Photodegradation of Sulfadiazine by Goethite ... - ACS Publications

Mar 22, 2010 - Photodegradation of Sulfadiazine by Goethite-Oxalate Suspension under UV. Light Irradiation. Yan Wang,†,‡,§ Juan Boo Liang,*,† X...
0 downloads 0 Views 390KB Size
Download PDF
Ind. Eng. Chem. Res. 2010, 49, 3527–3532

3527

Photodegradation of Sulfadiazine by Goethite-Oxalate Suspension under UV Light Irradiation Yan Wang,†,‡,§ Juan Boo Liang,*,† Xin Di Liao,§ Lu-song Wang,| Teck Chwen Loh,⊥ Jun Dai,| and Yin Wan Ho† Institute of Bioscience and Department of Animal Science, UniVersiti Putra Malaysia, 43400 Serdang, Selangor, Malaysia, Guangdong Institute of Eco-EnVironmental and Soil Sciences, Guangzhou 510650, P.R. China, Key Laboratory of Ecological Agriculture of Ministry of Agriculture of the People’s Republic of China, and College of Nature Resources and EnVironment, South China Agriculture UniVersity, Guangzhou 510640, P.R. China

Sulfadiazine, a potent antibacterial agent belonging to the group of antibiotics called sulfonamides, has been reported to be present in surface and groundwater. This study investigated the degradation of sulfadiazine in a goethite (R-FeOOH)-oxalate Fenton-like system under UV irradiation. The results showed that sulfadiazine could be effectively photodegraded by the goethite-oxalate Fenton-like system as a result of the formation of the highly oxidizing hydroxyl radicals, •OH. Among the iron oxides tested (R-FeOOH, γ-Fe2O3, γ-FeOOH, and R-Fe2O3), R-FeOOH was found to be the most effective. Degradation of sulfadiazine depended significantly on the pH and initial concentration of oxalic acid in the system, with optimal values of 3.5 and 4.0 mM, respectively, under UV irradiation. Five intermediate products of sulfadiazine degradation were identified using high-performance liquid chromatography-mass spectrometry (HPLC-MS), gas chromatography-mass spectrometry (GC-MS), and ion chromatography (IC), and a possible sulfadiazine degradation pathway in such a system was proposed. Organic sulfur and organic nitrogen mineralization were also observed, and the results indicated that cleavage of the sulfonylurea bridge was easier than the other potential cleavage bonds under the goethite-oxalate system. In addition, results from Biolog assays suggested that the ecological toxicity of the sulfadiazine solution was effectively reduced after degradation. 1. Introduction Antibiotics have been extensively used in human and veterinary medicines to treat and prevent bacterial infections. Because of their low and incomplete absorption in human and animal bodies, antibiotics are excreted in urine or feces in their original form and/ or as their metabolites.1 Sulfadiazine [SD, 4-amino-N-(2-pyrimidinyl)benzenesulfonamide] is a potent antibacterial agent belonging to a large group of structurally related antibiotics, namely, sulfonamides. Sulfadiazine has been detected at varying concentrations, ranging between nanograms and milligrams per liter in the environment.2 With recent increasing concerns about the adverse effects of veterinary drug residues on public health, studies have been conducted on antibiotic residues in animal products (meat, eggs, and milk) and in the environment, including those in surface and groundwater.3 Although antibiotic residues are often present at trace levels in the environment, most of the current wastewater treatment techniques do not incorporate the degradation of antibiotic residues, and therefore, the continuous release of these hazardous compounds into the environment could result in antibiotic resistance and toxicity effects. Thus, the search for new alternatives to prevent water contamination by antibiotic residues is necessary, considering the risks that they pose for the environment and for human health.4 Although biodegradation remains the most important route for degrading organic pollutants in the environment, photodegradation could serve as an alternative elimination process for these compounds.5 Currently, the photo-Fenton process is * To whom correspondence should be addressed. Tel.: +6-0389472132. Fax: +6-03-89472101. E-mail: [email protected]. † Institute of Bioscience, Universiti Putra Malaysia. ‡ Guangdong Institute of Eco-Environmental and Soil Sciences. § Key Laboratory of Ecological Agriculture of Ministry of Agriculture of the People’s Republic of China. | South China Agriculture University. ⊥ Department of Animal Science, Universiti Putra Malaysia.

considered a good alternative because of its high efficiency in the generation of the highly oxidizing hydroxyl radicals, •OH, which can catalyze the oxidation of organic pollutants in the environment.6–9 It has been reported that iron oxide and oxalic acid have strong abilities to chelate with multivalent cations10 and can set up a Fenton-like system without additional H2O2 to photodegrade organic pollutants in aqueous solutions.11,12 Oxalic acid (a polycarboxylic acid) is widely exuded by plants in the natural environment, and iron oxides are natural minerals found in soils, rocks, lakes, rivers, air, and organisms, as well as on the seafloor.13 Major iron oxides include goethite (R-FeOOH), hematite (R-Fe2O3), maghemite (γ-Fe2O3), and lepidocrocite (γFeOOH). The band gaps (Eg) of R-FeOOH, γ-FeOOH, γ-Fe2O3, and R-Fe2O3 are 2.10, 2.06, 2.03, and 2.02 eV, respectively,14 and when excited by UV light, the electrons receive energy from photons and their activities are further promoted. The principle of the process can be described as follows11,15 iron oxide + hν f e- + h+

(1)

iron oxide + nH2C2O4 T [Fe(C2O4)n]

3-2n

(2)

[≡Fe(C2O4)n](2n-3)- + hν f [FeII(C2O4)(n-1)]4-2n + (C2O4)•(3) (C2O4)•- f CO2 + CO2•-

(4)

CO2•- + O2 f CO2 + O2-•

(5)

O2-• + 2H+ + Fe2+ f Fe3+ + H2O2

(6)

Fe2+ + H2O2 f Fe3+ + OH- + •OH

(7)

Some studies have been conducted on the photodegradation of antibiotic residues in wastewater using the photo-Fenton

10.1021/ie9014974  2010 American Chemical Society Published on Web 03/22/2010

3528

Ind. Eng. Chem. Res., Vol. 49, No. 8, 2010

process.16–18 However, compared to the photo-Fenton process, the Fe(III)-oxalate Fenton-like system is considered more suitable for the treatment of organic pollutants in a variety of wastewaters because of its low cost (without addition of H2O2) and the prevention of secondary pollution by iron ions. So far, there is little information on the use of the Fe(III)-oxalate Fenton-like system for the degradation of antibiotics. The present study uses sulfadiazine as the model antibiotic to investigate the degradation rate of antibiotics in the Fe(III)-oxalate complex system under UV irradiation. In addition, influencing factors, a possible pathway, and the change in ecological toxicity of such a system were investigated. 2. Materials and Methods 2.1. Chemicals. Sulfadiazine (99.0%) and methanol (HPLC grade) were purchased from Sigma-Aldrich (St. Louis, MO) and J&K Chemical Ltd. (Beijing, China), respectively. Other chemicals of analytical grade were purchased from Guangzhou Chemical Co., Guangzhou, Guangdong, China. All chemicals were used without further purification, and all solutions were prepared with ultrapure water (18.2 m) obtained using an ultrapure water system (EASYPure II RF/UV Barnstead, USA). 2.2. Preparation and Characterization of Iron Oxides. R-FeOOH (goethite) was prepared following the procedure of Schwertmamm and Cornell (1991),13 whereas γ-Fe2O3, γ-FeOOH, and R-Fe2O3 were prepared as reported by Wang et al. (2009).19 The X-ray powder diffraction patterns of the iron oxides were recorded on a Rigaku D/Max-III diffractometer at room temperature, operating at 30 kV and 30 mA, using Cu KR radiation (λ ) 0.15418 nm). The phases were identified by comparing diffraction patterns with those on the standard powder XRD cards compiled by the Joint Committee on Powder Diffraction Standards.20 The total surface areas and total pore volumes of the four samples were measured by the Brunauer-Emmett-Teller (BET) method using N2 adsorption at 77 K and a Carlo Erba Sorptometer.21 Pure R-FeOOH, γ-FeOOH, R-Fe2O3, and γ-Fe2O3 powders were obtained in accordance with their characteristic peaks (Figure S1, Supporting Information). The crystal sizes of γ-Fe2O3, γ-FeOOH, R-Fe2O3, and R-FeOOH were 43.2, 13.7, 54.7, and 41.9 nm, respectively, as determined using Sherrer’s formula with the corresponding strongest XRD peak. The specific surface areas of γ-Fe2O3, γ-FeOOH, R-Fe2O3, and R-FeOOH measured by the BET-BJH (Barrett-JoynerHalenda) method were 14.36, 115.44, 29.40, and 120.93 m2/g, respectively, and the total pore volumes were 0.05, 0.30, 0.27, and 0.16 cm3/g, respectively. 2.3. Experimental Procedure for Photodegradation Study. The photodegradation of sulfadiazine was carried out using a Pyrex cylindrical photoreactor, with an 8-W UV lamp (Luzchem Research Inc.) with its main emission at 350 nm positioned at the center of the reaction suspension. The photoreactor was surrounded by a Pyrex jacket with circulating water to control the temperature during the reaction period and covered with an aluminum foil to avoid indoor light irradiation. The reaction suspension was formed by adding 0.1 g of R-FeOOH powder to 250 mL of aqueous sulfadiazine solution with an initial concentration of 20 mg/L. The initial pH values of the suspensions in the photodegradation experiments were adjusted using NaOH and HClO4 solutions. Prior to the photoreaction, each suspension was stirred in the dark for 30 min using a magnetic stirrer to establish adsorption/desorption. During the photoreaction process, the aqueous suspension was irradiated with UV light, under constant aeration and stirring. At desig-

nated time intervals (0, 2, 5, 10, 15, 20, and 40 min), samples were withdrawn from the suspension, immediately centrifuged for 28 min at 4950 rpm in the dark, and then stored for further analysis. 2.4. Analysis of Sulfadiazine and Its Intermediate Compounds. Sulfadiazine was determined by high-performance liquid chromatography (HPLC), using an instrument with a Waters 1525 binary pump and a Waters 2487 dual λ absorbance UV/vis detector at 254 nm. The analytical reverse-phase column was a 5 µm Symmetry-C18, 4.6-mm i.d. × 25-cm length (Waters, Milford, MA). A methanol/1% (v/v) acetic acid mixture (25:75) was used at a mobile flow rate of 1.0 mL/min under isocratic conditions at room temperature. Samples of 20 µL were injected into the column through the sampling loop for analysis. The data were processed with Breeze software (Waters, Milford, MA), and the relative standard deviation for HPLC analysis was controlled to within 5%. The intermediate products of sulfadiazine degradation were analyzed by HPLC-mass spectrometry (MS) with a hypersil ODS2 column (250 mm × 4.0 mm, 5 µm), gas chromatography-mass spectrometry (GC-MS), and ion chromatography (IC). The operating parameters of the HPLC-MS analyses were the same as described above for HPLC. The gas chromatograph was equipped with a split/splitless injector, and a TRX-1MS column (30 m, 0.25-mm id, 0.25-µm film thickness) was used for separation. The initial temperature was set at 55 °C and was held for 2 min in splitless mode. When the splitter opened, the oven was heated at a heating rate of 15 °C/min to 180 °C and then at a rate of 10 °C/min to 260 °C, where it was held for 12 min. The solvent delay time was set for 4 min, and the temperature of the transfer line was set at 250 °C. Mass spectra were recorded at three scans under electron impact of 70 eV with a mass range of 50-350 amu. The concentrations of oxalic acid; released SO42-, NO3-, and HCOOH; and CH3COOH were determined with an IC analyzer (Dionex ICS-90, Sunnyvale, CA) equipped with a polysulfonate ion-exclusion column and an organic acid column. The eluent of the ion analytical column contained 8 mM Na2CO3, 1 mM NaHCO3, and 45 mM sulfate. The eluent of the organic acid column contained 0.4 mM HCl and 5 mM tetrabutylammonium hydroxide. The above procedure was performed under a mobile flow rate of 1.0 mL/min at 1700-1800 psi and room temperature. 2.5. Determination of Ecological Toxicity. Red soil was sampled at a depth of 0-20 cm from a field located at the Guangdong Institute of Eco-Environmental and Soil Science, Guangzhou, China. The collected soil was air-dried at 25 °C and sieved through a 2-mm screen to remove stones, root fragments, and other impurities. The sieved soil was later adjusted to 50% water-holding capacity (WHC) by the weighing method and incubated in a polythene bottle at 25 °C for 7 days prior to the soil microbial functional diversity analysis. The functional diversity of the soil microbial community was analyzed by developing community-level physiological profiles using Biolog assays. Ten grams of soil sample was mixed in 100 mL of sterile saline solution (0.85%, m/v) with 5 g of 3-mm glass beads on a rotary shaker at 200 rpm for 30 min at 25 °C. After being diluted 100-fold and allowed to settle for 10 min, 10 mL of each dilution was mixed with 90 mL of the sulfadiazine solution after 0, 20, and 40 min of the degradation reaction. Sterile distilled water was used as the control. Then, 150 µL of the suspension mixture was added to each well of a Biolog-ECO plate. The plates were then incubated at 25 °C in the dark and recorded every 24 h for 7 days. The light absorbance at 590 nm of each well was recorded using a Biolog

Ind. Eng. Chem. Res., Vol. 49, No. 8, 2010

Figure 1. Dependence of SD photodegradation on 0.4 g/L of different types of iron oxides in the presence of 4 mM initial concentration of oxalic acid (C0ox) under UV light irradiation. C0 is the initial concentration of SD, and Ct is the SD concentration at time t (min).

automated Biolog Microplate Reader (Biolog, Hayward, CA), and the data were collected by Microlog 4.01 software (Biolog, Hayward, CA). Average well color development (AWCD) was obtained from the average optical density of all of the wells in each plate. 3. Results and Discussion 3.1. Photodegradation of Sulfadiazine. 3.1.1. Selection of the Iron Oxide. The type of iron oxide could considerably influence the photodegradation of organic compounds by the iron oxide and oxalic acid Fenton-like system. For example, we recently reported that R-FeOOH is more efficient than γ-FeOOH, R-Fe2O3, and γ-Fe2O3 for the degradation of pyrene by photo-Fenton-like degradation in solid phase.19 To evaluate the influence of different iron oxides (R-FeOOH, γ-FeOOH, R-Fe2O3, and γ-Fe2O3) on Fenton-like degradation of sulfadiazine, a series of preliminary experiments was carried out to compare their influence on sulfadiazine oxidation under UV irradiation using a fixed concentration of iron oxide (0.4 g/L), initial sulfadiazine concentration (20 mg/L), initial oxalic acid concentration (4.0 mM), and initial pH value (3.5). The results (Figure 1) showed that all four iron oxides had high photocatalytic activities to degrade sulfadiazine, with R-FeOOH being the most active. A removal of 97% of sulfadiazine was achieved by R-FeOOH after 40 min of irradiation, as compared to 81% for γ-FeOOH, 48% for R-Fe2O3, and 83% for γ-Fe2O3 after the same irradiation time. To quantify the photodegradation rate of sulfadiazine, the kinetics was simulated, and the fitted results showed that sulfadiazine degradation could be described using a first-order model. The first-order kinetic constants (k) for the degradation were found to be 0.08394 min-1 (R ) 0.904), 0.05076 min-1 (R ) 0.936), 0.01233 min-1 (R ) 0.998), and 0.04099 min-1 (R ) 0.993) in the presence of R-FeOOH, γ-FeOOH, R-Fe2O3, and γ-Fe2O3, respectively. The half-lives (t1/2) of sulfadiazine degraded by R-FeOOH, γ-FeOOH, R-Fe2O3, and γ-Fe2O3 were determined as 5.96, 9.85, 40.55, and 12.20 min, respectively. These results demonstrate that R-FeOOH has the highest photocatalytic activity for degrading organic pollutants in the photo-Fenton-like degradation in this study. The high efficiency of R-FeOOH in degrading sulfadiazine can be attributed to its basic morphology, namely, an acicular crystal structure, and different BET areas to generate a high quantum yield of Fe(II).19,22 R-FeOOH was thus selected as the model iron oxide to study the photodegradation of sulfadiazine because of its high efficiency and abundance in nature.

3529

Figure 2. Photocatalytic SD photodegradation with UV light only and 0.4 g/L R-FeOOH in the presence of 0-5 mM initial concentrations of oxalic acid (Ox) at pH 3.5 under UV light irradiation. The inset presents the dependence of the first-order rate constant (k) for SD degradation on the value of C0ox.

3.1.2. Influence of Initial Oxalic Acid Concentration. The results of sulfadiazine photodegradation at a fixed initial sulfadiazine concentration of 20 mg/L, a pH of 3.5, and a reaction temperature of 30 °C under UV irradiation only, under UV irradiation with R-FeOOH, and at various initial oxalic acid concentrations are presented in Figure 2. It was found that, under UV light alone, only about 3% of the sulfadiazine was photodegraded after 40 min, whereas in the presence of UV light and R-FeOOH, 15% of sulfadiazine was removed. However, in the presence of UV together with R-FeOOH and 4.0 mM initial oxalic acid, the extent of sulfadiazine photodegradation increased tremendously to 93%. These results clearly followed a first-order kinetic model with constants (k) of 3.351 × 10-4 (R ) 0.930), 0.00419 (R ) 0.981), and 0.08394 (R ) 0.904) for UV only, UV with R-FeOOH, and UV with R-FeOOH plus 4.0 mM initial oxalic acid, respectively. The results also demonstrated that the oxalic acid concentration had an obvious effect on sulfadiazine degradation. When the concentration of oxalic acid increased to 1, 2, 3, and 4 mM, the corresponding k values increased to 0.04523, 0.0593, 0.07567, and 0.08394 min-1, respectively. These results indicate that, in combination with oxalate, the photodegradation of sulfadiazine by iron oxide is greatly accelerated. In the absence of oxalic acid, iron oxide merely acts as a photocatalyst, but in the presence of oxalic acid, the Fe(III) is mainly present as Fe(C2O4)2- and Fe(C2O4)33- in the Fe(III)-oxalate system when the concentration of oxalate is more than 0.18 mM.23 The rate of production of hydroxyl radicals by the photolysis of the Fe(III)-oxalate complex system is higher than that for the Fe(III) system.23,24 Thus, the presence of oxalic acid accelerates the degradation of organic pollutants.15,25 However, the degradation constant decreased to 0.06762 min-1 at higher oxalic acid content (5 mM). It has been reported that excess oxalate can occupy the adsorption sites on the surface of iron oxide and react competitively for the hydroxyl radicals, leading to increased formation of Fe(III), which would inhibit the formation of H2O2 and, therefore, inhibit the photodegradation of organic pollutants.26,28 The present results (Figure 2) demonstrate that 4.0 mM is the optimal initial concentration of oxalic acid for the photodegradation of sulfadiazine by R-FeOOH in the aqueous phase. During the photodegradation of sulfadiazine under UV irradiation with R-FeOOH and oxalic acid, the concentration of oxalic acid was reduced through the reaction of R-FeOOH and oxalic acid under UV light as described in eqs 2-5, and the pH of the reaction system increased rapidly (Figure S2,

3530

Ind. Eng. Chem. Res., Vol. 49, No. 8, 2010 Table 1

Figure 3. Dependence of SD photodegradation on initial pH with 0.4 g/L R-FeOOH in the presence of 4 mM C0ox under UV light irradiation. The inset presents the dependence of the first-order rate constant (k) for SD degradation on the initial pH value.

Supporting Information). These results indicate that sulfadiazine can be effectively photodegraded by the goethite-oxalate Fenton-like system as a result of the formation of •OH, as we reported previously.19 Because geothite, oxalate, and light are abundant in nature, the results of the present study provide a better understanding of the natural degradation of sulfadiazine, which could further enhance the reaction efficiency of degradation of organic pollutants, including sulfadiazine in the environment. 3.1.3. Influence of pH. pH has been identified as another important factor in the photo-Fenton-like degradation of organic pollutants.25–27 The present results show that sulfadiazine photolysis follows first-order kinetics at different initial pH values, and the constants (k) at initial pH values of 2.5, 3.0, 3.5, 4.0, and 4.5 containing 20 mM sulfadiazine were found to be 0.03503, 0.06512, 0.08394, 0.03151, and 0.01873 min-1, respectively (Figure 3). These results indicate that pH 3.5 is the optimal pH for the photodegradation of sulfadiazine, whereas higher or lower pH would cause the photooxidation efficiency to decrease. Zuo and Hoigne´ previously reported that pH affects the competition between the reactions of active intermediates,28 HO2/O2- and Fe(III)/Fe(II), which, in turn, affect the formation rate of H2O2 and •OH radicals in the Fe(III)-oxalate complex system. In this study, when the pH was higher than 3.5, Fe(III) was present as Fe(OH)2+, and Fe(OH)2+ has been found to photolyze with much lower quantum yields than Fe(III)-oxalate complexes.28 Zhou et al. reported that, when the pH was further increased to 6.0,25 the photooxidation efficiency hardly changed as compared to that at pH 5.0; however, higher pH values would cause unstable ferric ions to produce colloids and precipitates. Therefore, pH values higher than 5 were not included in this study. When the pH is lower than 3.5, the excess H+ ions might affect the generation of Fe(II) and, thus, reduce the concentration of hydroxyl radicals (eq 6), which might influence the photooxidation rate in the solution.25 Based on the above findings, we concluded that 3.5 was the optimal pH for the photodegradation of sulfadiazine by the goethite-oxalate complex system. 3.2. Photooxidation Pathway and Mineralization of Sulfadiazine. Five intermediate products generated during the photodegradation of sulfadiazine were identified in the 10- and 40-min-irradiated solutions using HPLC-MS, GC-MS, and IC (Table 1). As shown in Figure S3 (Supporting Information), two intermediate products were detected by HPLC and had higher hydrophilicities than the mother compound, sulfadiazine. The main peaks were intensified using LC-MS, and the typical chromatograms of 4-OH-sulfadiazine and 4-[2-iminopyrimidine-

1(2H)-yl]aniline sulfadiazine were detected at m/z 267 and 187 (Figure S4, Supporting Information). The other three intermediate products were 2-aminopyrimidine (detected by GC-MS; Figure S5, Supporting Information) and HCOOH and CH3COOH (both detected by IC; Figure S6, Supporting Information). Based on the results of this study and those in the literature, the proposed pathways for the photooxidation of sulfadiazine induced by Fe(III)-oxalate complexes are shown in Figure 4. According to Boreen et al.,29 the potential photolysis/cleavage sites of sulfonamides are as follows:

The observation of 2-aminopyrimidine in this study confirmed the cleavage of the δ bond. The presence of 4-[2-iminopyrimidine-1(2H)-yl]aniline, detected by HPLC-MS in this study, suggested the cleavage of the γ bond, resulting in the formation of sulfanilic acid,29 which reacted with 2-aminopyrimidine to form 4-[2-iminopyrimidine-1(2H)-yl]aniline. However, the R, β, and ε bonds appeared to be more difficult to cleave compared to the γ and δ bonds in the photodegradation of sulfadiazine by Fe(III)-oxalate complexes because no related intermediate products, such as aniline, hydroquinone, and sulfanilamide, were observed in this study. Normally, carbon dioxide (CO2) and total organic carbon (TOC) are used to determine the mineralization of antibiotics.30 However, because of the presence of oxalic acid (containing organic C in its structure) in the degradation system, TOC cannot be used to determine the sulfadiazine mineralization in this study. On the other hand, elements S and N as organic S and organic N are part of the chemical structure of sulfadiazine, and when sulfadiazine is mineralized, the organic S and organic N will be oxidized to SO42- and NO3-, respectively. Based on the above assumptions, the total mineralization of sulfadiazine was traced using the formation of nitrate and sulfate ions detected using IC in this study. The results showed an increase in the concentration of sulfate ions in solution, with 0.075 mM of SO42- detected after 40 min of irradiation (Figure 5). Based on its molecular structure, the 20 mg/kg of initial sulfadiazine (molecular weight ) 250) used should contain 0.08 mM organic S, and taking into account the quantity of SO42- formed, the results indicated 94% mineralization of organic S after 40 min of irradiation. However, the sulfadiazine degradation shown in Figure 2 was 97%. This discrepancy seems to suggest that the

Ind. Eng. Chem. Res., Vol. 49, No. 8, 2010

3531

Figure 4. Possible pathways for SD photodegradation in the goethite/oxalate suspension with 0.4 g/L R-FeOOH in the presence of 4 mM C0ox at pH 3.5 under UV light irradiation.

Figure 5. Concentrations of NO3- and SO42- produced during SD photodegradation with 0.4 g/L R-FeOOH in the presence of 4 mM C0ox at pH 3.5 under UV light irradiation.

mineralization reaction could continue in the intermediate products after the 40-min experimental period. In contrast, the mineralization of organic N was much lower, as only 26% (0.084 mM out of the total organic N concentration of 0.32 mM from the 20 mg/kg initial concentration of sulfadiazine) was mineralized. These results indicate an important photocatalytic cleavage of the sulfonylurea bridge; that is, in the degradation pathway, the δ and γ bonds are easier to cleave than the β and ε bonds. The fact that the structure of sulfadiazine consists of two organic N atoms in the ring increases the difficulty for mineralization of organic N because of the stability of the ring.31 3.3. Ecological Toxicity of Sulfadiazine during Photooxidation. Antibiotics can have a negative effect on the soil microbial community. The Biolog-ECO plate has been used to investigate the ecotoxicity of antibiotics on the functional diversity of soil microbial communities.32 As mentioned previously (section 3.2), SD was degraded, and five intermediate products were detected in this study, suggesting that the toxicity of sulfadiazine solution during photooxidation might be altered. Thus, it is essential to assess the impact of such changes on the environment. The Biolog-ECO plates were used to evaluate the effect of sulfadiazine samples after 0, 20, and 40 min of reaction with R-FeOOH on the functional diversity of soil microbial community, and AWCD was used as an indicator of microbial activity. The results are shown in Figure 6. AWCD in the control (sterile water only) increased rapidly after 48 h and achieved 0.97 within 144 h of incubation, but in the treatment without the photodegradation of sulfadiazine, AWCD increased slowly

Figure 6. AWCD of 95 carbon source after different photodegradation times with 0.4 g/L R-FeOOH in the presence of 4 mM C0ox at 20 mg/kg initial SD concentration and pH 3.5 under UV light irradiation. The bars represent the standard error of the means (n ) 3).

to 0.27 within 144 h, suggesting that the soil microbial community was sensitive to sulfadiazine. After 20 and 40 min of photodegradation of sulfadiazine, AWCD increased to 0.52 and 0.90, respectively. These results demonstrate that the soil microbial inhibition was lower after UV illumination than before, and the ecotoxicity of the sulfadiazine solution decreased with the progress of photodegradation. Baran et al. reported that all degradation products were less toxic than the mother sulfonamides during the biodegradation process.33 The results of this study thus suggest that the ecotoxicity of post-photodegraded sulfadiazine was reduced because of the decrease in the concentration of the mother sulfonamides during the photodegradation progress under the Fe(III)-oxalate system. 4. Conclusions In conclusion, the results of this study demonstrate that sulfadiazine can be efficiently degraded by the Fe(III)-oxalate system under UV irradiation. Photodegradation of sulfadiazine was found to depend strongly on the type of iron oxide, as well as the initial concentration of oxalic acid and pH. Goethite had the highest photochemical activity compared to the other iron oxides tested. The optimal initial oxalic acid concentration and pH for the photodegradation reaction were found to be 4.0 mM and 3.5, respectively. Five intermediate products and organic N and organic S mineralization were detected. The results also indicated that cleavage of the sulfonylurea bridge was easier than other potential cleavage points under the goethite-oxalate system. The ecotoxicity of the intermediate products was found

3532

Ind. Eng. Chem. Res., Vol. 49, No. 8, 2010

to be less than that of the mother sulfadiazine compound, thus suggesting a positive effect of photodegradation on the impact of organic pollutants such as sulfadiazine on the environment. These results can help to better understand the natural degradation of sulfadiazine and offer the possibility of treating effluents, including livestock wastewater, contaminated with antibiotics at low cost because goethite, oxalate, and light are available in the natural environment. Acknowledgment This study was supported by the International Foundation for Science (IFS) (Grant C/4079-1), the National Natural Science Foundation of China (Project 30900205), and the ScienceFund of Malaysia (Project 05-01-04-SF0790). Supporting Information Available: Detailed information as described in the text. This material is available free of charge via the Internet at http://pubs.acs.org. Literature Cited (1) Boxall, A. B. A.; Fogg, L. A.; Blackwell, P. A.; Kay, P.; Pemberton, E. J. Veterinary medicines in the environment. ReV. EnViron. Contam. Toxicol. 2004, 180, 1. (2) Sukul, P.; Spiteller, M. Sulfonamides in the environment as veterinary drugs: Present status and future scopessA review. ReV. EnViron. Contam. Toxicol. 2006, 187, 67. (3) Hirsch, R.; Ternes, T.; Haberer, K. Occurrence of antibiotics in the aquatic environment. Sci. Total EnViron. 1999, 225, 109. (4) Seiler, J. P. Pharmacodynamic activity of drugs and ecotoxicologyscan the two be connected. Toxicol. Lett. 2002, 131, 105. (5) Kot-Wasik, A.; Da¸browska, D.; Namies´nik, J. Photodegradation and biodegradation study of benzo(a)pyrene in different liquid media. J. Photochem. Photobiol. A 2004, 168, 109. (6) Li, F. B.; Li, X. Z.; Li, X. M.; Liu, T. X. Heterogeneous photodegradation of bisphenol A in the interface of iron oxides together with oxalate aqueous solution. J. Colloid Interface Sci. 2007, 311, 481. (7) Li, F. B.; Li, X. Z.; Liu, C. S.; Li, X. M.; Liu, T. X. Effect of Oxalate on Photodegradation of Bisphenol A at the Interface of Different Iron Oxides. Ind. Eng. Chem. Res. 2007, 46, 781. (8) Liu, C.; Zhang, L.; Li, F.; Wang, Y.; Gao, Y.; Li, X.; Cao, W.; Feng, C.; Dong, J.; Sun, L. Dependence of Sulfadiazine Oxidative Degradation on Physicochemical Properties of Manganese Dioxides. Ind. Eng. Chem. Res. 2009. (9) Yu, J. G.; Yu, J. C.; Leung, M. K. P.; Ho, W. K.; Cheng, B.; Zhao, X. J.; Zhao, J. C. Effects of acidic and basic hydrolysis catalysts on the photocatalytic activity and microstructures of bimodal mesoporous titania. J. Catal. 2003, 217, 69. (10) Kayashima, T.; Katayama, T. Oxalic acid is available as a natural antioxidant in some systems. Biochim. Biophys. Acta 2002, 1573, 1. (11) Faust, B. C.; Allen, J. Photochemistry of aqueous iron (III)polycarboxylate complexes: Roles in the chemistry of atmospheric and surface waters. EnViron. Sci. Technol. 1993, 27, 2517. (12) Siffert, C.; Sulzberger, B. Light-induced dissolution of hematite in the presence of oxalate. A case study. Langmuir 1991, 7, 1627. (13) Schwertmann, U.; Cornell, R. M. Iron Oxides in the Laboratory: Preparation and Characterization; VCH: New York, 1991. (14) Leland, J. K.; Bard, A. J. Photochemistry of colloidal semiconducting iron oxide polymorphs. J. Phys. Chem. 1987, 91, 5076.

(15) Zuo, Y. G.; Deng, Y. W. Iron(II)-catalyzed photochemical decomposition of oxalic acid and generation of H2O2 in atmospheric liquid phases. Chemosphere 1997, 35, 2051. (16) Bautitz, I. R.; Nogueira, R. F. P. Degradation of tetracycline by photo-fenton processsSolar irradiation and matrix effects. J. Photochem. Photobiol. A 2007, 187, 33. (17) Gartiser, S.; Urich, E.; Alexy, R.; Kummerer, K. Ultimate biodegradation and elimination of antibiotics in inherent tests. Chemosphere 2007, 67, 604. (18) Reyes, C.; Fernandez, J.; Freer, J.; Mondaca, M. A.; Zaror, C.; Malato, S.; Mansilla, H. D. Degradation and inactivation of tetracycline by TiO2 photocatalysis. J. Photochem. Photobiol. A 2006, 184, 141. (19) Wang, Y.; Liu, C. S.; Li, F. B.; Liu, C. P.; Liang, J. B. Photodegradation of polycyclic aromatic hydrocarbon pyrene by iron oxide in solid phase. J. Hazard. Mater. 2009, 162, 716. (20) Joint Committee for Powder Diffraction Standards, JCPDS, International Centre for Diffraction Data: Newtown Square, PA, 1979. (21) Yu, Y.; Yu, J. C.; Yu, J. G.; Kwok, Y. C.; Che, Y. K.; Zhao, J. C.; Ding, L.; Ge, W. K.; Wong, P. K. Enhancement of photocatalytic activity of mesoporous TiO2 by using carbon nanotubes. Appl. Catal. A 2005, 289, 186. (22) Trovo´, A. G.; Melo, S. A. S.; Nogueira, R. F. P. Photodegradation of the pharmaceuticals amoxicillin, bezafibrate and paracetamol by the photo-Fenton processsApplication to sewage treatment plant effluent. J. Photochem. Photobiol. A 2008, 198, 215. (23) Balmer, M. E.; Sulzberger, B. Atrazine degradation in irradiated iron/oxalate system: Effect on pH and oxalate. EnViron. Sci. Technol. 1999, 33, 2418. (24) Zepp, R. G.; Faust, B. C.; Hoigne′, J. Hydroxyl radical formation in aqueous reactions (pH 3-8) of iron(II) with hydrogen peroxide: The photo-Fenton reaction. EnViron. Sci. Technol. 1992, 26, 313. (25) Zhou, D.; Wu, F.; Deng, N. S. Fe(III)-oxalate complexes induced photooxidation of diethylstilbestrol in water. Chemosphere 2004, 57, 283. (26) Lei, J.; Liu, C. S.; Li, F. B.; Li, X. M.; Zhou, S. G.; Liu, T. X.; Gu, M. H.; Wu, Q. T. Photodegradation of orange I in the heterogeneous iron oxide-oxalate complex system under UVA irradiation. J. Hazard. Mater. 2006, 37, 1016. (27) Liu, C. S.; Li, F. B.; Li, X. M.; Zhang, G.; Kuang, Y. Q. The effect of iron oxides and oxalate on the photodegradation of 2-mercaptobenzothiazole. J. Mol. Catal. A 2006, 252, 40. (28) Zuo, Y.; Hoigne´, J. Formation of hydrogen peroxide and depletion of oxalic acid in atmospheric water by photolysis of iron(III)-oxalate complexes. EnViron. Sci. Technol. 1992, 26, 1014. (29) Boreen, A. L.; Arnold, X. A.; McNeill, K. Photochemical fate of sulfa drugs in the aquatic environment: Sulfa drugs containing fivemembered heterocyclic groups. EnViron. Sci. Technol. 2004, 38, 3933. (30) Garcı´a-Gala´n, M. J.; Dı´az-Cruz, M. S.; Barcelo´, D. Identification and determination of metabolites and degradation products of sulfonamide antibiotics. Trends Anal. Chem. 1999, 27, 1008. (31) Lan, Q.; Li, F.; Liu, C.; Li, X. Heterogeneous Photodegradation of Pentachlorophenol with Maghemite and Oxalate under UV Illumination. EnViron. Sci. Technol. 2008, 42, 7918. (32) Kong, W. D.; Zhu, Y. G.; Fu, B. J.; Marschner, P.; He, J. Z. The veterinary antibiotic oxytetracycline and Cu influence functional diversity of the soil microbial community. EnViron. Pollut. 2006, 143, 129. (33) Baran, W.; Sochacka, J.; Wardas, W. Toxicity and biodegradability of sulfonamides and products of their photocatalytic degradation in aqueous solutions. Chemosphere 2006, 65, 1295.

ReceiVed for reView September 23, 2009 ReVised manuscript receiVed February 25, 2010 Accepted February 26, 2010 IE9014974