Environ Sci Pollut Res DOI 10.1007/s11356-015-4338-5

RESEARCH ARTICLE

Photolytic and thin TiO2 film assisted photocatalytic degradation of sulfamethazine in aqueous solution Sandra Babić 1 & Mirta Zrnčić 1 & Davor Ljubas 2 & Lidija Ćurković 3 & Irena Škorić 4

Received: 24 July 2014 / Accepted: 9 March 2015 # Springer-Verlag Berlin Heidelberg 2015

Abstract This paper deals with the photolytic and the photocatalytic degradation of sulfonamide antibiotic sulfamethazine (SMT) dissolved in Milli-Q water and in synthetic wastewater. Besides the direct photolysis, oxidation processes including UV/H2O2, UV/TiO2, and UV/TiO2/H2O2 using UV-A and UV-C radiation were investigated. Pseudo-first-order kinetics was observed for the degradation of SMT in all investigated processes. Additions of an electron acceptor (H2O2) and a Responsible editor: Roland Kallenborn Sandra Babić is a full professor, Department of Analytical Chemistry, Faculty of Chemical Engineering and Technology, University of Zagreb. Mirta Zrnčić holds a PhD degree, Department of Analytical Chemistry, Faculty of Chemical Engineering and Technology, University of Zagreb. Davor Ljubas is an associated professor, Department of Energy, Power Engineering and Environment, Faculty of Mechanical Engineering and Naval Architecture, University of Zagreb. Lidija Ćurković is a full professor, Department of Materials, Faculty of Mechanical Engineering and Naval Architecture, University of Zagreb. Irena Škorić is an associated professor, Department of Organic Chemistry, Faculty of Chemical Engineering and Technology, University of Zagreb. Electronic supplementary material The online version of this article (doi:10.1007/s11356-015-4338-5) contains supplementary material, which is available to authorized users. * Sandra Babić [email protected] 1

Department of Analytical Chemistry, Faculty of Chemical Engineering and Technology, University of Zagreb, Marulićev trg 19, 10000 Zagreb, Croatia

2

Department of Energy, Power Engineering and Environment, Faculty of Mechanical Engineering and Naval Architecture, University of Zagreb, Zagreb, Croatia

3

Department of Materials, Faculty of Mechanical Engineering and Naval Architecture, University of Zagreb, Zagreb, Croatia

4

Department of Organic Chemistry, Faculty of Chemical Engineering and Technology, University of Zagreb, Zagreb, Croatia

catalyst (TiO2 film) accelerated the photolytic degradation of SMT for both the UV-A- and the UV-C-based processes. The most efficient process was UV-C/TiO2/H2O2 with complete degradation of SMT obtained in 10 min. The UV-A-based processes have been less efficient in terms of irradiation time required to totally degrade SMT than the UV-C-based processes. It was also confirmed that different wastewater components can significantly reduce the degradation rate of SMT. An almost ninefold reduction in the rate constant of SMT was observed for the specific synthetic wastewater. Although UVA radiation experiments need more time and energy (2.7 times more electrical energy was consumed per gram of demineralized SMT) than UV-C experiments, they have a potential for practical use since natural UV-A solar radiation could be used here, which lowers the overall cost of the treatment. Five degradation products were detected during the degradation processes, and their structural formulae are presented. The structural formulae were elucidated based on mass spectra fragmentation pattern obtained using the tandem mass spectrometry (MS/MS) and NMR analysis. Keywords Titanium dioxide film . Pharmaceuticals . Sulfamethazine . Photolysis . Photocatalysis . Photodegradation path . Wastewater

Introduction Pharmaceuticals are being continuously released into the aquatic environment through different anthropogenic activities. The frequent occurrence of pharmaceutically active compounds in different water bodies and associated environmental hazards has caused public concern due to their negative impact on human health and the environment. Research results have confirmed that the main points of collection and subsequent release of pharmaceuticals into the environment are

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wastewater treatment plants (WWTPs), where they come via domestic and hospital sewages or through industrial discharges (Carballa et al. 2004; Gros et al. 2010; Behera et al. 2011). As a result, pharmaceuticals have been found in WWTP effluents, and in surface, ground, and drinking waters (Gros et al. 2006; Grujić et al. 2009) as well as in river sediments and wastewater sludge (Jelić et al. 2009). Occurrence data of pharmaceuticals in WWTP are recently reviewed in several papers (Gros et al. 2010; Rivera-Utrilla et al. 2013; Verlicchi et al. 2013; Collado et al. 2014; Luo et al. 2014). Observed removal efficiencies vary in a wide range for the different compounds, as well as for the same substance, due to a number of factors, including different chemical and physical characteristics of pharmaceuticals, wastewater composition, technology used, design, and operational conditions of WWTP. In addition, reported concentrations of pharmaceuticals in WWTP influent and effluent point to the significant spatial and temporal variations, which could be due to the rate of production, specific sales and practices, excretion rate, and water consumption per person and per day (Petrovic et al. 2009; Jelic et al. 2012). Reported overall removal rates varied strongly between individual pharmaceuticals and therefore, it is difficult to establish a general trend for specific therapeutic group. In a general, three different behaviors were observed: (i) low removal (0–35 %) or an increase in concentration along the passage through the WWTPs (antibiotics (some sulfonamides, metronidazole, macrolide erythromycin), trimethoprim, anti-epileptic carbamazepine, benzodiazepines (lorazepam, diazepam), β-blocker propranolol, non-steroidal anti-inflammatory drug (NSAID) mefenamic acid, lipid-lowering agent fenofibric acid, and tramadol), (ii) medium removal efficiency (35–70 %) (psychiatric drug trazodone, analgesic phenazone, lipid regulator atorvastatin, antibiotics (ciprofloxacin, sulfamethoxazole, sulfamethazine), cholesterollowering statin drugs pravastatin and mevastatin, β-blockers timolol, β-agonist salbutamol, histamine H1 and H2 receptor antagonists, and anti-diabetic glibenclamide), and (iii) high removal efficiency (>70 %) (NSAIDs (naproxen, ketoprofen, salicylic acid, acetaminophen, and ibuprofen), lipid regulator bezafibrate, β-blockers atenolol, diuretic furosemide, psychiatric drug citalopram, histamine receptor antagonist ranitidine, antihypertensive valsartan, caffeine, antihypertensive enalapril, and estriol) (Gros et al. 2010; Jelic et al. 2011; Collado et al. 2014; Luo et al. 2014). Inconsistent removal was observed for diclofenac, whose removal rates varied from no elimination to up to 100 %. Also, according to reported data, removal of diuretics furosemide and hydrochlorothiazide, lipidlowering drug gemfibrozil, and β-blocker metoprolol vary from moderate to low removal category. These findings emphasize the necessity for further investigation on the removal possibilities of pharmaceuticals to minimize the environmental exposure of their residues.

Oxidation processes based on the generation of highly reactive hydroxyl radicals used for the unselective oxidation of pollutants have obtained promising results for the degradation of pharmaceuticals by producing more biologically degradable and less toxic degradation products. The most widely tested oxidation process for the degradation of pharmaceuticals include UV irradiation combined with H2O2 as a strong oxidant, the Fenton and the photo-Fenton oxidation, and heterogeneous photocatalysis (Klavarioti et al. 2009). Titanium dioxide (TiO2) was the most widely employed catalyst in studies dealing with the photocatalytic treatment of pharmaceuticals. Tong et al. (2012) reviewed the application of oxidation processes involving the TiO2 photocatalysis for the degradation of pharmaceuticals. The authors concluded that the TiO2-assisted photocatalysis is a potentially feasible wastewater treatment process, but careful consideration of treatment conditions is necessary for a successful application. In the most of photocatalytic studies, the TiO2 suspension was used for organic pollutant degradation. Efficiency of the suspended TiO2 catalyst has proved to be superior to the same catalyst immobilized on the substrate, which could be attributed to the enhanced mass transport in suspended form (Ahmed et al. 2011). Pablos et al. (2013) observed that the TiO2 potocatalysis in a fixed-bed reactor required a similar irradiation time to that in the slurry reactor to reach the complete oxidation of pharmaceuticals in simulated municipal wastewater. However, the cost of the catalyst recovery makes the slurry system impractical. The aim of this study was to investigate the photolytic and the photocatalytic degradation of sulfonamide antibiotic sulfamethazine (SMT). Sulfonamides are the most frequently detected antibiotics in wastewaters and surface waters, followed by fluoroquinolones, tetracyclines, and macrolides (Lindberg et al. 2005; Verlicchi et al. 2010). While the photochemical behavior of sulfamethoxazole was thoroughly investigated (Hu et al. 2007; Tong et al. 2012), investigations dealing with photochemical degradation of SMT are scarce (Kaniou et al. 2005; Ito et al. 2014). In these papers, the degradation of sulfonamides was investigated often in ultrapure water (Calza et al. 2004; Kaniou et al. 2005; Nasuhoglu et al. 2011). In real applications, the photocatalytic degradation could also be affected by many other factors, among them being the coexisting substances in aquatic system, such as humic acid, metal cations (Na+, K+, Ca2+, Mg2+, Cu2+, and Fe3+), anions (NO3−, HCO3−, Cl-, and SO42−) (Ahmed et al. 2011; Wang et al. 2012), and other wastewater components. To our knowledge, the TiO2 film-based photocatalysis of SMT in wastewater has not been investigated yet. This study reports the photolytic and the photocatalytic degradation of SMT using immobilized TiO 2 film in Milli-Q water and in synthetic wastewater (SWW), similar by its composition to the wastewater of the pharmaceutical industry. Apart from kinetic study, an important aspect

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of this research was to understand and suggest the degradation pathways and to investigate the influence of water components on the photodegradation of SMT.

Materials and methods

glass tube was cleaned thoroughly and dried. Then, the sol was poured into the borosilicate glass cylinder, kept there for 10 min, slowly poured out of it, and then dried, first at 100 °C for 60 min and then calcined at 500 °C for 4 h. This deposition process was repeated three times. The procedure for the film preparation as well as its characterization was described in details elsewhere (Šegota et al. 2011; Ćurković et al. 2014).

Chemicals and reagents Sulfamethazine of high purity (>99 %) was obtained from Veterina Animal Health (Kalinovica, Croatia). For the investigation on the photolytic and the photocatalytic degradation of SMT as well as for the identification of photodegradation products, the SMT solution in Milli-Q water with higher analyte concentration of 10 mg L−1 was used. To i n v e s t i g a t e t h e e ff e c t o f m a t r i x c o m p o n e n t s o n photodegradation, SWW with similar composition to that of the wastewater from the pharmaceutical industry with SMT concentrations of 10 and 0.1 mg L−1 was used. The SWW was prepared in tap water (pH=7.01, total organic carbon (TOC) 0.5474 mg L−1, conductivity 597 μS cm−1, Cl− 8.53 mg L−1, SO42− 7.31 mg L−1, NO3− 2.95 mg L−1, Na+ 9.71 mg L−1, NH4+ 0.2413 mg L−1, K+ 1.40 mg L−1, Mg2+ 17.48 mg L−1, and Ca 2+ 78.88 mg L −1 ) with the addition of NaCl (1000 mg L −1), citric acid (50 mg L −1 ), ascorbic acid (30 mg L −1), saccharose (100 mg L −1 ), and Na 2HPO 4 (230 mg L−1). The pH value of the SWW was 7.25. Chemicals used for the TiO2 film preparation were of high purity and were used as received from Kemika (Zagreb, Croatia): titanium(IV) isopropoxide (purity>98 %), ethanol, acetylacetone, acetic acid, and polyethylene glycol (PEG) (Mr =5000–7000). Chemicals used for the preparation of mobile phase in the liquid chromatography analysis were the HPLC grade acetonitrile and formic acid supplied by Kemika (Zagreb, Croatia). Ultrapure water was prepared by a Millipore Simplicity UV system (Millipore Corporation, Billerica, MA, USA).

Photolytic/photocatalytic experiments All experiments were carried out in two 0.11 L borosilicate glass cylinders: one with the TiO2 film, for photocatalytic experiments, and the other without the TiO2 film, for photolytic experiments. UV radiation lamps were placed in the middle of each reactor. Detailed experimental set-up is presented in (Ćurković et al. 2014). The radiation sources were 15 W mercury UV lamps: (a) model Pen-Ray CPQ-7427, with λmax =365 nm and (b) model Pen-Ray 90-0004-07, with λmax =185/254 nm, manufactured by UVP (Upland, CA, USA). Both lamps were used with the same electrical source PS-4, I = 0.54 A, from UVP, too. Experiments were performed at a temperature of 25±0.2 °C, with continuous purging with air (O2). Two-milliliter aliquots were withdrawn at regular time intervals and analyzed directly by liquid chromatography tandem mass spectrometry (LCMS/MS). For each condition, repetition tests were done to ensure reproducibility. Electrical energy consumption was measured by a VOLTCRAFT energy meter (model Energy Check 3000). The photolytic/photocatalytic experiments were performed under different conditions: under UV radiation alone (photolysis), under UV radiation with the addition of H2O2 (photolysis), under UV radiation in the presence of the sol-gel TiO2 film (photocatalysis), and under UV radiation in the presence of the sol-gel TiO2 film with the addition of H2O2 (photocatalysis). In addition, an experiment comprising of the TiO2 film and the H2O2 addition

Preparation of sol-gel TiO2 film The TiO2 nano film was prepared by dip coating using the solgel method. The TiO2 nanostructured film was deposited on a borosilicate glass cylinder (200 mm in height and 30 mm in diameter). Titanium(IV) isopropoxide (TIP), ethanol (EtOH), acetylacetone (AcAc), acetic acid (Aaacid), distilled water, and polyethylene glycol (PEG) were used as a Ti precursor, as a solvent, as a chelating agent, as a catalyst, as a gelling agent, and as an organic/polymer additive, respectively. The molar ratio of these reactants in starting mixture was TIP:EtOH:AcAc:AAcid:H 2 O = 1:40:0.9:1.3:12.5. The amount of PEG was 2 g per 100 mL of colloidal solution (sol). Prior to the deposition of the TiO2 film, borosilicate

Table 1 Photolytic/photocatalytic experiments (105 mL of SMT solution, 25±0.2 °C, purging with air) Experiment

UV radiation range

Reactor inner wall

H2O2 (30 %)

UVC-1 UVC-2 UVC-3 UVC-4 UVA-1 UVA-2 UVA-3 UVA-4 NOUV

UV-C UV-C UV-C UV-C UV-A UV-A UV-A UV-A Without radiation

Glass Glass TiO2 film TiO2 film Glass Glass TiO2 film TiO2 film TiO2 film

– 0.1 – 0.1 – 0.1 – 0.1 0.1

mL mL mL mL mL

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The degradation of SMT and the formation of its degradation products were monitored by LC-MS/MS. Liquid chromatography analysis was performed using an Agilent chromatograph (Santa Clara, CA, USA) series 1200 coupled to an Agilent 6410 triple quadrupole mass spectrometer equipped w i t h a n e l e c t r o s p r a y i o n i z a t i o n (E S I ) i nt e r f a c e . Chromatographic separation was achieved using a Synergy Fusion C18 embedded column (150 mm×2.0 mm, particle size 4 μm) supplied by Phenomenex. Mobile phase consisted of 0.1 % formic acid in Milli-Q water as eluent A and 0.1 % formic acid in acetonitrile as eluent B in gradient elution mode. Flow rate was 0.2 mL min−1 and injection volume was 5 μL for all analyses. The gradient elution started with 0 % of B, during 2.30 min it rose to 8 % of B followed with a raise to 10 % of B in the 6th min, in the 11th min to 30 %, in the 15th min to 60 %, and in the 18th min to 95 %. During a 10-min period, the content of B was constant at 95 %. In 28.10 min, the content of B dropped to 0 % and remained that way until the 40th min, which was the end of the run. Parameters for the mass spectrometer were the following: drying gas temperature 350 °C, capillary voltage 4.0 kV, drying gas flow 11 L min−1, and nebulizer pressure 35 psi. Positivemode ESI was used for ionization. Fragmentor voltages were optimized to obtain precursor ions of best abundance for the SMT and degradation products. After the selection of precursor ions, product ions were obtained based on optimized collision energies. Instrument control, data acquisition, and evaluation were done using the Agilent MassHunter 2003–2007 Data Acquisition for Triple Quad B.01.04 (B84) software. Inorganic anions (NO3−, NO2−, SO42−) were quantified by a Shimadzu ion chromatograph (Kyoto, Japan) HIC-20A Super using a Shim-Pack IC-SA3 column (Shimadzu), while inorganic cation (NH4+) were quantified using a SHODEX YS-50 column (Showa Denko America, NY, USA). Mobile phases were 3.6 mM Na2CO3 for the determination of anions and 4 mM methanesulfonic acid for the ammonium ion determination. The extent of mineralization was evaluated by measuring changes in TOC, using the non-purgeable organic carbon (NPOC) method, on a Shimadzu TOC analyzer (model TOC-VCPH).

Results and discussion Photolysis/photocatalysis of SMT Single UV treatment SMT was subjected to photolysis in ultrapure water under UVC (experiment UVC-1) and UV-A radiation (experiment UVA-1). The results from the Bdark^ control (experiment NOUV) demonstrated the absence of SMT degradation in the absence of radiation. Complete SMT degradation was observed in the UV-C-irradiated solution over a period of 60 min (Fig. 1a) while no SMT degradation was observed in the solution irradiated with UV-A light (Fig. 1b) over a period of 120 min. The results obtained for photolysis experiments are expected due to the SMT solution absorption spectra, with two absorption peaks (one at 240 nm and the other at 262 nm), which are close to the predominant radiation wavelength of applied

a 1 0.8 C/C0

Analytical procedures

DMSO-d6 and measured in 5 mm NMR tubes. The 1H chemical shift values (δ) are expressed in ppm referring to DMSO (2.54 ppm). The NMR peak assignments of the photomixture were analyzed in detail. Assignments of 1H NMR spectra were performed on the basis of chemical shifts, signal intensities, magnitude, and multiplicity of H–H coupling constants.

0.6

UVC-1 UVC-2

0.4

UVC-3 UVC-4

0.2

NOUV 0 0

20

40

60 Time (min)

80

100

40

60 Time (min)

80

100

b

120

1 0.8 C/C0

without UV radiation was carried out. A total of nine experiments were performed (Table 1.).

0.6

UVA-1 UVA-2

0.4

UVA-3

Nuclear magnetic resonance spectroscopy

0.2

UVA-4 NOUV

1

H nuclear magnetic resonance (NMR) spectra of SMT antibiotic were recorded in high-resolution NMR spectrometer Bruker Avance 300, operating at 300 MHz for 1H resonance and for 13C at 75 MHz. The samples were dissolved in

0 0

20

120

Fig. 1 Photolytic/photocatalytic degradation profiles of sulfamethazine under UV-C radiation (a) and under UV-A radiation (b)

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UV-C radiation (near 254 nm). Absorption of the SMT solution in the area of the predominant UV-A radiation (near 365 nm) is very low, which explains the absence of degradation in this case. It was assumed that the direct photolysis of SMT with UVC radiation follows the first-order kinetics and a plot of ln(C/ C0) vs time (t) was made. C is the concentration of the SMT at a particular time and C0 is the initial SMT concentration. Linear regression analysis confirmed our assumption (r2 > 0.98), and the first-order rate constant for the SMT photolysis under UV-C irradiation was determined, kUV =0.111 min−1 (Table 2). UV/H2O2 treatment Prior to UV/H2O2 experiments, SMT was allowed to contact H2O2 (1.86·10−2 M) for 24 h at 25 °C to check if there was any reactivity between SMT and H2O2. The results indicate that SMT had no reactivity with H2O2, i.e., H2O2 under defined experimental conditions does not oxidize SMT. Since the molar extinction coefficient of H2O2 is very small compared to the molar extinction coefficient of SMT (ε= 22083.33 L mol−1 cm−1 at 254 nm), the amount of UV light adsorbed by the SMT was larger than that absorbed by H2O2. The addition of H2O2 favored the SMT degradation in the investigated photolysis systems. Under UV-C radiation (experiment UVC-2), complete degradation of SMT occurred in only 10 min (Fig. 1a). With the addition of H2O2 to the SMT solution irradiated under UV-A radiation (experiment UVA2), degradation occurred but it was very slow in comparison with the UV-C experiment (only 50 % of SMT degraded in 120 min) (Fig. 1b). The mechanism of SMT oxidation by UV/H2O2 has been investigated applying the oxidation mechanism previously proposed in literature (Benitez et al. 2013; Chelme-Ayala et al. 2010). The authors supposed that the degradation of an organic compound by means of the UV/H2O2 process results from the contribution of two pathways: direct photolysis and hydroxyl radical attack: ln½C 0  0 ¼ k UV t þ k OH t ¼ k UV=H2 O2 t ½C 

ð1Þ

where k UV=H2 O2 represents the pseudo-first-order rate constant for the overall degradation during the UV/H2O2 treatment,

Table 2

Photolytic/photocatalytic degradation rate constants (min−1) kUV

UV-A UV-C

– 0.111

k UV=H2 O2 0.006 0.539

k'OH 0.006 0.428

k UV=TiO2

k UV=TiO2 =H2 O2

0.011 0.101

0.006 0.776

kUV is the first-order rate constant for the direct UV photolysis, and k'OH is the pseudo-first-order rate constant for the reaction between hydroxyl radicals and the substance to be oxidized. According to Eq. (1), the rate constant for radical reaction k'OH can be simply deduced by subtracting kUV from k UV=H2 O2 . These values are presented in Table 2. The results suggest that the presence of H2O2 enhanced the degradation rate of SMT under UV-C irradiation. In fact, the contribution to the overall reaction rate of the direct photolysis was only 20 % compared to 80 % of the radical pathway. The study of Benitez et al. (2013) also showed a greater influence of radical pathway in the combined UV/H2O2 treatment of several emerging contaminants. Although it was reported that radiation with a wavelength lower than 400 nm is able to photolyse the H2O2 molecule (Esplugas et al. 2002), for the UV-A experiments a relatively small rate constant of radical reaction (k'OH) was obtained when compared with the UV-C experiments (Table 2). This is due to the fact that the molar extinction coefficient of H2O2 in the UV-A region of electromagnetic spectrum is much lower (0.0066 L mol−1 cm−1 at 365 nm) than in the UV-C region (19.6 L mol−1 cm−1 at 253.7 nm) (Baxendale and Wilson 1957). UV/TiO2 treatment Photocatalytic experiments were performed using borosilicate glass cylinders with the TiO2 film immobilized on the inner reactor wall. In the preliminary experiment, without UV radiation, no significant dark adsorption of SMT onto the TiO2 film was observed (experiment NOUV in Fig. 1a, b). Under UV-C radiation (experiment UVC-3), the TiO2 film did not accelerate the SMT degradation with respect to photolysis experiments (Fig. 1a). However, for experiments performed under UV-A radiation (experiment UVA-3), the TiO2 film had an important effect and raised the SMT conversion rate, but complete mineralization was not achieved even after 120 min (Fig. 1b). Since no degradation was found when the SMT solution was irradiated with UV-A alone, the observed degradation from the UV-A/TiO2 process could be solely derived from the TiO2 photocatalysis of SMT. Photocatalytic degradation rate of organic substances can be described by the pseudo-first-order kinetics when a substance is present at a low concentration (Ahmed et al. 2011): ln

½C 0  ¼ k UV=TiO2 t ½C 

ð2Þ

where k UV=TiO2 is the apparent rate constant of the UV/TiO2 degradation process. Pseudo-first-order rate constants were calculated as 0.011 and 0.101 min−1 for the UV-A/TiO2 and the UV-C/TiO2 photocatalytic system, respectively (Table 2).

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The addition of H2O2 to the UV-C/TiO2 system (experiment UVC-4) resulted in the fastest degradation in all investigated systems. Complete SMT degradation was achieved in less than 10 min (Fig. 1a). It is well known that the addition of H2O2 to the reaction solution leads to the production of additional hydroxyl radicals. Moreover, the direct photolysis of H2O2 may also occur and this would also generate hydroxyl radicals, thus enhancing the SMT degradation. However, the addition of H 2 O 2 in the UV-A-based photocatalysis system (experiment UVA-4) inhibited the SMT degradation. The degradation of SMT in the UV-A/ TiO2/H2O2 system was slower in comparison to that in the system without H2O2 and equal to that of the UV-A/H2O2 system (Fig. 1b). This indicates that, in the UV-A/TiO2/ H2O2 system, the heterogeneous process might at some degree be inhibited by H2O2. Some other authors also observed that the addition of H2O2 to the UV/TiO2 system could inhibit the degradation of organic compounds. They found that the addition of low H2O2 concentrations would increase the degradation rate while higher concentrations would reduce the degradation rate of organic compounds (Kaniou et al. 2005; Tanaka et al. 1990; Wang and Hong 1999). The inhibition of the heterogeneous process could be explained by the fact that the H2O2 sorbed on the TiO2 surface can scavenge TiO2 surface-formed •OH radicals and photogenerated holes and can thus inhibit the major pathway of heterogeneous generation of •OH radicals (Wang and Hong 1999). Furthermore, H2O2 can react with TiO2 and can form surface peroxytitanium complexes that irreversibly block the active sites of the catalyst (Wang and Hong 1999). Also, there is a possibility of competition between organic compounds and H2O2 for the adsorption sites onto the TiO2 surface (Amalric et al. 1994). It seems that homogeneous photodegradation competes with heterogeneous photocatalysis in the UV/TiO2/H2O2 systems, which could be an explanation for our results obtained for the SMT photocatalysis under UV-A radiation. Pseudo-first-order rate constants for the UV/TiO2/H2O2 processes are given in Table 2.

a 1 0.8 TOC/TOC0

UV/TiO2/H2O2 treatment

importance in water treatment since it is an unequivocal evidence of the total decomposition of the compound in water. The extent of the TOC reduction vs the time of irradiation for the photolytic and the photocatalytic degradation of SMT solution is shown in Fig. 2. Under UV-C radiation, 65 % of the initial carbon present in SMT was converted to CO2 in 2 h, with the exception of photolytic experiment performed only under UV-C radiation (experiment UVC-1) where only a 35 % reduction was obtained. In the same period, the SMT degradation was completed (Fig. 1a) as well as the formation and the complete degradation of SMT degradation products (Fig. 3). The obtained results suggest that, despite the fact that almost the same rate constant of SMT degradation was obtained in UVC-1 (only UV-C) and UVC-3 (UV-C/TiO2) experiments, the presence of TiO2 film had a positive influence in terms of faster degradation of SMT degradation products and faster TOC reduction. Photolytic/photocatalytic processes under UV-A radiation were significantly less efficient in the TOC reduction. Only 10 % reduction was observed for processes that include TiO2 as a catalyst (UVA-3 and UVA-4), while no TOC reduction was observed in photolysis experiments (Fig. 2b). The complete degradation of SMT was not reached during the test (Fig. 1b), while the presence of intermediates was observed (Fig. 3).

0.6

UVC-1 UVC-2

0.4

UVC-3 UVC-4

0.2

NOUV 0 0

20

40

60 Time (min)

80

100

120

40

60 Time (min)

80

100

120

b 1 0.8 TOC/TOC0

Comparing the obtained k UV=TiO2 values for the SMT degradation with corresponding rate constants for the direct photolysis, one could conclude that there is no contribution of the TiO2 photocatalysis for the UV-C system. For the UV-A system, photocatalysis on the TiO2 film plays an importation role in the SMT degradation.

0.6

UVA-1 UVA-2

0.4

UVA-3 UVA-4

0.2

NOUV

Mineralization The complete degradation of an organic compound to CO2 via photolysis/photocatalysis (mineralization) is of great

0 0

20

Fig. 2 Changes of the TOC during the experiments under the UV-C radiation (a) and under the UV-A radiation (b) (TOC0 =5.46 mg L−1)

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A/A0

0.16

0.5

DP124

0.14

DP140

0.12

DP215

0.10

DP231

DP124 DP140 DP215

DP295

0.08

UVA-2

0.4

A/A0

0.18

0.06 0.04

0.3

DP231 DP295

0.2 0.1

0.02 0

20

40

0.14

60 Time (min)

80

100

UVC-3

0.12

DP231 DP295

0.02

0.01

20

40

60 Time (min)

80

100

120

80

100

120

80

100

120

UVA-3

DP295

0.03 0.02

0

60 Time (min)

DP215

0.04

0.04

0.00

40

DP140

0.05

DP215

0.06

20

DP124

0.06

DP140

0.08

0

0.07

DP124

0.10 A/A0

0

120

A/A0

0.00

0

0

20

40

60 Time (min)

0.12

UVA-4

0.10

A/A0

0.08

DP124 DP140

0.06

DP231

0.04

DP295

0.02 0.00

0

20

40

60 Time (min)

80

100

120

Fig. 3 Evolution of degradation products during degradation of sulfamethazine (A is the peak area observed for a certain DP and A0 is the peak area for the initial SMT concentration)

Comparing Fig. 2 and Fig. 3, one can conclude that in some cases (UV-C experiments) SMT and its degradation products were totally degraded in 120 min although their TOC was not. This is a result of the formation of additional degradation products that are not followed within this study, the socalled Bminor degradation products^ as it was suggested by Oppenländer (2003), which are usually in the range of C1–C5 molecules and their final mineralization would need additional reaction time. Degradation pathways In order to identify the SMT degradation products (DPs) of photolytic/photocatalytic degradation and follow up their

evolution and disappearance as well as define degradation pathways, single analyte solutions in Milli-Q water were prepared at a concentration of 10 mg L−1. Degradation profiles were plotted as normalized chromatographic peak areas in percentage against irradiation time. A normalized peak area represents the peak area observed for a certain DP (A) divided by the peak area for the initial SMT concentration (A0). Since response factors in the ESI-MS analysis are dependent on the chemical structure of the analyte and the matrix/solvent composition during ionization, no conclusions on the absolute concentrations can be done without appropriate standards. The structures obtained as SMT degradation products were suggested primarily on the basis of mass spectra.

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The first step in the identification of degradation products is a comparison between the LC-MS chromatograms of irradiated SMT samples and the chromatograms of the initial sample and standard solution. The appearance of new peaks on Fig. 4 Total ion chromatograms obtained at t=0 (a) and after 30 min of degradation under 254 nm and with TiO2 film (b)

chromatograms for samples obtained at different irradiation times is indicative of possible degradation products. Figure 4 shows the total ion chromatogram (TIC) obtained after 30 min of degradation under UV-C irradiation and with the TiO2 film

a SMT m/z 279

b

DP124 m/z 124

DP140 m/z 140

DP231 m/z 231

DP215 m/z 215

DP295 m/z 295 SMT m/z 279

Environ Sci Pollut Res

(experiment UVC-3) where new peaks can be observed. A total of four different peaks, not including SMT, can be observed. Based on the mass spectra of each peak, it was concluded that five different DPs were the result of degradation process. Two DPs have very similar retention time (m/z 124 and m/z 140) and therefore came as one peak on the TIC, but based on the mass spectra and increase in the peak area during the degradation, it was concluded that there are two different degradation products. All DPs have shorter retention times than SMT, which leads to a conclusion that intermediates are more polar than SMT. During the photocatalytic process, •OH radicals are being formed when water from the solution reacts with TiO2 under irradiation. Hydroxyl radicals are non-selective and can attack the investigated molecule at any functional group. According to that, several degradation pathways were suggested which involve the attack of hydroxyl radical at different places on the SMT molecule (Fig. 5). Based on current structure, it is not possible to conclude in which ring the radical attack occurred, but since the amino group activates the aromatic ring toward an electrophilic substitution, the attack probably occurs on the anilinic ring (Calza et al. 2004). The addition of a hydroxyl radical on the anilinic ring forms degradation product DP295, the hydroxyl-sulfamethazine. Pathway 1 describes the formation of degradation product DP124 where the hydroxyl radical attacks the sulfonamide bond, resulting in the cleavage of the S–N bond. The subsequent addition of •OH to the DP124 forms degradation product DP140. Degradation products DP215 and DP231 are probably formed when the •OH attacks the heterocyclic ring, resulting in its opening. For degradation product DP231, the hydroxylation of the SO2 group occurred together with the opening of the heterocyclic ring. The proposed degradation pathway is in good agreement with other

CH3

researches that studied the photocatalytic degradation of sulfonamides (Calza et al. 2004; Abellan et al. 2007; Hu et al. 2007; Trovo et al. 2009; Yang et al. 2010). On the other hand, there are alternative explanations for the degradation pathway of sulfonamides. It does not include the opening of the heterocyclic ring but the degradation is described through the extrusion of the SO2 group. These studies dealt with enzymatic degradation (Schwarz et al. 2010; García-Galán et al. 2011), biodegradation (García-Galán et al. 2012), and photolysis (Boreen et al. 2005; Garcia-Galan et al. Garcia-Galan et al. 2012; Periša et al. 2013; Zessela et al. 2014). Information about degradation products, i.e., proposed structural formulas and conditions on the mass spectrometer that include fragmentor voltage and collision energies, can be found in Table 3. Values of fragmentors and collision energies that are shown are those values that gave the highest abundance for a certain precursor ion or product ion. Fragmentation patterns (Fig. S1 in Supporting information) for each degradation product were used to confirm the proposed structural formulae. Degradation products DP124 and DP140 fragmented in that way that the heterocyclic ring opened. Fragments obtained from degradation product DP215 were used to confirm the proposed molecular structure. Degradation product DP295 fragmented in a typical sulfonamide fragment which, in this case, included the hydroxyl group. The fragment m/z 124 is typical for SMT, which corresponds to a heterocyclic ring with the amino group (Fig. S1). From the 1 H NMR structural analysis of the SMT photomixture (Figs. 6b and 7b), it was concluded that in comparison with the starting SMT (Figs. 6a and 7a), new signals are found (Figs. 6b and 7b). The presence of the five photodegradation products assumed by MS spectra can be approved by the chemical shifts, coupling constants, and

CH3

H2N 1

N

O

N

3

H2N

CH3

OH

+ • OH S

S H2N

N

CH3

NH

O

1

DP124

N

O

N

Sulfamethazin

CH3

2

NH

O

N

DP295

2

+ • OH H2N

3

O

CH3 S

OH

N

OH S

H2N

N DP140

CH3

NH2

HO

O

NH

NH

CH3

DP231

Fig. 5 Proposed transformation pathways for sulfamethazine photocatalytic degradation

NH DP215

NH

CH3

Environ Sci Pollut Res Table 3 Sulfamethazine and assumed structural formulas of degradation products together with retention times and conditions on MS/MS that showed highest abundances for precursor and product ions

integrals of the new signals of the products mixture. It was not possible to isolate the photodegradation products, due to the

concentrations of the samples needed in the experiments. According to the MS analyses of the photomixture after the

2.2452

Environ Sci Pollut Res

a SMT, 2CH3

b

2.230

2.220

2.210

2.200

2.190

2.180

2.170

2.160

2.150

2.1680

2.240

2.2050

2.250

2.2217

2.260

2.2270

2.270

2.2517

2.280

2.2574

PPM 2.290

DP295, 2CH3 DP124, 2CH3

DP140, 2CH3

DP231, 2CH3

PPM

2.290

2.280

2.270

2.260

2.250

2.240

2.230

2.220

2.210

2.200

2.190

2.180

2.170

2.160

2.150

1

Fig. 6 Aliphatic part of the H NMR spectrum of SMT in DMSO-d6 (a) and photomixture of SMT in DMSO-d6 (b)

UVC-1 experiment (the same experiment for the NMR analysis), it was concluded that the proportion of the starting SMT was less than 10 % (Fig. 1a). Although the signals of SMT is overlapped with the signals of the photodegradation products since the parts of the structures DP215 and DP295 are very similar to the starting compound, it is evident that there is a lack of a signal for the NH proton in the neighborhood of the SO2 functional group at about 11 ppm. The photomixture in the aliphatic part of the 1H NMR spectrum showed several new signals related to the symmetrical methyl groups of the photoproducts DP295, DP231, DP140, and DP124 (Fig. 6b). The photodegradation product with m/z 215 was confirmed by NMR analysis as the major photoproduct with an elemental composition of C12H15N4. For DP215, two doublets of the para-substituted aromatic ring and one singlet were detected in the aromatic region of the proton NMR spectrum in expected very similar chemical shifts in comparison to SMT (Fig. 7). The aromatic part of the molecule of DP295 shows also similar chemical shifts as well as expected similar splittings as SMT (Fig. 7a). Regardless the position of the hydroxy group on the aromatic ring of DP295, two doublets and one singlet are expected in the spectrum.

The third photodegradation product of SMT with an observed m/z 124 and an elemental composition of C6H10N3 is the result of the sulfonamide bond cleavage. In the 1H NMR spectrum of the SMT photomixture, the expected minor singlets in the aliphatic/aromatic region can be seen and attributed to the aromatic proton/methyl group of product DP124. Photodegradation product DP140 is confirmed by NMR analysis due to the presence of three new singlets which can be attributed to the amino, hydroxy, and methyl group of DP140. Another series of signals can be associated with the structure of DP231 due to the new small signals in the aromatic region and aliphatic region as the result of the cleavage of the heteroaromatic ring. It is important to notice that the integrals of the sets of signals attributed to the exact photodegradation products are in agreement with the proposed desulfonated and/or hydrohylated products. The combination of the information obtained from mass spectra and new/all signals in the 1 H NMR spectrum verify the proposed structures. The evolution of degradation products during the period of degradation was monitored. In all experiments where UV-C irradiation was used, all five degradation products occurred and were degraded completely during the experiment (Fig. 3). Presence of the TiO2 film as a catalyst had an impact

5 .93 77

6 .55 62 6 .54 16

7 .64 04 7 .62 59

6 .74 18

Environ Sci Pollut Res

a SMT, d, 2Har

SMT, d, 2Har

2 .26 7

2 .24 9

6.4

6.2

6.0

5.8

5.6

5.4

5 .67 30

6.6

5 .93 58

6.8

6 .26 22

7.0

SMT, s, 2H, NH2

6 .60 97 6 .60 05 6 .59 54 6 .59 41 6 .59 07 6 .58 67 6 .57 24 6 .56 33 6 .56 06 6 .55 78 6 .35 11

7.2

6 .77 11 6 .76 97 6 .76 25 6 .76 14 6 .75 30

7.4

7 .06 65

7.6

6 .96 57

7.8

7 .01 43

8.0

7 .28 67 7 .28 32 7 .25 35 7 .25 02 7 .23 97 7 .23 63

8.2

7 .64 76 7 .64 48 7 .63 83 7 .63 30 7 .61 92 7 .60 74 7 .60 49

PPM

0 .99 9

2 .12 4

SMT, s, 1Har

b DP215, s, 1Har DP215, d, 2Har

DP295 s, 1Har

DP124 DP295 d, 2Har DP215, s, NH DP295, s, NH

PPM

8.2

8.0

DP231

7.8

7.6

7.4

DP215, d, 2Har DP295 d, 2Har

DP295, s 2H, NH2 DP215, s, 2H, NH2

DP124

DP215, s, 1H, NH

DP140

DP124

DP231

DP231

DP231

DP140

7.2

7.0

6.8

6.6

6.4

6.2

6.0

5.8

5.6

5.4

Fig. 7 Aromatic part of the 1H NMR spectrum of SMT in DMSO-d6 (a) and photomixture of SMT in DMSO-d6 (b)

on the elimination of degradation products. It accelerated the degradation of all degradation products so they were completely degraded in 60 min. In all UV-A experiments, degradation products were not completely eliminated during the experiment. Also, in the UVA-1 experiment, the SMT did not degrade, so no degradation products were formed. In experiments UVA-3 and 4, only four degradation products were formed. Degradation product DP124 was the most abundant one in all experiments, proving that the degradation pathway of sulfonamide bond cleavage represents the main degradation pathway. During the SMT degradation, a sulfur atom is released from SMT and/or from the formed degradation products (DP215, DP231, and DP295) and in 120 min of irradiation, 76.4–80.4 % of the sulfur stoichiometric amount (depending on the UV-C-based process employed) was released as sulfate ions. In parallel, only a small amount of organic nitrogen was

released, 26.2–29.4 % of nitrogen was transformed to NH4+ and 0.8–5.2 % to NO3−. A trace amount of NO2− was detected, but only at the beginning of the experiments and after longer irradiation, it oxidized to NO3−. A significant amount of nitrogen is probably present in some stable compounds that hardly release nitrogen. The study of Maurino et al. (1997) showed that the photocatalytic degradation of structures like aminophenols is slow due to the back reduction of the initial oxidation intermediates of aminophenols. Furthermore, the authors concluded that the degradation pathway of aminophenols account for the slow reduction in TOC. Conversion of organic sulfur and nitrogen into corresponding inorganic ions during the UV-A based experiments was less efficient since the complete degradation of SMT was not achieved. Only 18.8–31.6 % of sulfur was transformed to SO42−, 5.5–8.9 % of nitrogen was transformed to NH4+, and less than 1 % to NO3−.

Environ Sci Pollut Res

The pH values during the photolysis/photocatalysis experiments decreased from 6.00±0.50 to 4.20±0.60, but no precipitation was observed. The decrease in the pH value of up to two units during the experiments is in accordance with the simultaneous formation of SO42− (i.e., H2SO4) and NH4+, with faster H2SO4 formation.

Effect of water quality The experimentally confirmed most effective SMT photodegradation system—UV-C/TiO2/H2O2—was finally used to investigate the SMT photocatalytic fate in wastewater whereas the wastewater contained different matrix components that could influence the efficiency of photocatalytic degradation process. For that purpose, the SWW with a composition similar to that of the wastewater from the pharmaceutical industry was used. Besides the SMT degradation, the evolution of degradation products was also monitored. The same five DPs as in the experiments with SMT dissolved in Milli-Q water were detected with the DP124 as most abundant. The degradation profile of SMT in the SWW is shown in Fig. 8. As can be seen, the photocatalytic degradation of SMT is significantly slower in SWW than in Milli-Q water, with almost ninefold reduction in the photocatalytic degradation rate constant (k= 0.089 min−1 in SWW). The inhibition of SMT degradation can be attributed to the effects of different water constituents on the photodegradation (Guillard et al. 2003; Wang et al. 2004; Lair et al. 2008). On the basis of the UV spectra of SWW constituents (citric acid, ascorbic acid, and saccharose), only the citric acid could impact the photolysis-decreasing photodegradation rate of SMT because it absorbs UV-C radiation (λmax at 264 and 261 nm) in the close wavelength range as the UV-C lamp. Furthermore, humic acid (HA), inorganic cations and anions present in the SWW could also affect the SMT photodegradation. It has been shown that higher concentrations of HA, Cl−, PO43−, and NO3− decrease the photocatalytic efficiency of organic compounds (Rabindranathan et al. 1 0.9

MilliQ water (10 ppm)

0.8

SWW (10 ppm)

0.7

SWW (0.1 ppm)

C/C0

0.6 0.5 0.4 0.3 0.2 0.1 0

0

10

20

30

40 50 Time (min)

60

70

80

90

Fig. 8 Photocatalytic (UV-C/TiO2/H2O2) degradation profiles of SMZ in Milli-Q and synthetic wastewater (SWW)

2003; Chen et al. 2010), which is attributed to the competition among these species for active sites on the TiO2 surface and the catalyst deactivation, which subsequently decreased the degradation rate. As explained in several studies (Mahmoodi et al. 2007; Wu et al. 2009), inhibition is due to the reaction of positive holes and hydroxyl radicals with anions that behave as radical scavengers, resulting in the inhibition of photocatalytic degradation. In addition, natural organic matter could act as an inner filter which absorbs radiation in the UV-Vis region (Tong et al. 2012). Investigations of Shifu and Yunzhang (2007) showed that several cations (for example, Na+, K+, Mg2+, Ca2+) have a significant effect on the photocatalytic degradation rate of pesticide glyphosate, and Naeem and Feng (2008) also reported that Ca2+, Mg2+, and Cu2+ hindered the photolytic degradation process of phenol. When a low initial amount of SMT (100 μg L−1) was added to the SWW, the complete degradation of SMT was achieved in 15 min (Fig. 8) with a higher degradation rate (k = 0.432 min−1). Furthermore, as the SMT concentration increases, some of the UV light photons are absorbed by a substantial amount of SMT molecules and the quantity of effective photons which could be absorbed by the TiO2 surface is reduced and therewith the overall degradation rate becomes lower. Electrical energy consumption The results show that the electrical energy consumption of the UV-C photocatalytic system is slightly higher than that of the UV-A system. Energy consumption for the mineralization of SMT (kW h per gram of TOC-reduction) was calculated following the procedure described in Kopf et al. (2000) and Bolton et al. (2001), separately for the UV-C and the UV-A system. The energy consumption of these two processes was compared according to the consumption of electric power (kW h) of the laboratory devices (UV lamp, magnetic stirrer, and air compressor) under the experimental conditions. Energy for the production of H2O2 was not included in the calculation. During the experiments, due to the gradual warming of the UV lamps, electrical energy consumption of the experimental system slightly increased. Average electric energy consumption during 2 h were 22.8 W h for UV-A experiments and 23.5 W h for UV-C experiments. Among the UV-C experiments, the most effective process was UVC-4 (UV-C/TiO2/H2O2) with 150.34 kW h g−1 TOC reduction, while among the UV-A experiments, the most effective experiment was UVA-4 (UV-A/TiO 2 ) with 406.63 kW h g−1 TOC reduction, which is 2.7 times higher than in the UVC-3 experiment. Compared with other photocatalytic studies (Kopf et al. 2000; Alaton et al. 2002), these electrical energy consumption rates are rather high, but in those studies, the photocatalyst was in the form of slurry,

Environ Sci Pollut Res

which enhances the kinetics of degradation and therewith the electrical energy efficiency per gram of degraded/ demineralized substance. It should also be mentioned that the UV-A radiation source for the process of SMT degradation (and of other organic pollutants) could be the UV-A part of solar radiation that reaches the Earth’s surface (Wöhrle et al. 1998; Ljubas 2005). If solar radiation is used, then the costs of ensuring UV-A radiation could be significantly lowered, which should be considered in future experiments and in the real-scale reactor design.

Conclusions Oxidation of the antibiotic SMT by photolytic and photocatalytic processes using the immobilized TiO2 film was investigated in Milli-Q water and in SWW. It has been assumed that the radical pathway is a dominant mode of SMT degradation in the combined UV-C/H2O2 photolysis system. It has been shown that the presence of the TiO2 film in the UV-C/TiO2 photocatalysis system had a positive influence in terms of faster TOC reduction and faster degradation of formed degradation products. Among the investigated oxidation processes, the one including UV-C radiation combined with the addition of H2O2 and TiO2 catalyst was the most efficient. The UV-Abased processes have proven to be significantly less efficient in the SMT degradation. In total, five degradation products were identified. The results indicate that the cleavage of sulfonamide bond represents the main degradation pathway. The SMT photocatalytic degradation was considerably reduced in SWW due to the effects of different wastewater components; an almost ninefold reduction in the rate constant was observed. Electrical energy consumption per gram of demineralized sulfamethazine was about 2.7 times higher for the UV-A experiments than for the UV-C experiments, but the expenses for the UV-A experiments could be considerably lowered if the natural solar radiation (with its UV-A part) was used. Acknowledgements This study was partially supported by the University of Zagreb within the framework of the Short-Term Research Funding 2013-No. 2: BAdvanced Water Treatment Technologies.^

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