm The experimental system for SAs degradation kinetics was in a photoreactor equipped with a low-pressure (254 nm Hg) TNN 15/32 Heraeus Noblelight mercury lamp (nominal power 15 W). Solutions were deposited in six quartz tubes 1 cm in diameter with 35 mL capacity, which were placed in parallel and equidistant to the Hg lamp. The tubes were immersed in recirculating distilled water to maintain a constant temperature of 25 °C, using a Frigiterm ultra-thermostat and magnetic agitation system in each tube. SAs initial concentration was 15 mg L−1, samples were drawn from each solution at regular time intervals for subsequent measurement of the SAs concentration. The influence of experimental parameters such as initial pH (1.5, 6.5 and 12), initial SMZ concentration (5, 10 and 15 mg L−1), oxidant dosage (1.4 × 10−4 and 4.4 × 10−4 M) and water matrix (distiller and tap) was investigated.
2.4. Quantification of OH% and SO4%- radical concentration The concentration of OH% radical generated was calculated using PCBA. PCBA degradation with OH% radicals were quantified in the UV radiation system, in which radicals are responsible for the oxidation. Total concentrations of OH% radical at different time points were calculated the same way as other authors (Velo-Gala et al., 2017). The concentration of SO4%- radicals generated were quantified by identifying the reaction by-product, benzoquinone (BQ) (Oh et al., 2017), obtained through oxidation of p-hydroxybenzoic acid (HBA) by the SO4•- radicals generated in the system under UV radiation. Based on reaction stoichiometry (equation (3)), 1 mol HBA reacts with 1 mol SO4%- to form hydroquinone, which immediately transforms into a stable by-product, BQ, due to excess PDS. The excess PDS is added outside the UV radiation system.
2.3. Determination of the rate constant of SMZ with OH% and SO4%- radical The rate constant of SMZ with OH% and SO4%- radical was determined in competitive kinetics experiments, using para-chlorobenzoic acid (PCBA) as reference compound for OH% (kOH%/ 9 −1 −1 S ) (Kelner et al., 1990) at molar ratio of 1:1 with pCBA=5.2 × 10 M excess H2O2, and using citarabine (CTB) as reference compound for SO4%- ( ) (Ocampo-Pérez et al., 2010) at molar ratio of 1:1 with excess K2S2O8. When ln against ln
(
[CTB ó pCBA] [CTB ó pCBA]0
)
(
[SMZ ] [SMZ ]0
(3) In general, BQ concentrations are proportional to SO4%- concentrations. The BQ generated is relatively stable up to pH 9.0, when excess CHBA˃CBQ, because HBA inhibits the reaction between BQ and SO4%-, whose constants are .
) is plotted
2.5. Analytical methods
, depending on the case, the slope of the graph 2.5.1. Determination of SAs, PCBA, CTB, and BQ concentration in aqueous solution Concentration of the SAs was determined by reverse-phase highperformance liquid chromatography (HPLC), using a liquid chromatographer (Thermo-Fisher) equipped with visible UV-detector and
(kob ) allows the reaction constants of kOH•/ SMZ and kSO4• −/ SMZ radicals to be calculated according to Equations (1) and (2): (1)
225
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
automatic autosampler with capacity for 120 vials. A PHENOMENEX Kinetex C-18 column was used (2.6 μm particle size; 4.6 × 150 mm). The mobile phase was 70% formic acid (0.1%, v/v) and 30% acetonitrile (0.1% v/v) in isocratic mode at a flow of 0.35 mL min−1; the detector wavelength was set at 270 nm. The same methodology was used to determine PCBA and CTB concentrations except that a wavelength of 240 nm with 97% formic acid (0.4%, v/v) was used for PCBA and a wavelength of 272 nm with 3% methanol for CTB. BQ was determined at a wavelength of 244 nm using 50% methanol and 50% deionized water. 2.5.2. Identification of degradation by-products Degradation by-products of SMZ were identified by using Acquity ultra performance liquid chromatography (UPLC) system (Waters) equipped with a CORTECS™ C18 column (2.1 × 75 mm, 2.7 μm) (Waters). The mobile phase, in gradient mode (baseline: 0% B, T8: 95% B, T8.1: 0% B), was channel A, water with 0.1% formic acid and channel B, acetonitrile with 0.1% formic acid, at flow of 0.4 mL min-1, injection volume of 10 μL, and column temperature of 40 °C.
Fig. 1. Degradation kinetics of the three SAs by UVC photolysis. [SAs]0 = 15 mg L−1, pH = 7 and T = 298 K. The lines represent the prediction of the first-order kinetic model.
exposure times are required for this purpose in water treatment plants. 2.5.3. Evaluation of by-products cytotoxicity By-products cytotoxicity were evaluated by using an MTS assay to determine the % viability of human embryo kidney cells (HEK-293) from the CIC cell bank of the University of Granada. Degradation kinetics were first studied in the presence of phosphate buffered saline (PBS), and 10,000 HEK-293 cells were then incubated for 24 h, subsequently changing the medium and adding SAs by-products (10:100 μL dilution); after incubation for a further 24 h, 20 μL MTS was added and its absorbance was measured at 2 h using INFINITENANOQUA equipment, with a sample reading of 9 times at 490 nm absorbance.
3.2. Influence of the different operational variables on UVC photolysis of SAs 3.2.1. Influence of the initial SAs concentration SMZ was selected as model SAs due to show the lowest degradation rate. Fig. 2 and Table 3 shows that the Φλ (1.72 × 10−5, 2.24 × 10−5 and 3.75 × 10−5 mol E−1) and therefore kλ (4.10 × 10−4, 5.95 × 10−4 and 9.16 × 10−4 s−1) is influenced by the initial concentration (15,10 and 5 mg L−1, respectively). The reduced photodegradation rate at higher concentrations is related to the energy absorbed by each SMZ molecule (Wan and Wang, 2017). A fixed energy per volume unit is deposited in the medium, and molecules can accept more radiant energy at lower concentrations because there are fewer molecules in the medium and their availability (Rahmani et al., 2014).
3. Results and discussion 3.1. Direct UVC photolysis of SAs: quantum yields The quantum yield (Φλ ) is a critical factor to quantify the photon efficiency of photolysis reactions (Zhang et al., 2018). Table 2 shows the values of the study parameters. The Φλ for SMZ, SDZ and SML degradation by UVC photolysis was found to be 1.72 × 10−5,3.02 × 10−5 and 6.02 × 10−5 mol E−1, respectively. Which is low values than reported by other authors to degradation of SAs (Lian et al., 2015; Luo et al., 2018) and similar to other (Baeza and Knappe, 2011), due to energy irradiated by the lamp (1.027 × 10−4 E s−1 m−1 for this lamp). Fig. 1 shows the degradation kinetics constant, the trend of decreasing photolysis rate as SML > SDZ > SML. SML was degraded rapidly (1.31 × 10−5 s−1), this is evident due to include a penta-heterocycle attributed to differences in the electron densities of the heterocyclic rings (Cui et al., 2016). Molar absorption coefficient (ελ ) values of SMZ, SDZ and SML were determined to be 1.68 × 103, 1.34 × 103 and 1.46 × 103 m2 mol−1, respectively. These values suggest that SAs may undergo direct photolysis to 254 nm (Ji et al., 2018). R254 is a parameter to determines the applicability UV radiation under real condition (Prados-Joya et al., 2011). The low R254 values (Table 2) obtained to SAs indicate that the dose (400 J m−2) commonly used for water disinfection is inadequate to remove SAs, demonstrating that higher UV radiation doses or longer
3.2.2. Influence of solution pH The solution pH values selected for study were 1.5, 6.5, and 12, at which SAs are in cationic (is protonated at its amine group), neutral, and anionic (deprotonated in NH group) form, respectively. Fig. 3 depicts the variation in global molar absorption coefficient (ελ ) and global photodegradation rate constant (k ′E ) as a function of solution pH. The values of ελ increase in order with the pH, being higher when the SAs are in their anionic form (Fan et al., 2015). The degradation rates were higher at pH 12 for the SMZ and SML, consistent with major value ελ .
Table 2 Parameters obtained from the direct UVC photolysis of the three SAs at 254 nm. Sulfonamide
ελ 10−3 (m2 mol−1)
kλ 104 (s−1)
Φλ 105 (mol E−1)
kE′ 102 (m2 E−1)
R254 103 (%)
SMZ SDZ SML
1.681 1.342 1.464
4.10 5.74 13.13
1.72 3.02 6.32
6.67 9.33 21.32
5.62 7.05 18.07
Fig. 2. Influence of initial SMZ concentration on photodegradation with UVC radiation. pH = 7 and T = 298 K. The lines represent the prediction of the firstorder kinetic model. 226
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Table 3 Parameters obtained with direct UVC radiation of SMZ at 254 nm.
Table 5 Results obtained by UVC photolysis of SMZ at 254 nm in distilled and tap water.
[SMZ]0
ελ 10−3 (m2 mol−1)
kλ 104 (s−1)
Φλ 105 (mol E−1)
kE′ 102 (m2 E−1)
R254 103 (%)
15 10 5
1.681 1.646 1.723
4.10 5.95 9.16
1.72 2.24 3.75
6.67 8.51 14.90
5.62 7.21 12.61
Sulfonamide
Water
pH
Φλ 105 (mol E−1)
kE′ 102 (m2 E−1)
R254 103 (%)
% SAs degradation (60 min)
SMZ
Distilled Tap Distilled Tap Distilled Tap
6.7 8.2 6.2 8.0 5.7 8.2
1.72 3.54 3.01 1.70 3.32 16.11
6.63 17.90 8.32 7.87 21.32 63.50
5.62 15.20 7.05 6.67 18.07 53.90
77.10 98.05 87.41 82.08 99.12 100
SDZ SML
3.4. Indirect UVC photolysis of SAs 3.4.1. SMZ photodegradation in presence of different dose of H2O2 The combined use of UV/H2O2 is known to break the H2O2 molecule, producing two OH% radicals for each molecule and each photon absorbed (Reaction (4)) due to the high quantum yield ( Φ= 0.98 mol E −1) (Velo-Gala et al., 2017). (4) The effect of the H2O2 on the oxidative degradation of SMZ is presented in Table 6. The percentage SMZ degradation after 60 min of treatment at higher initial H2O2 concentrations, rising from 77% in the absence of H2O2 to 100% at an initial H2O2 concentration of 4.4 × 10−4 M. The presence of H2O2 enhances degradation by forming highly oxidizing radical species (Reactions (5)–(10)), which react with SMZ and therefore favor a higher degradation rate. (5)
Fig. 3. Molar absorption coefficient (ελ ) and normalized apparent photodegradation constant of energy emitted by the lamp (kE' ) as a function of the initial solution pH. [SAs]0 = 15 mg L-1, T = 298 K.
Protonated SAs exhibit low reactivity toward electrophilic radicals but deprotonated are easily oxidized by reactive radicals (Cui et al., 2016). However, for SDZ the degradation rate decrease with increase the pH and ελ similar behavior was founded by other authors (Boreen et al., 2004).
(6) (7) (8) (9) RH + HO%→by – products degradation
3.3. Applicability of UVC photolysis of SAs in different water matrix
They can also react with other species besides SMZ (Reactions (11)–(13)) (Kurniawan et al., 2006; Sun and Bolton, 1996), but these reactions reduce the OH% species in the medium, diminishing the efficacy of the system.
The applicability of UVC radiation to remove SAs in aqueous solution was determined by identifying the influence of the chemical composition (Table 4) of tap water and distilled water on their photodegradation. HCO3− is considered to be a scavenger of free radicals (Tan et al., 2014), nevertheless is reacted with electron-rich species such as anilines (Zhao et al., 2018). Several studies mentioned that SO42− (Burgos-Castillo et al., 2018), NO3− (Reina et al., 2018) or halogens (Zhu et al., 2018) can contribute to pharmaceutical compounds degradation. The results are exhibited in Table 5, revealing a higher degradation rate of SMZ (98%) and SML (100%) in tap versus distilled water. SDZ show different behavior similar to influence of pH. Tap water has a lower transmittance value, favoring UVC radiation absorption and markedly reducing the number of photons reaching the SAs (Prados-Joya et al., 2011). In conclusion, the presence of nitrates, sulfates, and chlorates in tap water exert a synergic effect on SAs degradation.
(11) (12) OH% + OH−→ O%– + H2O
Distilled water Tap watera a
[HCO3−] (mg L−1)
[SO42−] (mg L−1)
[Cl−] (mg L−1)
[NO3−] (mg L−1)
T (%)
6.87
–
–
–
–
100
8.20
160
21
5
2
98.17
pH
(13)
The reaction rate constant of the OH% radical with SMZ was determined to identify the influence of these recombination reactions, obtaining a kSMZ/OH% value of 4.35 × 109 M−1 s−1, very similar to reported by other authors (Li et al., 2017). Given that the reaction rate constant of the OH% radical is lower with H2O2 (3.3 × 107 M−1s−1) (Reaction 5) than with SMZ, degradation of SMZ by OH% radicals may be favored until the initial H2O2concentration is reached. Table 6 Kinetic parameters obtained from the degradation of SMZ by indirect UVC in presence of radical promoting species. [SMZ]0 = 15 mg L−1, T = 298 K, time = 60 min.
Table 4 Chemical composition of the water matrix. Water
(10)
Data obtained from: www.emasagra.es/ESP/191.asp. 227
[H2O2]0 104 (M)
[PDS]0 104 (M)
k 104 (M−1 s−1)
% Degradation
1.47 4.4 0 0 0
0 0 1.47 4.4 0
6.20 12.27 4.42 12.89 4.10
89 100 80 100 77
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
3.4.2. SMZ photodegradation in presence of different dose of PDS PDS (S2O8−2) is a strong oxidant (E0 = 2.05 V) used to degrade pollutants such as SMZ (Ji et al., 2018). Table 6 shows that the SMZ degradation rate increase with increasing PDS concentration in solution. Apparently, UVC/PDS appeared to be more efficient than UVC/ H2O2 for SMZ degradation attributed to the quantum yield of 1.4 and 1.0 to SO4%- and H2O2, respectively (Ji et al., 2018). Based on the S2O8−2 reactions are very slow at room temperature, different methods have been proposed to activate organic molecule degradation, including the photochemical decomposition of S2O82− to generate sulfate radicals (SO4%-) (Reaction (14)) that can degrade SMZ by participating in highly oxidizing radical reactions (Reactions (14)–(17)). (14) (15) (16) (17)
Fig. 5. Influence of PDS as promoter of SO4%- radicals in the degradation of SMZ at different H2O2 concentrations. [SMZ]0 = 15 mg L−1, pH = 7, T = 298 K.
Recombination radical reactions (Reactions (18)–(23)) are available to compete in SMZ degradation. (18)
concentrations; initial H2O2 concentrations of 1.47 × 10−4 M and 4.4 × 10−4 M produce the maximum radical generation of 6.68 × 10−13 M and 1.54 × 10−12 M at 10 and 8 min, respectively. The generation of radicals subsequently decreases, although the SMZ degradation continues, attributable to direct photolysis by UVC radiation. On the other hand, at an initial SMZ concentration of 4.4 × 10−4 M PDS, a maximum of 3.4 × 10−5 M SO4%- radicals are generated (Fig. 5). Thus, in the UV/PDS system, the SMZ exposure to radicals is higher than system UV/H2O2.
(19) (20) (21) (22)
3.5. Economic analysis by different oxidation process
(23) The rate constant value of the SO4%- radical with SMZ was determined as kSO4%- = 7.88 × 108 M−1s−1 and in comparison, to the Reactions (18)–(23), all recombination radical reactions are sufficient to compete with the SMZ degradation reaction in the presence of this radical.
To demonstrate the possible application of the UVC, UVC/H2O2 and UVC/PDS for SAs degradation, we applied an economic analysis based in the EE/O concept (Yin et al., 2018). EE/O include the electrical energy of UVC lamp (EE/OUVC) and the consumption of the oxidants (Oxidant/O). Which can be calculated by the following equation (24):
EE / OTotal = EE / OUVC + EE /O =
3.4.3. Quantification of radicals generated Other way to define the efficiency of the different oxidation process is compare the quantification of radicals generated (Oh et al., 2017). Fig. 4 depicts the concentration of OH% radicals at different H2O2
P·t c + oxidant V ·log(ci/ ct ) log(ci/ ct )
(24)
−1
Where P is the energy input of UVC lamp (k Wh h ), t the irradiation time (h), V the reactor volume (L), ci and ct are the initial and the final concentration of SA, respectively and coxidant is the concentration of H2O2 or PDS. Therefore, the total cost of SMZ degradation was 1.235, 0.257, 0.181 kWhL−1 to UVC, UVC/H2O2 and UVC/PDS, respectively. SMZ is degraded mainly by reactive species oxidation, addition oxidant helped lower EE/O (Zhang et al., 2016a). UVC/PDS process cost less energy than UVC/H2O2 process because SMZ reacted rapidly with SO4−. 3.6. Degradation by-products By-products of SMZ degradation by different systems. Chromatography and mass spectrum studies identified by-products, listed in Table 7 alongside their molecular formula, molecular weight, and chemical structure. Most of these by-products were previously reported. Based on the identified by-products, the degradation pathway of SMZ is proposed in Fig. 6, showing that SO2 removal is first produced by direct SMZ photolysis through UVC radiation, obtaining by-product P1. Wan et al. (Wan and Wang, 2016) indicated that sulfate ions are the main intermediate by-products. Next, given the oxidation of OH% radical during the process, hydroxylation is expected to be a common reaction responsible for SMZ degradation, generating by-products P3-1,
Fig. 4. Influence of H2O2 as promoter of OH% radicals in the degradation of SMZ at different H2O2 concentrations. [SMZ]0 = 15 mg L−1, pH = 7, T = 298 K. 228
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Table 7 By-products of the direct photolysis of SMZ by UV radiation. Treatment System
By-product identified
Molecular formula
Molecular weight (g mol−1)
UVC UVC/ H2O2 UVC/PDS UVC UVC/ H2O2 UVC/PDS
P1-1
C12H14N4
214.25
(Batista et al., 2014)
P1-2
C6H10O2N4S
214.23
(Wan and Wang, 2017)
UVC UVC/PDS
P2
C6H9N3
123.14
(Batista et al., 2014; Liu and Wang, 2013)
UVC UVC/ H2O2 UVC/PDS
P3-1
C12H14N4O
230.25
(Dong et al., 2017)
UVC UVC/ H2O2
P3-2
C12H14N4O
230.25
(Dong et al., 2017)
UVC/PDS
P4
C12H12N4O2
244.24
–
UVC/PDS UVC/ H2O2
P5-1
C12H12N4O2
294.32
(Wan and Wang, 2017)
UVC/H2O2
P5-2
C12H14N4O3S
294.32
(Wan and Wang, 2017)
UVC/H2O2 UVC/PDS
P6
C12H12N4O3S
292.30.
(Yin et al., 2018)
UVC/H2O2 UVC/PDS
P7
C12H12N4O4S
308.30
(Yin et al., 2018)
Chemical structure
Ref.
those of normal human cells (Kadu et al., 2017). Cell viability < 75% is considered a toxic effect and cell viability > 75% a viable effect. Fig. 7 depicts the cytotoxicity results of the degradation by-products as a function of time. The viability of cell cultures was > 75% for all sulfonamides and was even higher for SDZ and SML than for the control. Comparison of by-products with the original products (time 0) shows a higher viability for the by-products, which are therefore considered less toxic than the original product. Similar behavior has founded to other of pharmaceutical products (Yaghmaeian et al., 2017).
P3-2, P5-1 or P5-2. Sági et al. (2015) reported that the basic initial reaction is the addition of the hydroxyl radical to the benzene ring or, in some cases, heterocyclic rings, forming radical intermediates of cyclohexadiene accompanied by a fall in pH due to the formation of SO42− and smaller organic acids. A break in the eSO2e and eNHe bond allows identification of by-product P2, of low molecular weight, as previously reported in the mineralization pathway (Dong et al., 2017). By attack nucleophilic and incision in the group eSO2e, the byproduct P4 was formed. The hydroxylation intermediates are the major products formed by sulfate radical attack on aromatics (Yuan et al., 2011). Adduction of the 7 N atoms, forming the nitroso- and nitro-substitutional SMZ through electrophilic reaction was the dominated pathways by PDS (P6 and P7) (Yin et al., 2018).
4. Conclusions In this analysis of the degradation of three sulfonamides (SMZ, SDZ, and SML) based on quantification of the UVC radiation, found low values of quantum yields at 1 h of treatment. The UVC dose widely used for water disinfection in treatment plants is inadequate to remove this type of antibiotic. Direct UVC photolysis of SAs is influenced by their
3.7. Cytotoxicity of degradation by-products The metabolic conditions of HEK-293 cells are highly similar to 229
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Fig. 6. Mechanism of degradation by direct SMZ photolysis in the presence of UVC radiation.
Fig. 7. HEK-293% Cell viability in SA samples after different periods of UVC radiation. [SAs]0 = 5.38 × 10−6 M.
initial concentration, the degradation rates were higher at pH 12 (anionic specie) for SMZ and SML. The presence of sulfates, nitrates, and chlorides in tap water can exert a synergic effect on the photodegradation, especially in the cases of SMZ and SML. A concentration of 4.4 × 10−4 M of H2O2 or PDS increases SMZ degradation up to 100%. UVC/PDS cost less energy than UVC and UVC/H2O2. Hydroxylation reactions, binding between eSO2e and eNHe groups and electrophilic attacks were the aim way to degraded SAs. Ten by-products of UV radiation-induced SMZ degradation were identified and present a less cytotoxic in comparison to the original products.
sulfamethazine degradation using UV-based technologies and products identification. J. Photochem. Photobiol. A Chem. 290, 77–85. Białk-Bielińska, A., Stolte, S., Arning, J., Uebers, U., Böschen, A., Stepnowski, P., Matzke, M., 2011. Ecotoxicity evaluation of selected sulfonamides. Chemosphere 85, 928–933. Boreen, A.L., Arnold, W.A., McNeill, K., 2004. Photochemical fate of sulfa drugs in the aquatic environment: sulfa drugs containing five-membered heterocyclic groups. Environ. Sci. Technol. 38, 3933–3940. Burgos-Castillo, R.C., Sirés, I., Sillanpää, M., Brillas, E., 2018. Application of electrochemical advanced oxidation to bisphenol A degradation in water. Effect of sulfate and chloride ions. Chemosphere 194, 812–820. Canonica, S., Jans, U., Stemmler, K., Hoigne, J., 1995. Transformation kinetics of phenols in water: photosensitization by dissolved natural organic material and aromatic ketones. Environ. Sci. Technol. 29, 1822–1831. Cizmas, L., Sharma, V.K., Gray, C.M., McDonald, T.J., 2015. Pharmaceuticals and personal care products in waters: occurrence, toxicity, and risk. Environ. Chem. Lett. 13, 381–394. Cui, C., Jin, L., Han, Q., Lin, K., Lu, S., Zhang, D., Cao, G., 2016. Removal of trace level amounts of twelve sulfonamides from drinking water by UV-activated peroxymonosulfate. Sci. Total Environ. 572, 244–251. Dong, H., Qiang, Z., Lian, J., Qu, J., 2017. Degradation of nitro-based pharmaceuticals by UV photolysis: kinetics and simultaneous reduction on halonitromethanes formation potential. Water Res. 119, 83–90. Fan, Y., Ji, Y., Kong, D., Lu, J., Zhou, Q., 2015. Kinetic and mechanistic investigations of the degradation of sulfamethazine in heat-activated persulfate oxidation process. J. Hazard Mater. 300, 39–47. Iglesias, A., Nebot, C., Miranda, J.M., Vázquez, B.I., Cepeda, A., 2012. Detection and quantitative analysis of 21 veterinary drugs in river water using high-pressure liquid chromatography coupled to tandem mass spectrometry. Environ. Sci. Pollut. Control Ser. 19, 3235–3249. Ji, Y., Yang, Y., Zhou, L., Wang, L., Lu, J., Ferronato, C., Chovelon, J.-M., 2018. Photodegradation of sulfasalazine and its human metabolites in water by UV and UV/ peroxydisulfate processes. Water Res. 133, 299–309. Kadu, B.S., Wani, K.D., Kaul-Ghanekar, R., Chikate, R.C., 2017. Degradation of doxorubicin to non-toxic metabolites using Fe-Ni bimetallic nanoparticles. Chem. Eng. J. 325, 715–724.
Acknowledgments The authors are grateful for the financial support of the Ministry of Science and Innovation (CTQ2016-80978-C2-1-R). A. Acosta-Rangel is thankful to CONACYT for their financial aid granted of through the scholarship (Beca Mixta) No. 291062. References Babić, S., Zrnčić, M., Ljubas, D., Ćurković, L., Škorić, I., 2015. Photolytic and thin TiO2 film assisted photocatalytic degradation of sulfamethazine in aqueous solution. Environ. Sci. Pollut. Res. 22, 11372–11386. Baeza, C., Knappe, D.R., 2011. Transformation kinetics of biochemically active compounds in low-pressure UV photolysis and UV/H 2 O 2 advanced oxidation processes. Water Res. 45, 4531–4543. Baran, W., Sochacka, J., Wardas, W., 2006. Toxicity and biodegradability of sulfonamides and products of their photocatalytic degradation in aqueous solutions. Chemosphere 65, 1295–1299. Batista, A.P.S., Pires, F.C.C., Teixeira, A.C.S., 2014. The role of reactive oxygen species in
230
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
N-nitrosodimethylamine in simulated drinking water. Environ. Sci. Technol. 37, 1933–1940. Souza, B.M., Souza, B.S., Guimarães, T.M., Ribeiro, T.F., Cerqueira, A.C., Sant'Anna, G.L., Dezotti, M., 2016. Removal of recalcitrant organic matter content in wastewater by means of AOPs aiming industrial water reuse. Environ. Sci. Pollut. Control Ser. 23, 22947–22956. Sun, L., Bolton, J.R., 1996. Determination of the quantum yield for the photochemical generation of hydroxyl radicals in TiO2 suspensions. J. Phys. Chem. 100, 4127–4134. Tan, C., Gao, N., Zhou, S., Xiao, Y., Zhuang, Z., 2014. Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes. Chem. Eng. J. 253, 229–236. Velo-Gala, I., Pirán-Montaño, J., Rivera-Utrilla, J., Sánchez-Polo, M., Mota, A.J., 2017. Advanced Oxidation Processes based on the use of UVC and simulated solar radiation to remove the antibiotic tinidazole from water. Chem. Eng. J. 323, 605–617. Wan, Z., Wang, J.-L., 2016. Removal of sulfonamide antibiotics from wastewater by gamma irradiation in presence of iron ions. Nucl. Sci. Tech. 27, 1–5. Wan, Z., Wang, J., 2017. Fenton-like degradation of sulfamethazine using Fe3O4/Mn3O4 nanocomposite catalyst: kinetics and catalytic mechanism. Environ. Sci. Pollut. Control Ser. 24, 568–577. Wang, J., Wang, S., 2016. Removal of pharmaceuticals and personal care products (PPCPs) from wastewater: a review. J. Environ. Manag. 182, 620–640. Yaghmaeian, K., Moussavi, G., Mashayekh-Salehi, A., Mohseni-Bandpei, A., Satari, M., 2017. Oxidation of acetaminophen in the ozonation process catalyzed with modified MgO nanoparticles: effect of operational variables and cytotoxicity assessment. Process Saf. Environ. Protect. 109, 520–528. Yang, Y., Lu, X., Jiang, J., Ma, J., Liu, G., Cao, Y., Liu, W., Li, J., Pang, S., Kong, X., 2017. Degradation of sulfamethoxazole by UV, UV/H 2 O 2 and UV/persulfate (PDS): formation of oxidation products and effect of bicarbonate. Water Res. 118, 196–207. Yin, R., Guo, W., Wang, H., Du, J., Zhou, X., Wu, Q., Zheng, H., Chang, J., Ren, N., 2018. Enhanced peroxymonosulfate activation for sulfamethazine degradation by ultrasound irradiation: performances and mechanisms. Chem. Eng. J. 335, 145–153. Yuan, R., Ramjaun, S.N., Wang, Z., Liu, J., 2011. Effects of chloride ion on degradation of Acid Orange 7 by sulfate radical-based advanced oxidation process: implications for formation of chlorinated aromatic compounds. J. Hazard Mater. 196, 173–179. Zhang, H., Wei, X., Song, X., Shah, S., Chen, J., Liu, J., Hao, C., Chen, Z., 2018. Photophysical and photochemical insights into the photodegradation of sulfapyridine in water: a joint experimental and theoretical study. Chemosphere 191, 1021–1027. Zhang, R., Yang, Y., Huang, C.-H., Zhao, L., Sun, P., 2016a. Kinetics and modeling of sulfonamide antibiotic degradation in wastewater and human urine by UV/H2O2 and UV/PDS. Water Res. 103, 283–292. Zhang, R., Yang, Y., Huang, C.-H., Zhao, L., Sun, P., 2016b. Kinetics and modeling of sulfonamide antibiotic degradation in wastewater and human urine by UV/H 2 O 2 and UV/PDS. Water Res. 103, 283–292. Zhao, T., Li, P., Tai, C., She, J., Yin, Y., Qi, Y.a., Zhang, G., 2018. Efficient decolorization of typical azo dyes using low-frequency ultrasound in presence of carbonate and hydrogen peroxide. J. Hazard Mater. 346, 42–51. Zhu, W., Liu, J., Yu, S., Zhou, Y., Yan, X., 2016. Ag loaded WO 3 nanoplates for efficient photocatalytic degradation of sulfanilamide and their bactericidal effect under visible light irradiation. J. Hazard Mater. 318, 407–416. Zhu, Y., Wu, M., Gao, N., Chu, W., Li, K., Chen, S., 2018. Degradation of phenacetin by the UV/chlorine advanced oxidation process: kinetics, pathways, and toxicity evaluation. Chem. Eng. J. 335, 520–529.
Kelner, M., Bagnell, R., Welch, K., 1990. Thioureas react with superoxide radicals to yield a sulfhydryl compound. Explanation for protective effect against paraquat. J. Biol. Chem. 265, 1306–1311. Kurniawan, T.A., Lo, W.-h., Chan, G., 2006. Radicals-catalyzed oxidation reactions for degradation of recalcitrant compounds from landfill leachate. Chem. Eng. J. 125, 35–57. Li, M., Wang, C., Yau, M., Bolton, J.R., Qiang, Z., 2017. Sulfamethazine degradation in water by the VUV/UV process: kinetics, mechanism and antibacterial activity determination based on a mini-fluidic VUV/UV photoreaction system. Water Res. 108, 348–355. Lian, J., Qiang, Z., Li, M., Bolton, J.R., Qu, J., 2015. UV photolysis kinetics of sulfonamides in aqueous solution based on optimized fluence quantification. Water Res. 75, 43–50. Liu, Y., Wang, J., 2013. Degradation of sulfamethazine by gamma irradiation in the presence of hydrogen peroxide. J. Hazard Mater. 250, 99–105. Luo, S., Wei, Z., Spinney, R., Zhang, Z., Dionysiou, D.D., Gao, L., Chai, L., Wang, D., Xiao, R., 2018. UV direct photolysis of sulfamethoxazole and ibuprofen: an experimental and modelling study. J. Hazard Mater. 343, 132–139. Magureanu, M., Mandache, N.B., Parvulescu, V.I., 2015. Degradation of pharmaceutical compounds in water by non-thermal plasma treatment. Water Res. 81, 124–136. Ocampo-Pérez, R., Sánchez-Polo, M., Rivera-Utrilla, J., Leyva-Ramos, R., 2010. Degradation of antineoplastic cytarabine in aqueous phase by advanced oxidation processes based on ultraviolet radiation. Chem. Eng. J. 165, 581–588. Oh, W.-D., Dong, Z., Ronn, G., Lim, T.-T., 2017. Surface–active bismuth ferrite as superior peroxymonosulfate activator for aqueous sulfamethoxazole removal: performance, mechanism and quantification of sulfate radical. J. Hazard Mater. 325, 71–81. Pomati, F., Castiglioni, S., Zuccato, E., Fanelli, R., Vigetti, D., Rossetti, C., Calamari, D., 2006. Effects of a complex mixture of therapeutic drugs at environmental levels on human embryonic cells. Environ. Sci. Technol. 40, 2442–2447. Pomati, F., Orlandi, C., Clerici, M., Luciani, F., Zuccato, E., 2007. Effects and interactions in an environmentally relevant mixture of pharmaceuticals. Toxicol. Sci. 102, 129–137. Prados-Joya, G., Sánchez-Polo, M., Rivera-Utrilla, J., Ferro-Garcia, M., 2011. Photodegradation of the antibiotics nitroimidazoles in aqueous solution by ultraviolet radiation. Water Res. 45, 393–403. Rahmani, H., Gholami, M., Mahvi, A., Alimohammadi, M., Azarian, G., Esrafili, A., Rahmani, K., Farzadkia, M., 2014. Tinidazole removal from aqueous solution by sonolysis in the presence of hydrogen peroxide. Bull. Environ. Contam. Toxicol. 92, 341–346. Reina, A.C., Martínez-Piernas, A.B., Bertakis, Y., Brebou, C., Xekoukoulotakis, N.P., Agüera, A., Pérez, J.A.S., 2018. Photochemical degradation of the carbapenem antibiotics imipenem and meropenem in aqueous solutions under solar radiation. Water Res. 128, 61–70. Sági, G., Bezsenyi, A., Kovács, K., Klátyik, S., Darvas, B., Székács, A., Mohácsi-Farkas, C., Takács, E., Wojnárovits, L., 2018. Radiolysis of sulfonamide antibiotics in aqueous solution: degradation efficiency and assessment of antibacterial activity, toxicity and biodegradability of products. Sci. Total Environ. 622, 1009–1015. Sági, G., Csay, T., Szabó, L., Pátzay, G., Csonka, E., Takács, E., Wojnárovits, L., 2015. Analytical approaches to the OH radical induced degradation of sulfonamide antibiotics in dilute aqueous solutions. J. Pharmaceut. Biomed. Anal. 106, 52–60. Sharpless, C.M., Linden, K.G., 2003. Experimental and model comparisons of low-and medium-pressure Hg lamps for the direct and H2O2 assisted UV photodegradation of
231
Contents lists available at ScienceDirect
Journal of Environmental Management journal homepage: www.elsevier.com/locate/jenvman
Research article
Sulfonamides degradation assisted by UV, UV/H2O2 and UV/K2S2O8: Efficiency, mechanism and byproducts cytotoxicity
T
A. Acosta-Rangela,b,∗, M. Sánchez-Poloa, A.M.S. Poloa, J. Rivera-Utrillaa, M.S. Berber-Mendozab a
Department of Inorganic Chemistry, Faculty of Science, University of Granada, 18071 Granada, Spain Center of Postgraduate Research and Studies, Faculty of Engineering, University Autonomous of San Luis Potosí, Av. Dr. M. Nava No. 8, San Luis Potosí, S.L.P., 78290, Mexico b
A R T I C LE I N FO
A B S T R A C T
Keywords: Sulfonamides Intermediates Ultraviolet radiation Cytotoxicity
The objective of this study was to analyze the effectiveness of UVC, UVC/H2O2 and UVC/K2S2O8 on the degradation of SAs. Rate constant values increased in the order SMZ < SDZ < SML and showed the higher photodegradation of sulfonamides with a penta-heterocycle. Quantum yields were 1.72 × 10−5 mol E−1, 3.02 × 10−5 mol E−1, and 6.32 × 10−5 mol E−1 for SMZ, SDZ and SML, respectively, at 60 min of treatment. R254 values show that the dose habitually utilized for water disinfection is inadequate to remove this type of antibiotic. The initial sulfonamide concentration has a major impact on the degradation rate. The degradation rates were higher at pH 12 for SMZ and SML. SMZ and SML photodegradation kλ values are higher in tap versus distilled water. The presence of radical promoters generates a greater increase in the degradation rate, UVC/ K2S2O8 cost less energy, a mechanism was proposed, and the degradation by-products are less toxic than the original product.
1. Introduction
determination parameter quantic (direct photolysis) or compare the efficacy of the methods (indirect photolysis). These methods are characterized by generation of oxidizing species such as OH% or SO4%- radicals, which have a redox potential of 1.8–2.7 V and 2.5–3.1 V, respectively, with the latter being a more selective oxidant (Yang et al., 2017), have been successfully applied to destruct organic micropollutants in different water matrices (Souza et al., 2016). The major of studies using the Vibrio fischeri bacteria test to determine the toxicity of SAs (Białk-Bielińska et al., 2011; Sági et al., 2018). Knowledge of their byproducts cytotoxicity is still basic and restricted to just a few SAs (Cizmas et al., 2015). HEK 293 cells lack the expression of hormone receptors an important characteristic for studying basal toxicity associated with drugs (Pomati et al., 2006, 2007). The studies that using UV photolysis and its combination with PDS to eliminated sulfonamides still deficient in terms of mechanism, by-products, toxicity and quantification of radicals generated. Therefore, in this study, we evaluate the effectiveness of UVC photolysis, UVC/H2O2 and UVC/PDS in the oxidation of three SAs; sulfamethazine (SMZ), sulfadiazine (SDZ) and sulfamethizole (SML). The objectives of our work were to: i) A kinetic study was conducted to determine the quantum yield (Φλ ) of the UVC photolysis, ii) The influence of different operational variables (initial SAs concentration, oxidant dosage, solution pH, and water matrix) was
Ultraviolet radiation (UVC-254nm) is frequently applied to disinfect water intended for human consumption and wastewaters because it is effective to remove organic pollutants due to occupies a small space, and is readily managed and maintained (Lian et al., 2015; OcampoPérez et al., 2010; Prados-Joya et al., 2011). Sulfonamides (SAs) are widely used as antibiotics in human and veterinary medicine due to their strong antimicrobial activity, stable chemical properties, and low cost (Magureanu et al., 2015; Wang and Wang, 2016). SAs have been detected in surface waters at concentrations of 148–2978 ng L−1 (Iglesias et al., 2012). Prolonged exposure to low concentrations of antibiotics that accumulate in waters can be toxic for cells or lead to resistance in bacterial strains, posing a major public health problem (Baran et al., 2006). Previous studies have found that the direct photolysis based in the fluence quantification can be photodegraded only part of SAs during the UVC disinfection of water, SAs with penta-heterocycle are degraded faster (Li et al., 2017; Yang et al., 2017). By UVCactivated hydrogen peroxide (H2O2), peroximonosulfate (PMS) and peroxydisulfate (PDS) the SAs oxidation have been studied (Babić et al., 2015; Cui et al., 2016; Zhang et al., 2016b; Zhu et al., 2016). However, these efforts have focused in developed and optimized methods to
∗ Corresponding author. Center of Postgraduate Research and Studies, Faculty of Engineering, University Autonomous of San Luis Potosí, Av. Dr. M. Nava No. 8, San Luis Potosí, S.L.P., 78290, Mexico. E-mail address: [email protected] (A. Acosta-Rangel).
https://doi.org/10.1016/j.jenvman.2018.06.097 Received 2 November 2017; Received in revised form 15 February 2018; Accepted 30 June 2018 0301-4797/ © 2018 Elsevier Ltd. All rights reserved.
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Table 1 Molecular structure and physicochemical properties of the studied SAs. Molecular weight (g mol−1)
LogKow
Solubility (mg/L)
pKa
Sulfamethazine C12H14N4O2S (SMN)
278.33
0.19
1500 (29 °C)
pKa1: 2.00 pKa2: 6.99
Sulfadiazine C10H10N4O2S (SDZ)
250.278
−0.09
77 (25 °C)
pKa1: 2.01 pKa2: 6.99
Sulfamethizole C9H10N4O2S2 (SML)
270.33
0.54
1050 (37 °C)
pKa1: 1.95 pKa2: 6.71
Sulfonamide (SA)
Molecular structure
evaluated, iii) The quantification of OH% and SO4%- radicals generated was determined, iv) Degradation by-products were identified, and v) by-product cytotoxicity in cell HEK-293 was measured.
(2) where kob is the mean value of the slopes,
2. Material and methods
, are the reaction rate constants, and are the SMZ reaction rate constants with each radical. The photodegradation rate constant was determined by applying a pseudo-first-order model and considering 1 h as reaction follow-up time, the quantum yield (Φλ ) of each SA, the apparent photodegradation rate constant normalized by the lamp energy (k ′E ) (constant independent of fluctuations), the molar absorption capacity (ελ ) and R254 parameter representing the elimination percentages of SAs in water treatment plants to 400 J m−2, the radiant energy emitted by the lamp was 8.49 × 10-4 E m−2 to 254 nm determined by actinometry (Canonica et al., 1995; Prados-Joya et al., 2011; Sharpless and Linden, 2003).
2.1. Chemical substances All chemical reagents used were high-purity analytical grade reagents supplied by Sigma-Aldrich. All solutions were prepared with ultrapure water obtained using the Milli-Q® system (Millipore). Table 1 lists the molecular structures and physicochemical properties of the sulfonamides under study. 2.2. SAs degradation by UVC-254nm The experimental system for SAs degradation kinetics was in a photoreactor equipped with a low-pressure (254 nm Hg) TNN 15/32 Heraeus Noblelight mercury lamp (nominal power 15 W). Solutions were deposited in six quartz tubes 1 cm in diameter with 35 mL capacity, which were placed in parallel and equidistant to the Hg lamp. The tubes were immersed in recirculating distilled water to maintain a constant temperature of 25 °C, using a Frigiterm ultra-thermostat and magnetic agitation system in each tube. SAs initial concentration was 15 mg L−1, samples were drawn from each solution at regular time intervals for subsequent measurement of the SAs concentration. The influence of experimental parameters such as initial pH (1.5, 6.5 and 12), initial SMZ concentration (5, 10 and 15 mg L−1), oxidant dosage (1.4 × 10−4 and 4.4 × 10−4 M) and water matrix (distiller and tap) was investigated.
2.4. Quantification of OH% and SO4%- radical concentration The concentration of OH% radical generated was calculated using PCBA. PCBA degradation with OH% radicals were quantified in the UV radiation system, in which radicals are responsible for the oxidation. Total concentrations of OH% radical at different time points were calculated the same way as other authors (Velo-Gala et al., 2017). The concentration of SO4%- radicals generated were quantified by identifying the reaction by-product, benzoquinone (BQ) (Oh et al., 2017), obtained through oxidation of p-hydroxybenzoic acid (HBA) by the SO4•- radicals generated in the system under UV radiation. Based on reaction stoichiometry (equation (3)), 1 mol HBA reacts with 1 mol SO4%- to form hydroquinone, which immediately transforms into a stable by-product, BQ, due to excess PDS. The excess PDS is added outside the UV radiation system.
2.3. Determination of the rate constant of SMZ with OH% and SO4%- radical The rate constant of SMZ with OH% and SO4%- radical was determined in competitive kinetics experiments, using para-chlorobenzoic acid (PCBA) as reference compound for OH% (kOH%/ 9 −1 −1 S ) (Kelner et al., 1990) at molar ratio of 1:1 with pCBA=5.2 × 10 M excess H2O2, and using citarabine (CTB) as reference compound for SO4%- ( ) (Ocampo-Pérez et al., 2010) at molar ratio of 1:1 with excess K2S2O8. When ln against ln
(
[CTB ó pCBA] [CTB ó pCBA]0
)
(
[SMZ ] [SMZ ]0
(3) In general, BQ concentrations are proportional to SO4%- concentrations. The BQ generated is relatively stable up to pH 9.0, when excess CHBA˃CBQ, because HBA inhibits the reaction between BQ and SO4%-, whose constants are .
) is plotted
2.5. Analytical methods
, depending on the case, the slope of the graph 2.5.1. Determination of SAs, PCBA, CTB, and BQ concentration in aqueous solution Concentration of the SAs was determined by reverse-phase highperformance liquid chromatography (HPLC), using a liquid chromatographer (Thermo-Fisher) equipped with visible UV-detector and
(kob ) allows the reaction constants of kOH•/ SMZ and kSO4• −/ SMZ radicals to be calculated according to Equations (1) and (2): (1)
225
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
automatic autosampler with capacity for 120 vials. A PHENOMENEX Kinetex C-18 column was used (2.6 μm particle size; 4.6 × 150 mm). The mobile phase was 70% formic acid (0.1%, v/v) and 30% acetonitrile (0.1% v/v) in isocratic mode at a flow of 0.35 mL min−1; the detector wavelength was set at 270 nm. The same methodology was used to determine PCBA and CTB concentrations except that a wavelength of 240 nm with 97% formic acid (0.4%, v/v) was used for PCBA and a wavelength of 272 nm with 3% methanol for CTB. BQ was determined at a wavelength of 244 nm using 50% methanol and 50% deionized water. 2.5.2. Identification of degradation by-products Degradation by-products of SMZ were identified by using Acquity ultra performance liquid chromatography (UPLC) system (Waters) equipped with a CORTECS™ C18 column (2.1 × 75 mm, 2.7 μm) (Waters). The mobile phase, in gradient mode (baseline: 0% B, T8: 95% B, T8.1: 0% B), was channel A, water with 0.1% formic acid and channel B, acetonitrile with 0.1% formic acid, at flow of 0.4 mL min-1, injection volume of 10 μL, and column temperature of 40 °C.
Fig. 1. Degradation kinetics of the three SAs by UVC photolysis. [SAs]0 = 15 mg L−1, pH = 7 and T = 298 K. The lines represent the prediction of the first-order kinetic model.
exposure times are required for this purpose in water treatment plants. 2.5.3. Evaluation of by-products cytotoxicity By-products cytotoxicity were evaluated by using an MTS assay to determine the % viability of human embryo kidney cells (HEK-293) from the CIC cell bank of the University of Granada. Degradation kinetics were first studied in the presence of phosphate buffered saline (PBS), and 10,000 HEK-293 cells were then incubated for 24 h, subsequently changing the medium and adding SAs by-products (10:100 μL dilution); after incubation for a further 24 h, 20 μL MTS was added and its absorbance was measured at 2 h using INFINITENANOQUA equipment, with a sample reading of 9 times at 490 nm absorbance.
3.2. Influence of the different operational variables on UVC photolysis of SAs 3.2.1. Influence of the initial SAs concentration SMZ was selected as model SAs due to show the lowest degradation rate. Fig. 2 and Table 3 shows that the Φλ (1.72 × 10−5, 2.24 × 10−5 and 3.75 × 10−5 mol E−1) and therefore kλ (4.10 × 10−4, 5.95 × 10−4 and 9.16 × 10−4 s−1) is influenced by the initial concentration (15,10 and 5 mg L−1, respectively). The reduced photodegradation rate at higher concentrations is related to the energy absorbed by each SMZ molecule (Wan and Wang, 2017). A fixed energy per volume unit is deposited in the medium, and molecules can accept more radiant energy at lower concentrations because there are fewer molecules in the medium and their availability (Rahmani et al., 2014).
3. Results and discussion 3.1. Direct UVC photolysis of SAs: quantum yields The quantum yield (Φλ ) is a critical factor to quantify the photon efficiency of photolysis reactions (Zhang et al., 2018). Table 2 shows the values of the study parameters. The Φλ for SMZ, SDZ and SML degradation by UVC photolysis was found to be 1.72 × 10−5,3.02 × 10−5 and 6.02 × 10−5 mol E−1, respectively. Which is low values than reported by other authors to degradation of SAs (Lian et al., 2015; Luo et al., 2018) and similar to other (Baeza and Knappe, 2011), due to energy irradiated by the lamp (1.027 × 10−4 E s−1 m−1 for this lamp). Fig. 1 shows the degradation kinetics constant, the trend of decreasing photolysis rate as SML > SDZ > SML. SML was degraded rapidly (1.31 × 10−5 s−1), this is evident due to include a penta-heterocycle attributed to differences in the electron densities of the heterocyclic rings (Cui et al., 2016). Molar absorption coefficient (ελ ) values of SMZ, SDZ and SML were determined to be 1.68 × 103, 1.34 × 103 and 1.46 × 103 m2 mol−1, respectively. These values suggest that SAs may undergo direct photolysis to 254 nm (Ji et al., 2018). R254 is a parameter to determines the applicability UV radiation under real condition (Prados-Joya et al., 2011). The low R254 values (Table 2) obtained to SAs indicate that the dose (400 J m−2) commonly used for water disinfection is inadequate to remove SAs, demonstrating that higher UV radiation doses or longer
3.2.2. Influence of solution pH The solution pH values selected for study were 1.5, 6.5, and 12, at which SAs are in cationic (is protonated at its amine group), neutral, and anionic (deprotonated in NH group) form, respectively. Fig. 3 depicts the variation in global molar absorption coefficient (ελ ) and global photodegradation rate constant (k ′E ) as a function of solution pH. The values of ελ increase in order with the pH, being higher when the SAs are in their anionic form (Fan et al., 2015). The degradation rates were higher at pH 12 for the SMZ and SML, consistent with major value ελ .
Table 2 Parameters obtained from the direct UVC photolysis of the three SAs at 254 nm. Sulfonamide
ελ 10−3 (m2 mol−1)
kλ 104 (s−1)
Φλ 105 (mol E−1)
kE′ 102 (m2 E−1)
R254 103 (%)
SMZ SDZ SML
1.681 1.342 1.464
4.10 5.74 13.13
1.72 3.02 6.32
6.67 9.33 21.32
5.62 7.05 18.07
Fig. 2. Influence of initial SMZ concentration on photodegradation with UVC radiation. pH = 7 and T = 298 K. The lines represent the prediction of the firstorder kinetic model. 226
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Table 3 Parameters obtained with direct UVC radiation of SMZ at 254 nm.
Table 5 Results obtained by UVC photolysis of SMZ at 254 nm in distilled and tap water.
[SMZ]0
ελ 10−3 (m2 mol−1)
kλ 104 (s−1)
Φλ 105 (mol E−1)
kE′ 102 (m2 E−1)
R254 103 (%)
15 10 5
1.681 1.646 1.723
4.10 5.95 9.16
1.72 2.24 3.75
6.67 8.51 14.90
5.62 7.21 12.61
Sulfonamide
Water
pH
Φλ 105 (mol E−1)
kE′ 102 (m2 E−1)
R254 103 (%)
% SAs degradation (60 min)
SMZ
Distilled Tap Distilled Tap Distilled Tap
6.7 8.2 6.2 8.0 5.7 8.2
1.72 3.54 3.01 1.70 3.32 16.11
6.63 17.90 8.32 7.87 21.32 63.50
5.62 15.20 7.05 6.67 18.07 53.90
77.10 98.05 87.41 82.08 99.12 100
SDZ SML
3.4. Indirect UVC photolysis of SAs 3.4.1. SMZ photodegradation in presence of different dose of H2O2 The combined use of UV/H2O2 is known to break the H2O2 molecule, producing two OH% radicals for each molecule and each photon absorbed (Reaction (4)) due to the high quantum yield ( Φ= 0.98 mol E −1) (Velo-Gala et al., 2017). (4) The effect of the H2O2 on the oxidative degradation of SMZ is presented in Table 6. The percentage SMZ degradation after 60 min of treatment at higher initial H2O2 concentrations, rising from 77% in the absence of H2O2 to 100% at an initial H2O2 concentration of 4.4 × 10−4 M. The presence of H2O2 enhances degradation by forming highly oxidizing radical species (Reactions (5)–(10)), which react with SMZ and therefore favor a higher degradation rate. (5)
Fig. 3. Molar absorption coefficient (ελ ) and normalized apparent photodegradation constant of energy emitted by the lamp (kE' ) as a function of the initial solution pH. [SAs]0 = 15 mg L-1, T = 298 K.
Protonated SAs exhibit low reactivity toward electrophilic radicals but deprotonated are easily oxidized by reactive radicals (Cui et al., 2016). However, for SDZ the degradation rate decrease with increase the pH and ελ similar behavior was founded by other authors (Boreen et al., 2004).
(6) (7) (8) (9) RH + HO%→by – products degradation
3.3. Applicability of UVC photolysis of SAs in different water matrix
They can also react with other species besides SMZ (Reactions (11)–(13)) (Kurniawan et al., 2006; Sun and Bolton, 1996), but these reactions reduce the OH% species in the medium, diminishing the efficacy of the system.
The applicability of UVC radiation to remove SAs in aqueous solution was determined by identifying the influence of the chemical composition (Table 4) of tap water and distilled water on their photodegradation. HCO3− is considered to be a scavenger of free radicals (Tan et al., 2014), nevertheless is reacted with electron-rich species such as anilines (Zhao et al., 2018). Several studies mentioned that SO42− (Burgos-Castillo et al., 2018), NO3− (Reina et al., 2018) or halogens (Zhu et al., 2018) can contribute to pharmaceutical compounds degradation. The results are exhibited in Table 5, revealing a higher degradation rate of SMZ (98%) and SML (100%) in tap versus distilled water. SDZ show different behavior similar to influence of pH. Tap water has a lower transmittance value, favoring UVC radiation absorption and markedly reducing the number of photons reaching the SAs (Prados-Joya et al., 2011). In conclusion, the presence of nitrates, sulfates, and chlorates in tap water exert a synergic effect on SAs degradation.
(11) (12) OH% + OH−→ O%– + H2O
Distilled water Tap watera a
[HCO3−] (mg L−1)
[SO42−] (mg L−1)
[Cl−] (mg L−1)
[NO3−] (mg L−1)
T (%)
6.87
–
–
–
–
100
8.20
160
21
5
2
98.17
pH
(13)
The reaction rate constant of the OH% radical with SMZ was determined to identify the influence of these recombination reactions, obtaining a kSMZ/OH% value of 4.35 × 109 M−1 s−1, very similar to reported by other authors (Li et al., 2017). Given that the reaction rate constant of the OH% radical is lower with H2O2 (3.3 × 107 M−1s−1) (Reaction 5) than with SMZ, degradation of SMZ by OH% radicals may be favored until the initial H2O2concentration is reached. Table 6 Kinetic parameters obtained from the degradation of SMZ by indirect UVC in presence of radical promoting species. [SMZ]0 = 15 mg L−1, T = 298 K, time = 60 min.
Table 4 Chemical composition of the water matrix. Water
(10)
Data obtained from: www.emasagra.es/ESP/191.asp. 227
[H2O2]0 104 (M)
[PDS]0 104 (M)
k 104 (M−1 s−1)
% Degradation
1.47 4.4 0 0 0
0 0 1.47 4.4 0
6.20 12.27 4.42 12.89 4.10
89 100 80 100 77
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
3.4.2. SMZ photodegradation in presence of different dose of PDS PDS (S2O8−2) is a strong oxidant (E0 = 2.05 V) used to degrade pollutants such as SMZ (Ji et al., 2018). Table 6 shows that the SMZ degradation rate increase with increasing PDS concentration in solution. Apparently, UVC/PDS appeared to be more efficient than UVC/ H2O2 for SMZ degradation attributed to the quantum yield of 1.4 and 1.0 to SO4%- and H2O2, respectively (Ji et al., 2018). Based on the S2O8−2 reactions are very slow at room temperature, different methods have been proposed to activate organic molecule degradation, including the photochemical decomposition of S2O82− to generate sulfate radicals (SO4%-) (Reaction (14)) that can degrade SMZ by participating in highly oxidizing radical reactions (Reactions (14)–(17)). (14) (15) (16) (17)
Fig. 5. Influence of PDS as promoter of SO4%- radicals in the degradation of SMZ at different H2O2 concentrations. [SMZ]0 = 15 mg L−1, pH = 7, T = 298 K.
Recombination radical reactions (Reactions (18)–(23)) are available to compete in SMZ degradation. (18)
concentrations; initial H2O2 concentrations of 1.47 × 10−4 M and 4.4 × 10−4 M produce the maximum radical generation of 6.68 × 10−13 M and 1.54 × 10−12 M at 10 and 8 min, respectively. The generation of radicals subsequently decreases, although the SMZ degradation continues, attributable to direct photolysis by UVC radiation. On the other hand, at an initial SMZ concentration of 4.4 × 10−4 M PDS, a maximum of 3.4 × 10−5 M SO4%- radicals are generated (Fig. 5). Thus, in the UV/PDS system, the SMZ exposure to radicals is higher than system UV/H2O2.
(19) (20) (21) (22)
3.5. Economic analysis by different oxidation process
(23) The rate constant value of the SO4%- radical with SMZ was determined as kSO4%- = 7.88 × 108 M−1s−1 and in comparison, to the Reactions (18)–(23), all recombination radical reactions are sufficient to compete with the SMZ degradation reaction in the presence of this radical.
To demonstrate the possible application of the UVC, UVC/H2O2 and UVC/PDS for SAs degradation, we applied an economic analysis based in the EE/O concept (Yin et al., 2018). EE/O include the electrical energy of UVC lamp (EE/OUVC) and the consumption of the oxidants (Oxidant/O). Which can be calculated by the following equation (24):
EE / OTotal = EE / OUVC + EE /O =
3.4.3. Quantification of radicals generated Other way to define the efficiency of the different oxidation process is compare the quantification of radicals generated (Oh et al., 2017). Fig. 4 depicts the concentration of OH% radicals at different H2O2
P·t c + oxidant V ·log(ci/ ct ) log(ci/ ct )
(24)
−1
Where P is the energy input of UVC lamp (k Wh h ), t the irradiation time (h), V the reactor volume (L), ci and ct are the initial and the final concentration of SA, respectively and coxidant is the concentration of H2O2 or PDS. Therefore, the total cost of SMZ degradation was 1.235, 0.257, 0.181 kWhL−1 to UVC, UVC/H2O2 and UVC/PDS, respectively. SMZ is degraded mainly by reactive species oxidation, addition oxidant helped lower EE/O (Zhang et al., 2016a). UVC/PDS process cost less energy than UVC/H2O2 process because SMZ reacted rapidly with SO4−. 3.6. Degradation by-products By-products of SMZ degradation by different systems. Chromatography and mass spectrum studies identified by-products, listed in Table 7 alongside their molecular formula, molecular weight, and chemical structure. Most of these by-products were previously reported. Based on the identified by-products, the degradation pathway of SMZ is proposed in Fig. 6, showing that SO2 removal is first produced by direct SMZ photolysis through UVC radiation, obtaining by-product P1. Wan et al. (Wan and Wang, 2016) indicated that sulfate ions are the main intermediate by-products. Next, given the oxidation of OH% radical during the process, hydroxylation is expected to be a common reaction responsible for SMZ degradation, generating by-products P3-1,
Fig. 4. Influence of H2O2 as promoter of OH% radicals in the degradation of SMZ at different H2O2 concentrations. [SMZ]0 = 15 mg L−1, pH = 7, T = 298 K. 228
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Table 7 By-products of the direct photolysis of SMZ by UV radiation. Treatment System
By-product identified
Molecular formula
Molecular weight (g mol−1)
UVC UVC/ H2O2 UVC/PDS UVC UVC/ H2O2 UVC/PDS
P1-1
C12H14N4
214.25
(Batista et al., 2014)
P1-2
C6H10O2N4S
214.23
(Wan and Wang, 2017)
UVC UVC/PDS
P2
C6H9N3
123.14
(Batista et al., 2014; Liu and Wang, 2013)
UVC UVC/ H2O2 UVC/PDS
P3-1
C12H14N4O
230.25
(Dong et al., 2017)
UVC UVC/ H2O2
P3-2
C12H14N4O
230.25
(Dong et al., 2017)
UVC/PDS
P4
C12H12N4O2
244.24
–
UVC/PDS UVC/ H2O2
P5-1
C12H12N4O2
294.32
(Wan and Wang, 2017)
UVC/H2O2
P5-2
C12H14N4O3S
294.32
(Wan and Wang, 2017)
UVC/H2O2 UVC/PDS
P6
C12H12N4O3S
292.30.
(Yin et al., 2018)
UVC/H2O2 UVC/PDS
P7
C12H12N4O4S
308.30
(Yin et al., 2018)
Chemical structure
Ref.
those of normal human cells (Kadu et al., 2017). Cell viability < 75% is considered a toxic effect and cell viability > 75% a viable effect. Fig. 7 depicts the cytotoxicity results of the degradation by-products as a function of time. The viability of cell cultures was > 75% for all sulfonamides and was even higher for SDZ and SML than for the control. Comparison of by-products with the original products (time 0) shows a higher viability for the by-products, which are therefore considered less toxic than the original product. Similar behavior has founded to other of pharmaceutical products (Yaghmaeian et al., 2017).
P3-2, P5-1 or P5-2. Sági et al. (2015) reported that the basic initial reaction is the addition of the hydroxyl radical to the benzene ring or, in some cases, heterocyclic rings, forming radical intermediates of cyclohexadiene accompanied by a fall in pH due to the formation of SO42− and smaller organic acids. A break in the eSO2e and eNHe bond allows identification of by-product P2, of low molecular weight, as previously reported in the mineralization pathway (Dong et al., 2017). By attack nucleophilic and incision in the group eSO2e, the byproduct P4 was formed. The hydroxylation intermediates are the major products formed by sulfate radical attack on aromatics (Yuan et al., 2011). Adduction of the 7 N atoms, forming the nitroso- and nitro-substitutional SMZ through electrophilic reaction was the dominated pathways by PDS (P6 and P7) (Yin et al., 2018).
4. Conclusions In this analysis of the degradation of three sulfonamides (SMZ, SDZ, and SML) based on quantification of the UVC radiation, found low values of quantum yields at 1 h of treatment. The UVC dose widely used for water disinfection in treatment plants is inadequate to remove this type of antibiotic. Direct UVC photolysis of SAs is influenced by their
3.7. Cytotoxicity of degradation by-products The metabolic conditions of HEK-293 cells are highly similar to 229
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
Fig. 6. Mechanism of degradation by direct SMZ photolysis in the presence of UVC radiation.
Fig. 7. HEK-293% Cell viability in SA samples after different periods of UVC radiation. [SAs]0 = 5.38 × 10−6 M.
initial concentration, the degradation rates were higher at pH 12 (anionic specie) for SMZ and SML. The presence of sulfates, nitrates, and chlorides in tap water can exert a synergic effect on the photodegradation, especially in the cases of SMZ and SML. A concentration of 4.4 × 10−4 M of H2O2 or PDS increases SMZ degradation up to 100%. UVC/PDS cost less energy than UVC and UVC/H2O2. Hydroxylation reactions, binding between eSO2e and eNHe groups and electrophilic attacks were the aim way to degraded SAs. Ten by-products of UV radiation-induced SMZ degradation were identified and present a less cytotoxic in comparison to the original products.
sulfamethazine degradation using UV-based technologies and products identification. J. Photochem. Photobiol. A Chem. 290, 77–85. Białk-Bielińska, A., Stolte, S., Arning, J., Uebers, U., Böschen, A., Stepnowski, P., Matzke, M., 2011. Ecotoxicity evaluation of selected sulfonamides. Chemosphere 85, 928–933. Boreen, A.L., Arnold, W.A., McNeill, K., 2004. Photochemical fate of sulfa drugs in the aquatic environment: sulfa drugs containing five-membered heterocyclic groups. Environ. Sci. Technol. 38, 3933–3940. Burgos-Castillo, R.C., Sirés, I., Sillanpää, M., Brillas, E., 2018. Application of electrochemical advanced oxidation to bisphenol A degradation in water. Effect of sulfate and chloride ions. Chemosphere 194, 812–820. Canonica, S., Jans, U., Stemmler, K., Hoigne, J., 1995. Transformation kinetics of phenols in water: photosensitization by dissolved natural organic material and aromatic ketones. Environ. Sci. Technol. 29, 1822–1831. Cizmas, L., Sharma, V.K., Gray, C.M., McDonald, T.J., 2015. Pharmaceuticals and personal care products in waters: occurrence, toxicity, and risk. Environ. Chem. Lett. 13, 381–394. Cui, C., Jin, L., Han, Q., Lin, K., Lu, S., Zhang, D., Cao, G., 2016. Removal of trace level amounts of twelve sulfonamides from drinking water by UV-activated peroxymonosulfate. Sci. Total Environ. 572, 244–251. Dong, H., Qiang, Z., Lian, J., Qu, J., 2017. Degradation of nitro-based pharmaceuticals by UV photolysis: kinetics and simultaneous reduction on halonitromethanes formation potential. Water Res. 119, 83–90. Fan, Y., Ji, Y., Kong, D., Lu, J., Zhou, Q., 2015. Kinetic and mechanistic investigations of the degradation of sulfamethazine in heat-activated persulfate oxidation process. J. Hazard Mater. 300, 39–47. Iglesias, A., Nebot, C., Miranda, J.M., Vázquez, B.I., Cepeda, A., 2012. Detection and quantitative analysis of 21 veterinary drugs in river water using high-pressure liquid chromatography coupled to tandem mass spectrometry. Environ. Sci. Pollut. Control Ser. 19, 3235–3249. Ji, Y., Yang, Y., Zhou, L., Wang, L., Lu, J., Ferronato, C., Chovelon, J.-M., 2018. Photodegradation of sulfasalazine and its human metabolites in water by UV and UV/ peroxydisulfate processes. Water Res. 133, 299–309. Kadu, B.S., Wani, K.D., Kaul-Ghanekar, R., Chikate, R.C., 2017. Degradation of doxorubicin to non-toxic metabolites using Fe-Ni bimetallic nanoparticles. Chem. Eng. J. 325, 715–724.
Acknowledgments The authors are grateful for the financial support of the Ministry of Science and Innovation (CTQ2016-80978-C2-1-R). A. Acosta-Rangel is thankful to CONACYT for their financial aid granted of through the scholarship (Beca Mixta) No. 291062. References Babić, S., Zrnčić, M., Ljubas, D., Ćurković, L., Škorić, I., 2015. Photolytic and thin TiO2 film assisted photocatalytic degradation of sulfamethazine in aqueous solution. Environ. Sci. Pollut. Res. 22, 11372–11386. Baeza, C., Knappe, D.R., 2011. Transformation kinetics of biochemically active compounds in low-pressure UV photolysis and UV/H 2 O 2 advanced oxidation processes. Water Res. 45, 4531–4543. Baran, W., Sochacka, J., Wardas, W., 2006. Toxicity and biodegradability of sulfonamides and products of their photocatalytic degradation in aqueous solutions. Chemosphere 65, 1295–1299. Batista, A.P.S., Pires, F.C.C., Teixeira, A.C.S., 2014. The role of reactive oxygen species in
230
Journal of Environmental Management 225 (2018) 224–231
A. Acosta-Rangel et al.
N-nitrosodimethylamine in simulated drinking water. Environ. Sci. Technol. 37, 1933–1940. Souza, B.M., Souza, B.S., Guimarães, T.M., Ribeiro, T.F., Cerqueira, A.C., Sant'Anna, G.L., Dezotti, M., 2016. Removal of recalcitrant organic matter content in wastewater by means of AOPs aiming industrial water reuse. Environ. Sci. Pollut. Control Ser. 23, 22947–22956. Sun, L., Bolton, J.R., 1996. Determination of the quantum yield for the photochemical generation of hydroxyl radicals in TiO2 suspensions. J. Phys. Chem. 100, 4127–4134. Tan, C., Gao, N., Zhou, S., Xiao, Y., Zhuang, Z., 2014. Kinetic study of acetaminophen degradation by UV-based advanced oxidation processes. Chem. Eng. J. 253, 229–236. Velo-Gala, I., Pirán-Montaño, J., Rivera-Utrilla, J., Sánchez-Polo, M., Mota, A.J., 2017. Advanced Oxidation Processes based on the use of UVC and simulated solar radiation to remove the antibiotic tinidazole from water. Chem. Eng. J. 323, 605–617. Wan, Z., Wang, J.-L., 2016. Removal of sulfonamide antibiotics from wastewater by gamma irradiation in presence of iron ions. Nucl. Sci. Tech. 27, 1–5. Wan, Z., Wang, J., 2017. Fenton-like degradation of sulfamethazine using Fe3O4/Mn3O4 nanocomposite catalyst: kinetics and catalytic mechanism. Environ. Sci. Pollut. Control Ser. 24, 568–577. Wang, J., Wang, S., 2016. Removal of pharmaceuticals and personal care products (PPCPs) from wastewater: a review. J. Environ. Manag. 182, 620–640. Yaghmaeian, K., Moussavi, G., Mashayekh-Salehi, A., Mohseni-Bandpei, A., Satari, M., 2017. Oxidation of acetaminophen in the ozonation process catalyzed with modified MgO nanoparticles: effect of operational variables and cytotoxicity assessment. Process Saf. Environ. Protect. 109, 520–528. Yang, Y., Lu, X., Jiang, J., Ma, J., Liu, G., Cao, Y., Liu, W., Li, J., Pang, S., Kong, X., 2017. Degradation of sulfamethoxazole by UV, UV/H 2 O 2 and UV/persulfate (PDS): formation of oxidation products and effect of bicarbonate. Water Res. 118, 196–207. Yin, R., Guo, W., Wang, H., Du, J., Zhou, X., Wu, Q., Zheng, H., Chang, J., Ren, N., 2018. Enhanced peroxymonosulfate activation for sulfamethazine degradation by ultrasound irradiation: performances and mechanisms. Chem. Eng. J. 335, 145–153. Yuan, R., Ramjaun, S.N., Wang, Z., Liu, J., 2011. Effects of chloride ion on degradation of Acid Orange 7 by sulfate radical-based advanced oxidation process: implications for formation of chlorinated aromatic compounds. J. Hazard Mater. 196, 173–179. Zhang, H., Wei, X., Song, X., Shah, S., Chen, J., Liu, J., Hao, C., Chen, Z., 2018. Photophysical and photochemical insights into the photodegradation of sulfapyridine in water: a joint experimental and theoretical study. Chemosphere 191, 1021–1027. Zhang, R., Yang, Y., Huang, C.-H., Zhao, L., Sun, P., 2016a. Kinetics and modeling of sulfonamide antibiotic degradation in wastewater and human urine by UV/H2O2 and UV/PDS. Water Res. 103, 283–292. Zhang, R., Yang, Y., Huang, C.-H., Zhao, L., Sun, P., 2016b. Kinetics and modeling of sulfonamide antibiotic degradation in wastewater and human urine by UV/H 2 O 2 and UV/PDS. Water Res. 103, 283–292. Zhao, T., Li, P., Tai, C., She, J., Yin, Y., Qi, Y.a., Zhang, G., 2018. Efficient decolorization of typical azo dyes using low-frequency ultrasound in presence of carbonate and hydrogen peroxide. J. Hazard Mater. 346, 42–51. Zhu, W., Liu, J., Yu, S., Zhou, Y., Yan, X., 2016. Ag loaded WO 3 nanoplates for efficient photocatalytic degradation of sulfanilamide and their bactericidal effect under visible light irradiation. J. Hazard Mater. 318, 407–416. Zhu, Y., Wu, M., Gao, N., Chu, W., Li, K., Chen, S., 2018. Degradation of phenacetin by the UV/chlorine advanced oxidation process: kinetics, pathways, and toxicity evaluation. Chem. Eng. J. 335, 520–529.
Kelner, M., Bagnell, R., Welch, K., 1990. Thioureas react with superoxide radicals to yield a sulfhydryl compound. Explanation for protective effect against paraquat. J. Biol. Chem. 265, 1306–1311. Kurniawan, T.A., Lo, W.-h., Chan, G., 2006. Radicals-catalyzed oxidation reactions for degradation of recalcitrant compounds from landfill leachate. Chem. Eng. J. 125, 35–57. Li, M., Wang, C., Yau, M., Bolton, J.R., Qiang, Z., 2017. Sulfamethazine degradation in water by the VUV/UV process: kinetics, mechanism and antibacterial activity determination based on a mini-fluidic VUV/UV photoreaction system. Water Res. 108, 348–355. Lian, J., Qiang, Z., Li, M., Bolton, J.R., Qu, J., 2015. UV photolysis kinetics of sulfonamides in aqueous solution based on optimized fluence quantification. Water Res. 75, 43–50. Liu, Y., Wang, J., 2013. Degradation of sulfamethazine by gamma irradiation in the presence of hydrogen peroxide. J. Hazard Mater. 250, 99–105. Luo, S., Wei, Z., Spinney, R., Zhang, Z., Dionysiou, D.D., Gao, L., Chai, L., Wang, D., Xiao, R., 2018. UV direct photolysis of sulfamethoxazole and ibuprofen: an experimental and modelling study. J. Hazard Mater. 343, 132–139. Magureanu, M., Mandache, N.B., Parvulescu, V.I., 2015. Degradation of pharmaceutical compounds in water by non-thermal plasma treatment. Water Res. 81, 124–136. Ocampo-Pérez, R., Sánchez-Polo, M., Rivera-Utrilla, J., Leyva-Ramos, R., 2010. Degradation of antineoplastic cytarabine in aqueous phase by advanced oxidation processes based on ultraviolet radiation. Chem. Eng. J. 165, 581–588. Oh, W.-D., Dong, Z., Ronn, G., Lim, T.-T., 2017. Surface–active bismuth ferrite as superior peroxymonosulfate activator for aqueous sulfamethoxazole removal: performance, mechanism and quantification of sulfate radical. J. Hazard Mater. 325, 71–81. Pomati, F., Castiglioni, S., Zuccato, E., Fanelli, R., Vigetti, D., Rossetti, C., Calamari, D., 2006. Effects of a complex mixture of therapeutic drugs at environmental levels on human embryonic cells. Environ. Sci. Technol. 40, 2442–2447. Pomati, F., Orlandi, C., Clerici, M., Luciani, F., Zuccato, E., 2007. Effects and interactions in an environmentally relevant mixture of pharmaceuticals. Toxicol. Sci. 102, 129–137. Prados-Joya, G., Sánchez-Polo, M., Rivera-Utrilla, J., Ferro-Garcia, M., 2011. Photodegradation of the antibiotics nitroimidazoles in aqueous solution by ultraviolet radiation. Water Res. 45, 393–403. Rahmani, H., Gholami, M., Mahvi, A., Alimohammadi, M., Azarian, G., Esrafili, A., Rahmani, K., Farzadkia, M., 2014. Tinidazole removal from aqueous solution by sonolysis in the presence of hydrogen peroxide. Bull. Environ. Contam. Toxicol. 92, 341–346. Reina, A.C., Martínez-Piernas, A.B., Bertakis, Y., Brebou, C., Xekoukoulotakis, N.P., Agüera, A., Pérez, J.A.S., 2018. Photochemical degradation of the carbapenem antibiotics imipenem and meropenem in aqueous solutions under solar radiation. Water Res. 128, 61–70. Sági, G., Bezsenyi, A., Kovács, K., Klátyik, S., Darvas, B., Székács, A., Mohácsi-Farkas, C., Takács, E., Wojnárovits, L., 2018. Radiolysis of sulfonamide antibiotics in aqueous solution: degradation efficiency and assessment of antibacterial activity, toxicity and biodegradability of products. Sci. Total Environ. 622, 1009–1015. Sági, G., Csay, T., Szabó, L., Pátzay, G., Csonka, E., Takács, E., Wojnárovits, L., 2015. Analytical approaches to the OH radical induced degradation of sulfonamide antibiotics in dilute aqueous solutions. J. Pharmaceut. Biomed. Anal. 106, 52–60. Sharpless, C.M., Linden, K.G., 2003. Experimental and model comparisons of low-and medium-pressure Hg lamps for the direct and H2O2 assisted UV photodegradation of
231