Photocatalytic degradation of ciprofloxacin using mono- (Au, Ag and Cu) and bi- (Au–Ag and Au–Cu) metallic nanoparticles supported on TiO2 under UV-C and simulated sunlight

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Photocatalytic degradation of ciprofloxacin using mono- (Au, Ag and Cu) and bi- (Au–Ag and Au–Cu) metallic nanoparticles supported on TiO2 under UV-C and simulated sunlight Juan C. Durán-Álvarez a , Edwin Avella a , Rosa María Ramírez-Zamora b , Rodolfo Zanella a,∗ a Centro de Ciencias Aplicadas y Desarrollo Tecnológico, Universidad Nacional Autónoma de México, Circuito Exterior S/N, Ciudad Universitaria, P.O. Box 70-186, Coyoacán 04510, México, D.F., Mexico b Instituto de Ingeniería, Universidad Nacional Autónoma de México, Avenida Universidad 3000, Coyoacán, México, D.F., Mexico

a r t i c l e

i n f o

Article history: Received 2 May 2015 Received in revised form 2 July 2015 Accepted 17 July 2015 Available online xxx Keywords: Antibiotics By-products Metallic nanoparticles Photocatalysis Toxicity

a b s t r a c t Heterogeneous photocatalysis using TiO2 can effectively remove antibiotics from water using UV light; however, its performance is notably reduced under sunlight irradiation. The use of metallic nanoparticles deposited on TiO2 may result in the photo-activation of the catalyst within the visible spectrum. In this study mono- (Au, Ag and Cu) and bi-metallic Au–Ag and Au–Cu nanoparticles were deposited on TiO2 to photocatalytically degrade the antibiotic ciprofloxacin in pure water using either UV-C or simulated sunlight. The optimal loading of mono-metallic nanoparticles on TiO2 was determined as 1.5 wt.% for Au and Ag, and 1.0 wt.% for Cu; first order degradation rates (kapp ) of 0.06, 0.117 and 0.072 min–1 , respectively, were determined for these materials. In UV-C tests, the complete degradation of ciprofloxacin was achieved upon 90 min of irradiation, whilst complete mineralization was reached in <180 min for all of the tested catalysts. In simulated sunlight photocatalysis, ciprofloxacin was only partially removed upon 360 min of irradiation when using mono-metallic materials, while complete mineralization was achieved when bi-metallic nanoparticles on TiO2 were tested. A group of by-products were identified and degradation paths were elucidated for photolysis and photocatalysis. Toxicity tests using V. fischeri showed the non-toxicity of the by-products remaining after 360 min of simulated sunlight irradiation. Even though toxicity was low, ciprofloxacin by-products showed some residual antibiotic activity. No catalyst deactivation was observed after 3 consecutive reaction cycles. © 2015 Elsevier B.V. All rights reserved.

1. Introduction The occurrence of antibiotics in wastewater and their potential to reach drinking water brings concern to scientists and water policy makers. Antibiotics are mostly synthetic compounds designed to cause biological effects in specific organisms, thus when these substances get into the environment, unpredictable effects may be caused to non-target organisms, such as genotoxicity in aquatic organisms [1–3] and expression of antibiotic resistance in certain microorganisms, some of them pathogenic for humans [4,5]. Wastewater treatment schemes have attempted to completely remove antibiotics; however, most of these pharmaceuticals are only partially removed by primary and secondary treatment systems, and concentrations of tens of ng/L persist in effluents [6,7]. Advanced oxidation processes may be a good alternative to remove

∗ Corresponding author. Tel.: +52 55 56228601; fax: +52 55 55500654. E-mail address: [email protected] (R. Zanella).

antibiotics from water [8], although several by-products persist upon long reaction periods. Some of these by-products may be more genotoxic than parent compounds as well as maintain antibiotic activity when active moieties in molecules remain unchanged [9,10]. Heterogeneous photocatalysis has shown to remove and mineralize some second and third generation antibiotics such as fluoroquinolones [11], cephalosporins [12,13] and vancomycins [14]. High mineralization is achieved by the continuous generation of reactive oxygen substances (ROS) upon the irradiation of the photocatalyst. Reactive oxygen substances jointly with the photo-produced holes on the semiconductor surface can oxidize both the target antibiotics and their by-products [15]. Titanium dioxide is the semiconductor most commonly used as photocatalyst since it is non-toxic, unexpensive, efficient and easy handling [15,16]. Nonetheless, due to its high band gap value (3.2 eV), TiO2 can be only activated as photocatalyst at  < 400 nm, i.e., UVA light [15,17], which limits its use in solar plants (<5% of the solar spectrum corresponds to UV-A light). In this respect, deposition of well-dispersed metallic nanoparticles (2–20 nm) on the

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Please cite this article in press as: J.C. Durán-Álvarez, et al., Photocatalytic degradation of ciprofloxacin using mono- (Au, Ag and Cu) and bi- (Au–Ag and Au–Cu) metallic nanoparticles supported on TiO2 under UV-C and simulated sunlight, Catal. Today (2015), http://dx.doi.org/10.1016/j.cattod.2015.07.033

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semiconductor surface has shown favorable results under sunlight conditions. Metallic nanoparticles (Au, Ag, Pt, Pd, Cu and Ni) act as a major component for harvesting visible light due to the surface plasmon resonance (SPR) effect [18]. Additionally, nanoparticles act as electron traps due to the differences in the Fermi levels, dragging the photo-produced electrons in the conduction band and thus diminishing recombination [19]. The captured electrons in the metallic nanoparticle can chemically reduce the molecular oxygen (O2 ), producing ROS such as superoxide (• O2 – ) and hydroxyl (• OH) radicals [18]. Surface modified semiconductors with metallic nanoparticles have been used to degrade antibiotic compounds in water, showing auspicious results. Oros-Ruiz et al. [20] reported the complete degradation and 80% of mineralization of trimethoprim upon 6 h of UV-C light irradiation using TiO2 modified with Au, Ag, Cu and Ni nanoparticles. Complete degradation and 90% of mineralization of chloramphenicol was achieved after 18 min of UV-C light irradiation when (1 wt.%) Ag/TiO2 was tested [21]. Pugazhenthiran et al. [12,13] used Au and Ag nanoparticles deposited on TiO2 nanotubes to degrade ceftiofur sodium, obtaining the complete degradation and 50% of mineralization upon 90 min and 10 h of UV–visible–NIR (360–2000 nm) irradiation, respectively. Chlortetracycline was effectively degraded (80% upon 90 min irradiation) using bio-inspired hollow spherical Cu/TiO2 materials and simulated sunlight [22]. Such results have been attributed to the high production of • OH radicals on the surface of the metallic nanoparticles [23]. On the other hand, deposition of bi-metallic nanoparticles on TiO2 has been extensively reported for selective oxidation/reduction of gas streams [24] but barely in water purification. Reduction of NO3 − into N2 and Cr6+ into Cr+3 in natural waters using Pd–Cu/TiO2 and Au–Cu/TiO2 , respectively have been recently reported [25,26], while efficient oxidation of phenol was observed in wastewater using Au–Cu alloys deposited on TiO2 [27]. Bi-metallic alloys display additional advantages compared to their mono-metallic counterparts, such as increased photocatalytic activity and selectivity [25,27–29], as well as the decrease in the deactivation process [24]. Oxidation of Ag, Ni and Cu nanoparticles during photocatalytic process can be hindered when Au is used in the alloy, since gold is less oxophilic than other metals [30,31]. Moreover, it is reported that carbonaceous species are efficiently adsorbed onto Au nanoparticles in bi-metallic alloys, while the second metal is able to convert dissolved O2 into ROS. Resulting in higher oxidation rates compared with mono-metallic materials [29,32,33]. Gold is miscible with Ag and Cu, thus Au–Ag and Au–Cu alloyed nanoparticles can be synthesized [34]. Au and Ag are miscible in all the proportions, while Au–Cu is less miscible due to the differences in their lattice constants and atomic radii [35,36]. For the best of our knowledge, this is the first study using bi-metallic nanoparticles deposited on TiO2 for the removal of antibiotics in water. The aim of this work was to evaluate the impact of the superficial modification of TiO2 by deposition of mono- (Au, Ag and Cu) and bi-metallic (Au–Ag and Au–Cu) nanoparticles on the photocatalytic degradation and mineralization of ciprofloxacin using either UV-C light (254 nm) or simulated sunlight (290–800 nm). Additionally, degradation mechanisms were elucidated and the toxicity of the by-products was determined in effluents when mineralization was not completed at the end of the experiments.

2. Materials and methods 2.1. Catalysts preparation The TiO2 was synthesized by acidic hydrolysis through the sol–gel procedure. To obtain 5 g of TiO2 , 23 mL of titanium(IV)

butoxide was dissolved in 26 mL of 1-butanol (Sigma–Aldrich). The solution was sonicated at 139 kHz for 5 min. Subsequently, the solution was stabilized at 70 ◦ C and 9 mL of tridistilled water, adjusted at pH 3 with HNO3 , were slowly added to promote the hydrolysis. The reaction mixture was stirred for 24 h; then, the gel was dried at 100 ◦ C for 24 h and grounded to obtain a fine powder. Anatase formation was promoted by annealing at 500 ◦ C for 3 h using a temperature ramp of 2 ◦ C/min. The preparation of Au/TiO2 and Cu/TiO2 materials was carried out via the deposition–precipitation method using urea as basifying agent [37]. For this, the corresponding amount of Au and Cu precursors (HAuCl4 ·3H2 O and Cu(NO3 )2 ·2.5H2 O, respectively) were dissolved in tridistilled water to achieve the optimum concentration of 4.2 × 10–3 M [37]; then urea (0.42 M) was added to the solution followed immediately by TiO2 . The suspension was vigorously stirred and the temperature was kept constant at 80 ◦ C for 16 h. The Ag/TiO2 catalyst was obtained by deposition–precipitation using NaOH as basifying agent. In a typical preparation, 1 g of TiO2 and 4.2 × 10–3 M of AgNO3 solution were mixed. The pH of the suspension was raised up from ∼3 to 9 by addition of 0.05 M NaOH. Then the suspension was stirred for 4 h at 80 ◦ C. After deposition, all of the materials were washed four times with distilled water (100 mL of water for each gram of material) and dried under vacuum at 80 ◦ C for 2.5 h. The mono- and bimetallic nanoparticles were formed by thermal treatment at 350 ◦ C for 3 h under H2 flow (1 mL of H2 /mg of powder); a heating rate of 2 ◦ C/min was used to achieve this temperature. The same thermal treatment has been previously reported by our group [20]. In order to determine the optimal loading of metallic nanoparticles for the photocatalytic degradation of the target antibiotic, three loadings were tested, namely 0.5, 1.0 and 1.5 wt.%, for each metal. Regarding bi-metallic catalysts, the amount of gold deposited on TiO2 was fixed at 1 wt.%, while Ag and Cu loadings were fixed at 0.5 wt.%. Preparation of bi-metallic materials was performed by sequential deposition–precipitation, as previously reported [29,33]. Ag or Cu was initially deposited as described above; then, after the drying step, Au was deposited. The catalysts were washed, dried, and thermally treated as stated above.

2.2. Characterization Elemental analysis of the photocatalysts was performed by energy-dispersive X-ray spectroscopy (EDS) using a scanning electron microscope JEOL 5900-LV with micro analysis on an EDS Oxford ISIS. The size of the metallic nanoparticles was determined by transmission electron microscopy (TEM) on a JEOL JEM 2010 microscope, operated at 200 keV. Particle size distribution and average particle size were determined by measuring at least 500 particles of each sample. The specific surface area was calculated by the Brunauer–Emmett–Teller (BET) method on a Quantachrome Autosorb 1. Before nitrogen adsorption, the materials were dried and outgassed at 80 ◦ C for 24 h. Determination of the band-gap energy of the photocatalysts (Eg ) was performed by diffuse refractance spectroscopy (DRS), in an Agilent Cary 5000 spectrophotometer. The Eg values were determined using the Kubelka–Munk model [38]. Solid-state photoluminescence (PL) spectra were obtained using a Horiba Fluorolog-3 modular system. Excitation at 286 nm was supplied to the samples using a Xenon lamp at room temperature, and a major peak at 366 nm was followed to estimate the band-to-band transition [13]. Hydrogen temperature-programmed reduction (H2 -TPR) of the dried samples was performed in a RIG-150 in situ research unit using a flow of 30 mL/min of 10% H2 /Ar gas mixture and a heating rate of 10 ◦ C/min from 50 ◦ C to 600 ◦ C.

Please cite this article in press as: J.C. Durán-Álvarez, et al., Photocatalytic degradation of ciprofloxacin using mono- (Au, Ag and Cu) and bi- (Au–Ag and Au–Cu) metallic nanoparticles supported on TiO2 under UV-C and simulated sunlight, Catal. Today (2015), http://dx.doi.org/10.1016/j.cattod.2015.07.033

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2.3. Photocatalytic tests The photocatalytic performance of the surface modified materials was assessed via the degradation of ciprofloxacin using either UV-C or simulated sunlight. The target compound was used as a 30 mg/L solution in tridistilled water. Preliminary tests showed that the optimal concentration of the catalyst in the reaction system was 0.5 g/L. Photocatalytic reactions were carried out in a 250 mL batch cylindrical glass reactor equipped with an external jacket connected to a recirculation bath to maintain the temperature at 25 ◦ C. A 15 W Pen-Ray low pressure Hg lamp with primary emission at 254 nm was put into a quartz tube at the center of the reactor. 250 mL of the antibiotic solution was initially stirred in the reactor to achieve the reaction temperature, and then the photocatalyst was added. The suspension was bubbled with air at 100 mL/min and stirred in the dark for 1 h to ensure the adsorption/desorption equilibrium; after this time lapse, the lamp was turned on. The reaction time was 180 min and 7 mL samples were withdrawn after 0, 5, 10, 15, 30, 45, 60, 90, 120 and 180 min of irradiation. UV-C light was supplied for 3 h at an intensity of 44 W/m2 , resulting in a dose of 475 kJ/m2 . The materials that showed the best performance in UV-C experiments were tested for the degradation of ciprofloxacin using simulated sunlight. Such experiments were performed in a solar simulator SUNTEST CPS+, equipped with a 1500 W Xenon lamp and 560 cm2 of exposed area. Simulated sunlight was applied to samples for 6 h using an intensity of 500 W/m2 , resulting in a dose of 10,800 kJ/m2 . Throughout experiments, samples were agitated using an orbital shaker at 150 rpm. Temperature in all the experiments was maintained below 35 ◦ C. Disposable 7 mL samples were taken at the beginning of experiments, after 1 h of agitation, when adsorption equilibrium was reached, and upon 30, 60, 90, 120, 180, 240, 300 and 360 min of irradiation. All the samples were filtered using nylon membranes (0.2 ␮m pore size) prior to the analysis. Experiments with no addition of catalyst were carried out for both UV-C and simulated sunlight in order to assess the degradation of the antibiotic by photolysis. Degradation data were fitted to the pseudo-first order kinetic approach. The apparent photocatalytic reaction rate for each treatment was calculated using Eq. (1): Ct = C0 × e−kt

(1)

where C0 and Ct are the concentrations of the antibiotic at time 0 and t (in mg/L), respectively; kapp is the pseudo-first order degradation rate constant (min−1 ), and t is the experimental time (min). Given that photolysis experiments were carried out, the increment in the degradation rate by the use of the photocatalysts was determined for each material and expressed as the degradation rate increase factor. At the end of the photocatalytic tests, the photocatalyst was recovered by centrifugation at 10,500 rpm for 10 min, followed by a drying step at 80 ◦ C for 2.5 h under vacuum conditions. The recovered photocatalysts were reused in three consecutive cycles in order to assess the decay in photocatalytic performance by deactivation. 2.4. Instrumental analysis The quantification of the target compound as well as the identification of a group of by-products was carried out using liquid chromatography in tandem with electrospray ionization and triple quadrupole detector (HPLC–ESI–MS/MS), Agilent Technologies 6400 series. Chromatographic separation was achieved using a C18 Eclipse Plus (2.1 mm × 150 mm × 3.5 ␮m) column. 0.1% formic acid aqueous solution (Solvent A) and acetonitrile (Solvent B) were used as mobile phase at a flow rate of 0.4 mL/min. The

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composition of the mobile phase throughout analysis was as follows: 90% of A and 10% of B for 25 min; then 40% of A and 60% of B for 5 min; and, 90% of A and 10% of B for the last 5 min. MS/MS analysis was performed in the positive mode; 332 was followed as the molecular ions [M+H]. Fragmentation conditions were as follows: cone voltage was maintained at 60 V; drying gas (nitrogen) was supplied at 13 L/min and 350 ◦ C; the voltage of the capillary was set at 3000 V; the collision cell voltage was kept at 16 V. In MS/MS analysis, the transitions 332–314, 332–288 and 332–231 were followed for identification. Levofloxacin was used as internal standard for quantification of ciprofloxacin. For identification of the by-products, chromatographic separation and mass analysis conditions were the same as described for ciprofloxacin. Table 1 in the Supporting Information section shows the chemical structure and mass transitions of the degradation byproducts selected for this study. Mineralization of ciprofloxacin was determined by analysis of total organic carbon. A Shimadzu TOC–LCSH/CPH analyzer was used for determination of organic carbon by wet combustion. The mineralization percentage at the end of the experiments was determined using Eq. (2). (%) Mineralization =

TOCinitial − TOCfinal × 100 TOCinitial

(2)

where TOCinitial and TOCfinal refer to the organic carbon concentration at the beginning and at the end of the photocatalytic experiments.

2.5. Toxicity tests In those cases mineralization was incomplete at the end of the photocatalytic tests both the toxicity and the residual antibiotic activity were assessed in water samples. Toxicity was determined via inhibition of the bioluminescence of the marine bacteria V. fischeri, using a LUMIStox 300 kit (HACH LANGE). Freezy-dried V. fischeri was acclimated from −15 ◦ C to −5 ◦ C for 1 h and then suspended in 2% NaCl. The suspension was put in contact with samples at different dilution levels, in accordance to standard method ISO 11348-3, and natural bioluminescence of the bacteria was measured after 0, 5, 15 and 30 min of incubation at 25 ◦ C. The inhibition of bioluminescence was calculated as described by the ISO 11348-3 standard method. The concentration at which 50% of the bacteria population displayed inhibition effect (EC50 ) was determined using the dose–response curves resulting in the tests. The residual antibiotic activity in water samples was evaluated as a function of the growth of bacterial colonies in agar prior to and after the photocatalytic treatment by the Kirby–Bauer method. Briefly, E. coli ATCC bacteria were growth on TBX agar medium and transferred into 10 mL of physiological solutions to achieve the concentration of 107 CFU/100 mL. Then, the bacterial suspensions were spread on Mueller Hinton agar. Nitrocellulose sensi-discs impregnated with the water samples taken at 0, 180 and 360 min of simulated sunlight irradiation were placed on the surface of each inoculated plate. After 24 h of incubation at 37 ◦ C, the diameter of the growth inhibition zone around the sensi-discs was measured. As the bacterial colonies grow, a ring which marks the inhibition capacity of the antibiotic on the bacterial growth can be observed around the sensi-discs. The larger the inhibition diameter of the sensi-discs, the higher the antibiotic activity remains in water sample. The inhibition zone diameter at 0 min of irradiation represents 100% of antibiotic activity; then inhibition zone diameters measured throughout photocatalytic experiments are divided by the diameter of the inhibition zone at 0 min of irradiation and expressed as residual antibacterial activity in percentage.

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3. Results 3.1. Characterization of the materials According to XRD anatase was the main crystallographic phase obtained in the sol–gel procedure, traces of brookite were also found (see Fig. 1 in the Supporting Information section). Table 1 compares the nominal and the actual loading of metals deposited on the TiO2 surface. For both mono- and bi-metallic catalysts, the measured loadings of Au and Cu metals were close to the nominal ones, evidencing the high efficiency of deposition when urea is used as basifying agent; this is consistent with results previously reported [20,29,33,39]. Unlike Au and Cu, differences were found in the measured loadings of Ag compared with the nominal ones for both mono- and bi-metallic materials. Precipitation of Ag+ ions does not occur when urea is used as basifying agent; thus deposition–precipitation of Ag was carried out using NaOH as precipitation agent, which resulted in low loadings of the metal. For bi-metallic materials, differences were observed between the nominal and the measured loadings of Ag and Cu. This may find an explanation in the lixiviation of the initially deposited metal (Ag or Cu) during deposition of the second one (Au). Metallic nanoparticles were well-dispersed on the TiO2 surface (Fig. 1). The average size of the metallic nanoparticles was maintained below 4 nm (Au and Ag nanoparticles displayed the largest size), showing low standard deviation (see histograms in Fig. 1). The addition of the second metal showed no increase in the size of nanoparticles. Deposition of metallic nanoparticles slightly increased the specific surface area of the sol–gel TiO2 . The specific surface area was similar for the different mono-metallic catalysts, while a higher surface area was observed for the bi-metallic ones. H2 -TPR analysis (Fig. 2a) showed that Au atoms on the freshly synthesized material (i.e., prior to the thermal treatment) occurred in the oxidized state as Au (III), in both mono- and bi-metallic materials. A low temperature peak centered at 120 ◦ C (from 100 ◦ C to 165 ◦ C) evidenced the reduction of the Au (III) species [39]. Ag atoms

reached the metallic state during deposition–precipitation, thus no signal of H2 consumption was observed in the TPR spectra for both mono- and bi-metallic Ag modified materials. Decomposition of AgOH into Ag and O2 during the NaOH deposition–precipitation process has been previously reported [29]. Cu/TiO2 showed one low temperature reduction peak at 155 ◦ C (from 140 ◦ C to 180 ◦ C) and another broad peak at 400 ◦ C (from 300 ◦ C to 500 ◦ C). Both peaks have been attributed to reduction of Cu (II) species, firstly to Cu (I), then to metallic Cu [20]. The Au–Ag material showed only the characteristic reduction peak of Au, centered at 120 ◦ C, indicating the occurrence of metallic silver in the bi-metallic material prior to thermal treatment. For Au–Cu bi-metallic catalyst, a low temperature reduction peak was found between the reduction peaks of mono-metallic Au and Cu (reduction peak was centered at 180 ◦ C). This may be an indication of the interactions established by Au and Cu atoms in the freshly prepared material, which influences the reducibility of Au and Cu species [36]. Additionally, a broad reduction peak at 400 ◦ C as the observed here has also been observed in previous reports [20,36]. DRS spectra are shown in Fig. 3. All of the synthesized materials showed UV light absorption with a characteristic edge near 370 nm. This is typical of TiO2 , and evidences its low absorption in the visible light spectrum. Band gap values determined for the modified materials were similar to those calculated for bare sol–gel TiO2 (Table 1). Surface plasmon resonance was displayed by both mono- and bi-metallic materials (see insertion in Fig. 3), confirming that nanoparticles reached the zero valence state upon the thermal treatment at 350 ◦ C in H2 atmosphere. Plasmon resonance of Au–Ag bi-metallic material was blue shifted compared to that displayed by Au modified TiO2 (560 nm). Plasmonic band of the bi-metallic material was similar to that expressed by Ag/TiO2 (485 nm), which may be attributable to a limited dispersion of Ag atoms within the bi-metallic alloy [29]. Unlike Au–Ag/TiO2 , the plasmon resonance band of the Au–Cu bi-metallic material was expressed at 545 nm, which is near of the plasmonic band of Au. This may be due to the low proportion of Cu atoms in the alloy.

Fig. 1. Particle size distribution and TEM images for photocatalysts: (a) Au/TiO2 , (b) Ag/TiO2 , and (c) Cu/TiO2 .

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Table 1 Relevant characteristics of the as-synthesized catalysts and comparison with TiO2 sol–gel. Catalyst

Nominal loading of metal (wt.%)

Actual loading of metal (wt.%)

Eg (eV)

Average particle size (nm)

Surface area (m2 /g)

Au/TiO2

0.5 1.0 1.5

0.6 ± 0.1 0.9 ± 0.2 1.4 ± 0.4

3.16

3

52

Ag/TiO2

0.5 1.0 1.5

0.4 ± 0.05 0.8 ± 0.2 1.3 ± 0.05

3.19

4

53

Cu/TiO2

0.5 1.0 1.5

0.6 ± 0.1 0.9 ± 0.1 1.5 ± 0.2

3.15

2.5

56

Au–Ag/TiO2

1–0.5

Au: 1.0 ± 0.1 Ag: 0.4 ± 0.1

3.1

2.5

63

Au–Cu/TiO2

1–0.5

Au: 1.0 ± 0.1 Cu: 0.4 ± 0.08

3.15

2.5

64

TiO2 sol–gel

–

–

3.2

–

48

Photoluminescence spectra of the catalysts obtained by excitation at 286 nm and ambient temperature are shown in Fig. 4. Emission at 366 nm was attributed to self-trapped exciton within the TiO6 octahedra (i.e., band gap transition). The decay in the

emission peak intensity displayed by surface-modified materials was attributed to the increased separation of the hole–electron pair caused by the migration of the electron to the metallic nanoparticles [31]. The lowest recombination rate was observed for Au and

A

B Au/TiO

Ag/TiO2

Ag/TiO

Cu/TiO2

Cu/TiO

0

H2 uptake (a.u.)

H2 uptake (a.u.)

Au/TiO2

100

200

300

400

Temperature (°C)

Au-Ag/TiO2

Au-Ag/TiO

Au-Cu/TiO2

Au-Cu/TiO

500

600

0

100

200

300

400

500

600

Temperature (°C)

Fig. 2. TPR profiles for (a) freshly synthesized photocatalysts, and (b) photocatalysts exposed to reaction conditions.

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Fig. 3. UV–vis absorption spectra of the surface-modified photocatalysts. Surface plasmon resonance is shown in insert.

Au–Ag modified TiO2 , while separation of the hole–electron pair decayed as the value of the work function of metals decreases; this has been previously reported by Ref. [20]. 3.2. Determination of the optimal loading of metallic nanoparticles In the first stage of the experiment, the optimal loading of metallic nanoparticles in terms of photocatalytic degradation of the antibiotic was determined using UV-C light. The optimal loadings of metallic nanoparticles were found as: 1.5 wt.% for Au and Ag, and 1.0 wt.% for Cu (Fig. 5). Higher loadings of metallic nanoparticles resulted in decay of the ciprofloxacin degradation rate. This is consistent with that previously reported [40], and is related with the screening effect of photons exerted by nanoparticles in the surface of the semiconductor [40,41]. The optimal loading of Cu nanoparticles was lower than those observed for Au and Ag. This may be explained by the oxidation of Cu nanoparticles into Cu (II) species during photocatalytic process (as shown in H2 -TPR, Fig. 2b), which resulted in the decrement of the photocatalytic activity as Cu loading increased [42]. On the other hand, given that Au and Ag nanoparticles are not oxidized during the photocatalytic process (Fig. 2b), higher loadings of these metals can be used with no significant decrease in the photocatalytic performance. High degradation rates of organic pollutants in water have been previously reported using similar loadings of metallic nanoparticles on TiO2 . For instance, Sangpour et al. [43] observed high degradation rates of methylene blue by using Au, Ag and Cu nanoparticles deposited

100 0

TiO2 sol-gel Ag/TiO2 Au-Ag/TiO2

on TiO2 thin films at a molar concentration of 2.7%. Pugazhenthiran et al. [12,13] reported high degradation yields of cephalosporin antibiotics using TiO2 nanotubes loaded with Au and Ag nanoparticles at 2 wt.%. On the other hand, it is noteworthy that in some cases acceptable degradation rates of organic pollutants in water have been obtained using loadings of metallic nanoparticles lower than those observed in this work [20,21]. 3.3. Photocatalytic degradation and mineralization kinetics under UV-C light irradiation Adsorption of ciprofloxacin onto the catalyst surface was determined for all of the synthesized materials. As shown in Fig. 2 of the Supporting Information section, adsorption of ciprofloxacin upon 30 min of stirring in dark conditions varied from 10% to 20%, depending on the material. No further adsorption was observed after 30 min of stirring in dark conditions. A notable increment in the adsorption of ciprofloxacin molecules was observed upon the deposition of metallic nanoparticles on the TiO2 surface. UV-C light photolysis played an important role in the degradation rate of the antibiotic, and higher degradation rate was observed upon the addition of bare TiO2 (Fig. 6). The increment factor was determined as 1.18. This is in line with that reported in other works [44,45]. Further increment in the degradation rate of ciprofloxacin was observed when mono- and bi-metallic materials were tested (Table 2). The highest apparent photocatalytic degradation rate (kapp ) of ciprofloxacin was found when (1.5 wt.%) Ag/TiO2 was

Au/TiO2 Cu/TiO2 Au-Cu/TiO2

Intensity (counts)

800

600

400

200

0 300

350

400

450

500

Wavelength (nm) Fig. 4. Photoluminescence absorption spectra of the surface-modified photocatalysts.

Fig. 5. Pseudo-first order photocatalysis degradation rates of ciprofloxacin using different metal loading in mono-metallic materials.

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Table 2 Photolytic and photocatalytic degradation parameters of ciprofloxacin in UV-C light experiments. Energy dose for complete removal (kJ/m2 )

System Photolysis TiO2 sol–gel 1.5% Au/TiO2 1.5% Ag/TiO2 1.0% Cu/TiO2 Au–Ag/TiO2 Au–Cu/TiO2

237.5 237.5 158.5 79.2 158.5 237.5 237.5

Concentration (mg/L)

30

TiO2 sol-gel

0.048 0.057 0.06 0.117 0.072 0.053 0.099

0.992 0.997 0.993 0.991 0.988 0.995 0.991

Degradation rate increment by photocatalysis – 1.18 1.25 2.4 1.5 1.1 2

1.5% Au/TiO2 1.5% Ag/TiO2

20

3.4. Photocatalytic degradation and mineralization kinetics using simulated sunlight

1% Cu/TiO2 1% Au-0.5% Ag/TiO2

15

1% Au-0.5% Cu/TiO2

10 5 0 -60

r2

end of experiments was obtained when Au/TiO2 and Au–Cu/TiO2 materials were tested (see Fig. 7).

Photolysis

25

kapp (min−1 )

Srringin dark -30

Photocatalysis 0

30

60

90

120

150

180

Time (min) Fig. 6. Photolysis and photocatalysis kinetics of ciprofloxacin using selected materials under UV-C light irradiation.

30

Photolysis TiO2 sol-gel

25

Concentration (mg/L)

used, while lower degradation rates were obtained for (1.0 wt.%) Cu/TiO2 and (1.5 wt.%) Au/TiO2 catalysts. Au–Ag/TiO2 displayed lower activity than their mono-metallic counterparts (Table 2). The kapp displayed by this bi-metallic material was similar than that obtained for Au/TiO2 , while Au–Cu/TiO2 catalyst was as active as Ag/TiO2 . The complete mineralization of ciprofloxacin was achieved upon 180 min of UV-C light irradiation for all of the modified materials (Fig. 7). In contrast, minimal mineralization (removal of TOC below 20%) was observed in the photolysis assay. Mineralization followed a two-step kinetic (Fig. 3, in the Supporting Information section). In the first step, low-rate mineralization was observed, indicating that the parent compound was transformed into by-products and only a small fraction of these by-products were transformed into mineral products. In the second mineralization stage, by-products obtained in the first stage were rapidly converted into mineral products, i.e., CO2 and H2 O, but also F– , NO3 – and NH4 + . In accordance with the mineralization kinetics obtained for ciprofloxacin, the first mineralization step lasted from 45 to 60 min; then complete mineralization was attained upon near 180 min of irradiation. The highest mineralization percentage at the

Optimal loading of metallic nanoparticles observed in UV-C assays were used to synthesize the photocatalysts tested in simulated sunlight experiments; however, it is noteworthy that results obtained using UV-C are not necessarily transposable to low energy light conditions, thus future research must be done to determine the optimal loading under simulated sunlight. Fig. 8 depicts the photolytic and photocatalytic degradation kinetics of ciprofloxacin using simulated sunlight. The temperature of the reaction in the SUNTEST device was kept below 35 ◦ C, thus lower adsorption of the compound on the TiO2 surface (see Fig. 4 in the Supporting Information section) and low concentration of dissolved O2 (4.6 mg/L) than those determined in UV-C light experiments (6.3 mg/L) were observed. The degradation rates and the time at which complete degradation was achieved were significantly lower than those observed in the UV-C light tests (Table 3). Ciprofloxacin was not fully degraded by photolysis upon 360 min of irradiation, which is consistent with that is reported in the literature [45]. The low degradation rate observed in photolysis tests evidenced the low activity of sunlight to break molecules with high electron density (i.e., quinolone structure). Other flouroquinolones have shown low removal rates by sunlight photolysis [11]. Photocatalysis, on the other hand, showed to significantly increase the degradation rate of ciprofloxacin (degradation increase factor was above 5 for all the photocatalysts). Reaction rates were higher for mono-metallic materials compared to the bi-metallic ones (Table 3), Au and Ag modified catalysts showed the highest degradation rates (Fig. 8). Degradation rates achieved by the

1.5% Au/TiO2 1.5% Ag/TiO2 1% Cu/TiO2

20

1% Au-0.5% Ag/TiO2 1% Au-0.5% Cu/TiO2

15 10 5

Srring in dark

Photocatalysis

0 -60

0

60

120

180

240

300

360

Time (min) Fig. 7. Mineralization of ciprofloxacin in photolysis and photocatalysis upon 90 and 180 min of UV-C light irradiation.

Fig. 8. Photolysis and photocatalysis kinetics of ciprofloxacin using selected materials under simulated sunlight irradiation.

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Mineralization (%)

100

360 min

80 60 40 20 0 Photolysis

TiO2 sol-gel 1.5% Au/TiO2 1.5% Ag/TiO2 1% Cu/TiO2

Au-Ag/TiO2

Au-Cu/TiO2

Fig. 9. Mineralization of ciprofloxacin in photolysis and photocatalysis upon 180 and 360 min of simulated sunlight irradiation.

Table 3 Photolytic and photocatalytic degradation parameters of ciprofloxacin in simulated sunlight experiments. System

Energy dose for complete removal (kJ/m2 )

k (min−1 )

r2

Degradation rate increment by photocatalysis

Photolysis TiO2 sol–gel 1.5% Au/TiO2 1.5% Ag/TiO2 1% Cu/TiO2 Au–Ag/TiO2 Au–Cu/TiO2

>10,800 10,800 7200 7200 10,800 10,800 10,800

0.0036 0.021 0.042 0.04 0.023 0.021 0.022

0.963 0.954 0.926 0.913 0.964 0.995 0.983

– 5.8 11.7 11.1 6.4 5.8 6.1

bi-metallic materials were similar to that obtained for unmodified TiO2 . The complete mineralization of ciprofloxacin was achieved upon 360 min of irradiation only when bi-metallic catalysts were tested (Fig. 9). When mono-metallic materials were used, a trace of ciprofloxacin remained after 360 min of irradiation along with several by-products. As occurred in UV-C assays, two-step mineralization kinetics was observed for all of the tested materials. 3.5. Stability of the catalytic materials The stability of the synthesized photocatalysts was evaluated by comparing the degradation rate constants obtained in three consecutive reaction cycles. No significant decay in rate constants was observed for the mono- and bi-metallic materials upon several reaction cycles (Fig. 10). As observed in H2 -TPR analysis (Fig. 2b), Au and Ag nanoparticles are not oxidized during the photocatalytic process, additionally no carbonate ions were detected in the catalysts at the end of the experiments (data not shown); thus deactivation of these materials is not expected. Contrary to Au and Ag nanoparticles, Cu was oxidized into Cu (II) species during the photocatalytic processes in both mono- and bi-metallic Cycle 1

Cycle 2

Cycle 3

Degradation rate (min-1)

0.16

0.12

0.08

0.04

0 1.5% Au/TiO2

1.5% Ag/TiO2

1.0% Cu/TiO2

Au-Ag/TiO2

Au-Cu/TiO2

Fig. 10. Degradation kinetic constant of ciprofloxacin after three reaction cycles using UV-C light.

materials (Fig. 2b). Moreover, DRS analysis of Cu/TiO2 at the end of the tests showed the absence of the plasmon band at 670 nm; thus deactivation of the catalyst was expected. The photocatalytic activity of the Cu/TiO2 material after several reaction cycles may be due to the formation of the TiO2 –CuO binary composite in which the photo-formed electron in the CuO semiconductor migrate to the conduction band of TiO2 , resulting in the production of ROS on the TiO2 surface. The stability of surface-modified TiO2 with mono-metallic nanoparticles upon several reaction cycles has been reported elsewhere [12,13,23,43]. 3.6. By-products identification Identification of by-products was carried out using HPLC–ESI–MS/MS. By-products were identified only for photolysis and photocatalysis using (1.5 wt.%) Ag/TiO2 and UV-C light irradiation; further studies will aim to compare the chemical nature of the by-products produced by different catalysts and light sources. In the photolytic process four by-products were identified: CB1 (m/z = 330), CB2 (m/z = 334), CB3 (m/z = 362) and CB4 (m/z = 306). Transitions used to identify each by-product were based in what is reported by Calza et al. [46]. The four identified by-products persisted in water samples throughout the photolytic process and they corresponded to the TOC concentrations found in water samples (see low mineralization rates in UV-C photolysis tests showed in Fig. 7). By-product CB1 resulted from the substitution of the fluorine atom by OH− in the position 6 of the fluoroquinolone moiety (bond dissociation energy = 120 kcal/mol). CB1 molecule is then hydroxylated and carboxyl is substituted by OH− to form CB2. Hydroxylation is the most common reaction in photolytic processes, which has been reported for several pharmaceutical compounds in water [47]. The fluoroquinolone structure can reach the triplet state upon UV light absorption, resulting in a strong electrophile that favors the hydroxylation of the molecule [48,49], and contributing in turn to formation of CB2. A second reaction mechanism was elucidated in the photolytic process of ciprofloxacin. In this mechanism the substitution of the fluorine atom does not occur, while cleavage of the piperazine ring takes place in positions 5 and 6. Oxidation of the open piperazine ring occurs to form CB3. Lastly, 2 C O moieties are released from the

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Fig. 11. Photolysis and photocatalysis by-products of ciprofloxacin identified by HPLC–MS/MS analysis and proposed reaction mechanisms. Appearance/disappearance kinetics of the degradation products in inserts.

piperazine ring and CB4 is produced. Abstraction of the piperazine ring may also occur upon the triplet formation [49], resulting in CB4. Other by-products were identified in the photolytic reaction, such as m/z 360 and m/z 261, probably corresponding to hydroxylation of CB3, followed by cleavage of the piperazine ring. However, consistent fragmentation was not obtained to bear the premise of their presence in water samples and thus they were not included in the by-product list. Given the lack of commercially available standards of the ciprofloxacin by-products, concentration of these molecules was not determined in water samples, instead normalized peak areas were obtained by dividing the ion peak area by the maximum peak area detected during the reaction time course and thus a semi-quantitative study was performed. Fig. 11 shows, on the one hand, the proposed reaction mechanisms of ciprofloxacin in the photolytic and photocatalytic processes, and on the other hand, the appearance/disappearance kinetics of the by-products identified by MS/MS analysis. The highest abundance was observed for the defluorinated by-product CB2, which appeared from 10 min and increased via the hydroxylation of CB1. A plateau is reached in CB1 appearance from 45 min, then formation of CB2 significantly decreased. Lastly, CB1 formation decreased while CB2 appearance continued. Regarding CB3, it initially appeared after 10 min of irradiation and concentration increased until 60 min. CB4 was produced through the cleavage of the piperazine ring from 30 min. Upon 1 h of irradiation the production rates of CB3 and CB4 were sustained. After 180 min of UV-C irradiation no decay of both by-products was observed. In the photocatalytic process, four by-products were identified, namely CB3 (m/z = 362), CB4 (m/z = 306), CB5 (m/z = 291) and CB6 (m/z = 263). Substitution of the fluorine atom in the fluoroquinolone moiety was not observed in the photocatalytic process;

instead degradation of ciprofloxacin occurs by changes in the piperazine ring. This indicates that photocatalytic oxidation of the ciprofloxacin molecule occurs more rapidly than the formation of the triplet state. According to Witte et al. [50], the enhanced reactivity of fluoroquinolones for hydroxyl radical attack is found in position 2 of the quinolone moiety as well as in the N atoms of the piperazine ring. Our results indicate that photocatalytic transformation of ciprofloxacin occurred via piperazine ring cleavage rather than the breakage of the quinolone moiety. The reaction started with the cleavage of the piperazine ring resulting in CB3. This byproduct was consistently formed until 60 min of irradiation, then its disappearance occurred by the release of 2 C O moieties, resulting in CB4. As CB4 was formed, CB5 appeared, via the hydroxylation and the release of CH NH2 from the CB4 molecule. Appearance of CB5 started after 15 min of reaction and disappeared after nearly 180 min, at the end of the assay. Further liberation of C O from the modified piperazine ring led to the formation of CB6. This byproduct prevailed until the end of the reaction. The structure of fluoroquinolone prevailed until late stages in the photocatalytic reaction. This was demonstrated by UV–vis spectrophotometric analysis of water samples taken throughout the photocatalytic process. The prevalence of a shoulder between 310 and 330 nm, the characteristic peak of the fluoroquinolone structure, was observed in water samples taken at 90 and even 120 min (Fig. 5 in the Supporting Information section). Some by-products, which transitions suggest the cleavage of the fluoroquinolone moiety, such as anthranilic acid (m/z = 137) and benzoic acid (m/z = 122), were identified toward the end of the photocatalytic process, although identification occurred only in water samples taken at 90 and 120 min of reaction; thus such structures were not included in the reaction mechanism. Nevertheless, these by-products may be

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60 40 20 0 10

20

30

40

50

80 60 40 20 0

60

0

10

Concentration (%)

40

50

60 40 20 0 20

30

40

Concentration (%)

20 0

10

50

60

80 60 40 20 0

10

20

30

40

20

30

40

50

60

Au-Cu/TiO2

100

0 10

40

Concentration (%)

% Inhibition

80

0

60

0

60

Au-Ag/TiO2

100

% Inhibition

% Inhibition

30

80

Concentration (%)

1.0 % Cu/TiO2

100

20

1.5% Ag/TiO2

100

% Inhibition

80

0

1.5% Au/TiO2

100

% Inhibition

% Inhibition

100

50

60

80 60 40 20 0 0

10

Concentration (%)

20

30

40

50

60

Concentration (%)

Fig. 12. Dose–response charts of ciprofloxacin obtained in acute toxicity tests.

the basis for elucidation of the fluoroquinolones final degradation mechanism in the photocatalytic process. 3.7. Toxicity tests Mineralization of ciprofloxacin was not fully achieved in simulated sunlight tests (as occurred in UV-C irradiation tests) and by-products (as seen for mono-metallic materials) were present in water samples at the end of photolytic and photocatalytic process. Acute toxicity was evaluated by the decay of natural bio-luminescence of the marine bacteria V. fischeri when it is exposed to water samples taken upon 360 min of simulated sunlight irradiation. Fig. 12 shows the dose–response charts obtained for ciprofloxacin treated with all the modified materials. Toxicity was determined as the GL20 factor, which represent the reciprocal of the dilution value at which 20% of the bio-luminescence is inhibited by bacterial death. The higher the value of GL20 factor the higher the toxicity of the water sample. The highest toxicity was determined for the photolysis process, which was expected given the low mineralization observed. On the other hand, toxicity in water samples after the photocatalytic process was significantly lower, notably for the bi-metallic photocatalysts. Even when the use of mono-metallic materials did not resulted in the complete mineralization, toxicity results can be classified as negligible. These results are consistent with those reported in the literature for photocatalytic degradation of ciprofloxacin using visible light [46]. Determination of residual antibiotic activity in water samples at the end of photolysis and photocatalysis tests using simulated sunlight were carried out through bacterial growth inhibition tests. E. coli ATCC strain was grown in Mueller Hinton agar and sensidiscs, impregnated with water samples taken at different reaction times (0, 180 and 360 min), were applied on the agar plate. Residual antibiotic activity in water samples was related with the diameter of the growth inhibition zone. Table 4 and Fig. 6 in the Supporting Information section show the residual antibiotic activity of water samples taken in photolysis and photocatalysis experiments. Incipient decay of the antibiotic concentration in photolysis tests resulted in 58% of the antibiotic activity remaining in water samples upon 360 min of irradiation. In photocatalysis, a clear decay of the antibiotic activity was observed upon 180 min of irradiation. After such time lapse, ciprofloxacin was completely degraded but not mineralized, thus it is concluded that ciprofloxacin by-products maintain some biocide effect capable to inhibit bacterial growth. However, there is no evidence to assert that such residual antibiotic activity is enough to lead pathogens to express antibiotic resistance. At the end of the photocatalytic process, all of the water

samples corresponding to mono-metallic catalysts displayed low levels of residual antibiotic activity (Table 4). On the contrary, the use of surface modified TiO2 with bi-metallic nanoparticles resulted in the complete removal of the antibiotic activity. This is consistent with the complete mineralization of the antibiotic achieved upon 360 min of simulated sunlight irradiation (Fig. 9). From these results, it is possible to conclude that the tested photocatalysts were able to both satisfactorily remove the target pollutant and to produce safe effluents from the perspective of toxicity and the potential to disseminate resistance to antibiotics in the bacteria communities living in the receiving water. 4. Discussion 4.1. Photocatalytic degradation of ciprofloxacin using UV-C light Due to the chemical nature of fluoroquinolones, pH is the parameter ruling the adsorption of ciprofloxacin onto the catalyst surface. Fig. 7 in the Supporting Information section shows the speciation of the antibiotic at different pH values. Considering that the isoelectric point of TiO2 is 6.3 [51], it is possible to predict the extent of adsorption of ciprofloxacin on the catalyst by knowing the pH of the suspension. Photocatalysis tests were performed in unbuffered suspensions; thus pH was 6.1 in antibiotic standard solutions, decreasing to 5.5 at the end of experiments. Under these pH values, 50% of the ciprofloxacin molecules displayed double positive charge, 45% occurred with one single positive charge and less than 5% was present in the suspension in its zwitterionic form [45]. Regarding TiO2 , it occurred as zwitterion at the pH values of the experiment. Ciprofloxacin molecules are adsorbed onto TiO2 by electrostatic forces between the positive charges of the molecule and the TiO– fraction of the catalyst; hence the highest adsorption rates of ciprofloxacin have been reported at neutral pH, in which the zwitterion form of both catalyst and antibiotic prevails [45]. Adsorption of fluoroquinolone antibiotics on TiO2 surface has been previously reported as around 15% upon 0.5–1 h of stirring in pure water [45,52], which occurs majorly when anatase is used [53]. In our experiments 10–20% of ciprofloxacin was adsorbed. The high adsorption rate of ciprofloxacin may be explained by the deposition of metallic nanoparticles onto the TiO2 surface. High adsorption of fluoroquinolones on metallic nanoparticles has been previously observed; for instance, Tom et al. [54] reported that up to 65 molecules of ciprofloxacin can be adsorbed onto a 4 nm Au nanoparticle. The adsorption of ciprofloxacin onto metallic nanoparticles can be attributed to binding with the NH moiety in the piperazine ring [54].

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Table 4 Residual antibiotic activity of ciprofloxacin after 6 h of simulated sunlight irradiation using different modified catalysts. Antibiotic

Residual antibacterial activity (%) Photolysis

1.5% Au/TiO2

1.5% Ag/TiO2

1% Cu/TiO2

Au–Ag/TiO2

Au–Cu/TiO2

Ciprofloxacin

58.33

4.17

12.50

9.09

0

0

In photocatalysis tests, degradation of ciprofloxacin effectively increased when modified TiO2 was used. The increase in the photocatalytic degradation rate of ciprofloxacin when surface modified catalysts were used can be attributed to the electron trap effect exerted by metallic nanoparticles, resulting in the lower recombination of the hole–electron pair (see photoluminescence spectra shown in Fig. 4). The migration of photo-produced electrons from the conduction band to the metallic nanoparticles resulted in the higher formation of ROS on the surface of the metallic nanoparticles [40]. According to the reported in literature, photocatalytic degradation of fluoroquinolone antibiotics is mainly caused by the oxidation through photo-formed holes (near 63% of degradation), whereas oxidation reactions with ROS occur to a lesser extent (less than 25% degradation) [52,55]. According to the aforementioned, it is plausible to attribute the increase in the degradation rates of the antibiotic to both the extended life of holes and the formation of ROS on the surface of the metallic nanoparticles. In accordance to the hypothesis of the increased degradation rate as a result of the inhibition in the recombination, Au/TiO2 catalyst was expected to display the highest degradation rates, as has been previously observed in our group for other antibiotics [20]; surprisingly, this did not occur. The higher degradation rates observed for Ag/TiO2 , and Cu/TiO2 , compared with that obtained for Au/TiO2 suggest that, in addition to the charge separation, other phenomena are contributing to the photocatalytic degradation of the antibiotics. One of these phenomena may be the increment in the production of superoxide radicals (• O2 − ) on the surface of Ag and Cu metallic nanoparticles. Due to their lower electronegativity, Ag and Cu have greater potential than Au nanoparticles to adsorb dissolved O2 ; then the electron trapped in the nanoparticle is transferred to the dissolved O2 acting as an electrophile [43]. High photocatalytic degradation of organic pollutants using Ag and Cu modified TiO2 has been previously reported. For instance, Sangpour et al. [43] reported higher degradation rates of methylene blue using Cu/TiO2 compared with Au/TiO2 . The authors explained this behavior by the increased presence of oxygen deficiencies in the Cu modified materials, which in turns favored the adsorption of dissolved oxygen and thus the production of superoxide radicals. Other studies comparing different mono-metallic surface modified semiconductors have reported higher degradation of antibiotics using Ag/TiO2 and Cu/TiO2 compared to Au/TiO2 catalyst [43]. Shokri et al. [21] found the complete degradation of chloramphenicol in only 20 min using (1 wt.%) Ag/TiO2 , while Bu and Zhang [22] reported total degradation of chlortetracycline upon 2 h of UV irradiation using Cu/TiO2 hollow spheres. Oxidation of Cu nanoparticles during photocatalysis (as seen in the H2 -TPR analysis, Fig. 2b) can result in the decrease of the antibiotic degradation rate by deactivation of metallic nanoparticles. However, in our experiments this catalyst remained active upon oxidation of the Cu nanoparticles. It is possible to hypothesize that the oxidized material may behave as a binary TiO2 –CuO composite, which may display better catalytic activity for the oxidation of organic compounds than bare TiO2 [56]. Because of the low band gap value of CuO (ca. 1.5 eV), the generation of the hole–electron pair can easily occur, with the subsequent generation of ROS. Additionally, CuO may work as an efficient electron trap from the TiO2 conduction band (work function of CuO is 5.3 eV vs. 4.2 eV for TiO2 ). Regarding the Au–Cu/TiO2 material, it is reported that when Au–Cu alloys are used, copper nanoparticles may be oxidized producing a

CuO@Au/TiO2 core@shell structure, in which the oxidation of the organic molecules is carried out by the oxygen associated to the copper oxide core [57]. 4.2. Photocatalytic degradation of ciprofloxacin under simulated sunlight Photolysis and photocatalysis under simulated sunlight showed degradation results significantly lower than those observed in UVC experiments because of the lower energy of the light source ( = 290–800 nm). Nevertheless, some other facts may contribute to the decrease in the photodegradation rates of ciprofloxacin. The drop of the dissolved O2 concentration in water may result in higher recombination of the photo-formed hole–electron pairs due to the lack of electrophiles to scavenge the electron in the conduction band of the TiO2 , which in turn resulted in the decrease of ROS formation [45]. As stated above, the temperature achieved in the SUNTEST equipment (35 ◦ C) resulted in lower adsorption of ciprofloxacin on the catalyst surface compared to UV-C experiments (at 25 ◦ C). Further studies are necessary to elucidate the effects of temperature increase in the photocatalytic process. Given that the band gap of TiO2 was not modified upon deposition of mono- and bi-metallic nanoparticles, the photocatalytic degradation of ciprofloxacin in experiments using simulated sunlight can be attributable to the harvesting of visible light by metallic nanoparticles via the superficial plasmon resonance effect (all of the synthesized materials expressed plasmon resonance, as showed in Fig. 3), and to a lesser extent by the electron trap effect in mono or bi-metallic nanoparticles. Photolysis of ciprofloxacin under simulated sunlight was notably slow. The complete removal of the antibiotic was not achieved upon 360 min of simulated sunlight irradiation. As most of molecules containing the 4-quinolone moiety, ciprofloxacin can reach the triplet state under sunlight irradiation (at 330–360 nm) and be degraded upon the formation of the triplet state. Degradation of ciprofloxacin via self-excitation occurs by substitution of the fluorine atom and further hydroxylation. In this mechanism, ciprofloxacin zwitterion displays quantum yield over a magnitude order higher than that observed for its cationic form [58]. Given that our experiments were carried out at pH values below 6.9, the cationic form of ciprofloxacin prevailed and thus degradation efficiency was low, the decrease in pH values may improve the photolytic degradation of ciprofloxacin. Regarding photocatalysis, the degradation rate of ciprofloxacin using unmodified TiO2 was substantially higher than that observed in photolysis (around one magnitude order). This increment may be explained by the sensitizing effect of TiO2 upon the adsorption of ciprofloxacin molecules. According to Paul et al. [53], ciprofloxacin–TiO2 complex is able to activate by visible light irradiation reaching the triplet state; then the excited electrons of the ciprofloxacin triplet migrate to the conduction band of TiO2 and ROS are produced via the transfer of electron to dissolved O2 . As stated above, this mechanism occurs more effectively at neutral pH, thus optimal results would be expected using buffered suspensions. The complete mineralization of ciprofloxacin was achieved only when bi-metallic materials were tested. The mineralization mechanism of ciprofloxacin in bi-metallic nanoparticles may be similar to that described for CO oxidation. In the first step, ciprofloxacin is adsorbed onto the Au nanoparticles; in parallel, due to its

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lower electronegativity, Ag and Cu nanoparticles adsorb dissolved oxygen and transfer the photo-formed electron to O2 producing ROS. Lastly, ROS attack ciprofloxacin molecules to form stable byproducts. Since in most of the identified ciprofloxacin by-products the quinolone structure remained unchanged, these molecules may be adsorbed onto TiO2 and undergo the sensitizing process aforementioned. 4.3. Toxicity tests It is known that fluoroquinolones and its photo-products can express toxic effects [59,60]. However, results obtained in acute toxicity tests showed very low toxicity of ciprofloxacin by-products in both photolysis and photocatalysis experiments. Overall, the biocide effect of ciprofloxacin is expressed by the quinolone moiety, although other components in the molecule contribute to the antibacterial activity. Fluorine atom increases the inhibition of the gyrase enzyme, while the piperazine ring favors antibiotic selectivity to topoisomerases II and IV, and the carboxyl moiety is essential for the intracellular transportation of the fluoroquinolone. During the photolytic process, the reduction of the antibiotic activity may be explained by the substitution of the fluorine atom as well as by the hydroxylation of the carboxyl moiety. On the other hand, the cleavage of the piperazine ring is responsible for the decrease in antibiotic activity in photocatalysis. Due to the modification of these moieties within the ciprofloxacin molecule the antibiotic activity decreased in the early stages of the photocatalytic treatment even when the fluoroquinolone moiety remained unchanged and mineralization was barely occurring. Paul et al. [61] reported that the removal of the antibiotic activity in the early stages of the photocatalytic process using UV-A light displayed a nearly stoichiometric relation with ciprofloxacin degradation; however as by-products appeared the referred stoichiometric relation was lost and slow reduction of the antibiotic activity was observed. It is possible to express the elimination of antibiotic activity in water by the photocatalytic process in terms of energy efficiency. The fluence (the product of incident irradiance and the elapsed reaction time) was calculated to estimate the energy demand for the elimination of antibiotic activity under the experimental conditions. According with our calculations <1.2 kW h/m3 were required to completely remove the antibacterial activity of ciprofloxacin when bi-metallic materials were used. Paul et al. [61] estimated a fluence value of 1.3 kW h/m3 to decrease antibacterial activity, expressed as antibacterial potency, in a magnitude order using bare TiO2 under visible light irradiation. Higher energy efficiencies (i.e. lower energy requirements to reduce the antibacterial activity) were found in this study due to the use of surface-modified materials. 5. Conclusion Surface-modified TiO2 using Au, Ag and Cu mono- and bi-metallic nanoparticles were synthesized by deposition– precipitation method and tested for the photocatalytic degradation of ciprofloxacin. The mono- and bi-metallic materials showed higher activity than unmodified TiO2 in simulated sunlight assays, while when UV-C light was used only bi-metallic materials displayed a notable increment in the photodegradation of the antibiotic compared with bare TiO2 . In experiments using UV-C light, photocatalytic activity can be attributed to the efficient separation of the hole–electron pair and, to lower extent, by the increased adsorption of the antibiotic on the surface of the modified materials. On the other hand, the surface plasmon resonance effect can be the responsible of the increased activity observed for the modified materials in sunlight assays. Bi-metallic modified

photocatalysts showed the highest mineralization percentage at the end of experiments both in UV-C and simulated sunlight assays. Ciprofloxacin by-products maintained antibiotic activity during the photocatalytic process as the fluoroquinolone structure was kept unchanged; although the cleavage of piperazine ring and the substitution of fluorine showed to decrease in some extent the antibiotic activity in early stages of the photocatalytic process. Even though complete mineralization was not achieved by using of the tested materials in sunlight conditions, the obtained effluents showed to be non-toxic and antibiotic activity was fully removed in water. Based on these results, it is concluded that surface-modified TiO2 with bi-metallic nanoparticles at low loads (0.5–1 wt.%) can be used to the complete mineralization of ciprofloxacin in sunlight irradiated systems. Optimization of process parameters, such as pH, may lead to complete removal of residual antibiotic in a lower time. Further investigations will aim in scaling up the photocatalytic process as well as in elucidate the complete reaction mechanisms for these and other organic micro-pollutants. Acknowledgments The authors would like to thank to Dirección General de Asuntos del Personal Académico (DGAPA)-UNAM and to the Consejo Nacional de Ciencia y Tecnología (CONACYT) for funding this work within the framework of projects IN103513 and 194017, respectively. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.cattod.2015.07. 033 References [1] M. Isidori, M. Lavorgna, A. Nardelli, L. Pascarella, A. Parrella, Sci. Total Environ. 346 (2005) 87–98. [2] A.A. Robinson, J.B. Belden, M.J. Lydy, Environ. Toxicol. Chem. 24 (2005) 423–430. [3] C.H. Johansson, L. Janmar, T. Backhaus, Aquat. Toxicol. 156 (2014) 248–258. [4] S. Rodriguez-Mozaz, S. Chamorro, E. Marti, B. Huerta, M. Gros, A. Sànchez-Melsió, C.M. Borrego, D. Barceló, J.L. Balcázar, Water Res. 69 (2015) 234–242. [5] F. Baquero, J.L. Martínez, R. Cantón, Curr. Opin. Biotech. 19 (2008) 260–265. [6] A.J. Watkinson, E.J. Murby, S.D. Costanzo, Water Res. 41 (2007) 4164–4176. [7] N. Le-Minh, S.J. Khan, J.E. Drewes, R.M. Stuetz, Water Res. 44 (2010) 4295–4323. [8] M. Klavarioti, D. Mantzavinos, D. Kassinos, Environ. Int. 35 (2009) 402–417. [9] Y. Ji, C. Ferronato, A. Salvador, X. Yang, J.M. Chovelon, Sci. Total Environ. 472 (2014) 800–808. [10] M.S. Yahya, N. Oturan, K. El Kacemi, M. El Karbane, C.T. Aravindakumar, M.A. Oturan, Chemosphere 117 (2014) 447–454. [11] M. Sturini, A. Speltini, F. Maraschi, A. Profumo, L. Pretali, E.A. Irastorza, E. Fasani, A. Albini, Appl. Catal. B: Environ. 119–120 (2012) 32–39. [12] N. Pugazhenthiran, S. Murugesan, S. Anandan, J. Hazard. Mater. 263 (2013) 541–549. [13] N. Pugazhenthiran, S. Murugesan, P. Sathishkumar, S. Anandan, Chem. Eng. J. 241 (2014) 401–409. [14] G. Lofrano, M. Carotenuto, C.S. Uyguner-Demirel, A. Vitagliano, A. Siciliano, M. Guida, Environ. Technol. 35 (2013) 1234–1242. [15] M.N. Chong, B. Jin, C.W.K. Chow, C. Saint, Water Res. 44 (2010) 2997–3027. [16] M. Pelaez, N.T. Nolan, S.C. Pillai, M.K. Seery, P. Falaras, A.G. Kontos, P.S.M. Dunlop, J.W.J. Hamilton, J.A. Byrne, K. O’Shea, M.H. Entezari, D.D. Dionysiou, Appl. Catal. B: Environ. 125 (2012) 331–349. [17] A.L. Linsebigler, G. Lu, J.T. Yates, Chem. Rev. 95 (1995) 735–758. [18] P. Wang, B. Huang, Y. Dai, M.H. Whangbo, Phys. Chem. Chem. Phys. 14 (2012) 9813–9825. [19] M. Jakob, H. Levanon, P.V. Kamat, Nano Lett. 3 (2003) 353–358. [20] S. Oros-Ruiz, R. Zanella, B. Prado, J. Hazard. Mater. 263 (2013) 28–35. [21] M. Shokri, A. Jodat, N. Modirshahla, M.A. Behnajady, Environ. Technol. 34 (2012) 1161–1166. [22] D. Bu, H. Zhuang, Appl. Surf. Sci. 265 (2013) 677–685. [23] S. Anandan, N. Pugazhenthiran, G.J. Lee, J.J. Wu, J. Mol. Catal. A: Chem. 379 (2013) 112–116.

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Please cite this article in press as: J.C. Durán-Álvarez, et al., Photocatalytic degradation of ciprofloxacin using mono- (Au, Ag and Cu) and bi- (Au–Ag and Au–Cu) metallic nanoparticles supported on TiO2 under UV-C and simulated sunlight, Catal. Today (2015), http://dx.doi.org/10.1016/j.cattod.2015.07.033