Solar Photocaralytic Degradation of Antibiotics: Chemical, Ecotoxicological and Biodegradability Assessment

Solar Photocatalytic Degradation of Antibiotics: Chemical, Ecotoxicological and Biodegradability Assessment A Dissertation to the UNIVERSITY OF PORTO for the degree of Doctor in Environmental Engineering by João Henrique de Oliveira da Silva Pereira Supervisor: Dr. Rui Alfredo Rocha Boaventura Co-Supervisors: Dr. Vítor Jorge Pais Vilar Dr. Maria Teresa Martins Borges [FCUP] Associated Laboratory LSRE-LCM Department of Chemical Engineering Faculty of Engineering University of Porto June, 2014 i Acknowledgments My sincere gratitude goes to Dr. Rui Boaventura and Dr. Vítor Vilar, for the opportunity of developing my work with all the required conditions in their research group (LSRE-FEUP). Their scientific supervision, trust and support, ideas and patience, were essential to the concretization of this thesis. Notwithstanding the physical and “scientific” distance, the supervision of Dr. Maria Teresa Borges was greatly appreciated for many, many reasons. A mention must be made to the following institutions that supported this work: the Foundation for Science and Technology (FCT) (doctoral grant: SFRH/BD/62277/2009); the Laboratory of Separation and Reaction Engineering (LSRE) and Faculty of Engineering of the University of Porto (FEUP), for the technical resources; Coordenação de Aperfeiçoamento de Pessoas de Nível Superior (CAPES) and FCT (CAPES/FCT Proc. 308/11 project) and the University of Porto (AQUAPHOTOBIO - Multidisciplinary project). I also am very thankful for the collaborative work developed with Dr. Olga Nunes (LEPABE), and for the esteemed contribution of Ana Reis. Cheers to Daniel Birra Queirós for his assistance, partnership and patience. To all other people who also gave a hand, in one way or the other, you have not been forgotten. I am also very grateful to Prof. Dr. Santiago Esplugas for having accepted me in his AOP Engineering Group (Universitat de Barcelona, Facultat de Quimica). A special salute goes to Dr. Óscar González Alvarez for a most invaluable guidance during such short, but fruitful stay back in 2010. The other members of their research group will always be fondly remembered for the great moments and friendships, and also Anna May, for the Catalan translation. Fins aviat! To Jordi Bueso, for the treasurable hospitality, and the rest of the Bueso family: moltes gràcies a tots! An extended greeting goes to my esteemed past and present colleagues from LSRE-FEUP, Portuguese, Brazilian, German, Finnish, French, Argentinian, Mexican, Indian, et cetera, for all the good, bad, and overall funny moments spent together, for the coffee break sessions and for all the awkward lunch time conversations and dilemmas. I wish you the best of luck. To the many interesting people I met from all over the world in the scientific meetings I attended: I hope to see you again one day. A kind word is due towards all the good-willed individuals who shared their cherished time and spirit with me along the trodden Way, during these last four and a half years; particularly, Pedro Cunha, Luís Lomba, Cristiana Barbosa and Lívia Xerez. Obrigado, André Monteiro, Sérgio Mota and Rita Félix, pela vossa amizade genuína. Finally, my gratefulness goes to the Source of all things, both finite and infinite. ii iii With love to my parents, brothers, nieces and nephew. iv v To know that you do not know is highest to not know but think you know is flawed Only when one recognizes the fault as a fault can one be without fault The sages are without fault because they recognize the fault as a fault That is why they are without fault Chapter 71 of the Tao Teh King, by Laozi. Translation by Derek Lin. vi vii Abstract The extensive use of human and veterinary antibiotics has resulted in a ubiquitous presence of trace amounts in natural aquatic environments all over the world, since conventional wastewater treatments have been shown unable to remove these highly stable and/or non-biodegradable compounds. This has led to increasing concerns with the risks of potential ecotoxicological effects and of antibiotic resistance propagation in bacterial communities. The development of alternative processes to secure water quality and overall environmental health has thus become a topmost scientific priority. The use of Advanced Oxidation Processes (AOPs) to degrade recalcitrant pollutants such as antibiotics has been showing promising results in recent years. AOPs are characterized by the production of the highly reactive and non-selective hydroxyl radicals (•OH), leading to largely satisfactory results in the mineralization of pollutants to CO2, water and inorganic compounds, or at least in their partial degradation to less harmful and/or more biodegradable compounds. Two of these AOPs, heterogeneous photocatalysis mediated by titanium dioxide (TiO2/UV) and the photo-Fenton process, are regarded as of great interest. Their ability of using naturally available solar radiation as the source of ultraviolet/visible (UV-Vis) radiation greatly reduces energetic costs, while the required catalysts and reactants are fairly inexpensive. Due to the highly efficient use of both direct and diffuse UV solar radiation, Compound parabolic collectors (CPCs) are commonly used as the photoreactors of choice. The main aim of this thesis was the study of the detoxification of three selected antibiotics, Oxytetracycline (OTC), Oxolinic acid (OXA) and Amoxicillin (AMX), by means of these two solar-driven photocatalytic processes. Experiments were performed in lab-scale photocatalytic apparatus provided with a sunlight simulator (Suntest device) and in a solar pilot-plant, both equipped with CPCs. The following objectives were pursued: i) attainment of photocatalytic degradation rate constants with pure antibiotic solutions; ii) assessment of required phototreatment times to achieve antibiotic degradation levels below resistance-inducing concentrations and desirable mineralization; iii) evaluation of the role of photocatalytic process variables; iv) assessment of the influence of individual wastewater components and real matrices and v) development of an alternative multistage treatment combining the biological degradation of AMX by means of an enriched culture with a solar photocatalytic system. Results are compared in terms of required accumulated UV energy per liter of solution, QUV (kJUV L-1). xiv 4 Insights into Solar TiO2-Assisted Photocatalytic Oxidation of Two Antibiotics Employed in Aquatic Animal Production, Oxolinic acid and Oxytetracycline .................75 4.1 Introduction ...................................................................................77 4.2 Materials and Methods .......................................................................79 4.3 Results and discussion .......................................................................80 4.3.1 Solar photolytic and photocatalytic degradation of individual antibiotics .....80 4.3.2 Solar photocatalytic degradation of a mixed OXA and OTC solution ............81 4.3.3 Detailed characterization of the antibiotics degradation .........................82 4.3.3.1 OXA experiment ...................................................................82 4.3.3.2 OTC experiment ...................................................................83 4.3.4 Effects of inorganic ions and scavengers on the photocatalytic efficiency.....85 4.3.5 Solar photocatalytic efficiency index ................................................87 4.4 Conclusions ....................................................................................89 4.5 References .....................................................................................90 5 Assessment of Solar Driven TiO2-Assisted Photocatalysis Efficiency on Amoxicillin Degradation .............................................................................................93 5.1 Introduction ...................................................................................95 5.2 Materials and Methods .......................................................................96 5.3 Results and discussion .......................................................................97 5.3.1 Pilot-scale AMX photolysis and photocatalysis ......................................97 5.3.2 Evaluation of the AMX mineralization ................................................98 5.3.3 Influence of inorganic ions and scavengers ....................................... 101 5.3.4 Solar photocatalytic efficiency index .............................................. 103 5.4 Conclusions .................................................................................. 105 5.5 References ................................................................................... 106 6 Process Intensification at Near Neutral pH of a Homogeneous Photo-Fenton Reaction Using Ferricarboxylate Complexes: Application to Oxytetracycline Degradation ........................................................................................... 109 6.1 Introduction ................................................................................. 111 6.2 Materials and Methods ..................................................................... 113 6.3 Results and discussion ..................................................................... 114 6.3.1 Conventional Fe2+/H2O2/UV-Vis reaction .......................................... 114 6.3.2 Fe3+/Carboxylate/H2O2/UV-Vis OTC degradation reactions ..................... 117 6.3.2.1 Influence of iron concentration ............................................... 117 6.3.2.2 Influence of initial solution pH ................................................ 119 6.3.2.3 Influence of temperature and irradiance .................................... 123 6.3.2.4 Influence of inorganic anions and humic acids .............................. 125 6.3.2.5 Influence of the matrix ......................................................... 127 6.3.3 Solar pilot-plant experiment ........................................................ 128 6.4 Conclusions .................................................................................. 131 6.5 References ................................................................................... 132 7 Biodegradation of Amoxicillin by a Mixed Culture and Oxidation of Metabolic By-products by Solar Photocatalysis ............................................................... 137 7.1 Introduction ................................................................................. 139 7.2 Materials and methods ..................................................................... 141 7.2.1 Reagents ............................................................................... 141 7.2.2 Microbial growth media and conditions ............................................ 141 7.2.3 Culture enrichment ................................................................... 142 7.2.4 Bacteria isolation and identification ............................................... 142 7.2.5 Combined treatment process ....................................................... 142 7.2.5.1 Biological treatment ............................................................ 143 7.2.5.2 Solar Photocatalytic treatment ............................................... 144 xv 7.2.6 Analytical procedures ................................................................ 144 7.3 Results and discussion ..................................................................... 146 7.3.1 Characterization of the mixed culture (MC) ...................................... 146 7.3.2 Combined treatment for AMX removal ............................................. 147 7.3.2.1 Biological degradation step performance .................................... 147 7.3.2.2 TiO2/UV photocatalysis step performance ................................... 149 7.3.2.3 Photo-Fenton step performance .............................................. 151 7.4 Conclusions .................................................................................. 156 7.5 References ................................................................................... 157 8 Main conclusions and future work ............................................................. 161 8.1 Main conclusions ............................................................................ 163 8.1.1 Solar photolysis ........................................................................ 163 8.1.2 TiO2/UV system ....................................................................... 163 8.1.3 Photo-Fenton process ................................................................ 165 8.1.4 Combined treatment of AMX solutions ............................................. 167 8.2 Recommendations for future work ....................................................... 169 xvi xvii Table of Figures Figure 1.1. Sources and possible routes of exposure of antibiotics in the environment. Adapted from Kemper (2008). ................................................................................ 7 Figure 2.1. Molecular structures of a) OTC, b) OXA, c) AMX antibiotics. ............................43 Figure 2.2. SOLARBOX lab-scale experimental set-up: a) schematic representation (adapted from Méndez-Arriaga et al. (2008)); b) and c) views of the thermostatic bath, reservoir tank, peristaltic pump, sunlight simulator and photoreactor equipped with a parabolic reflector .......46 Figure 2.3. SUNTEST lab-scale experimental set-up: a) schematic representation; b) and c) views of the photoreactor equipped with a CPC, the peristaltic pump, the reservoir tank and the sunlight simulator ...................................................................................48 Figure 2.4. Solar pilot-plant experimental set-up: a) plant flowchart; b) front view; c) back view ....51 Figure 3.1. Molecular Structure of Oxytetracycline (OTC) .............................................64 Figure 3.2. OTC (, ), TOC (, ) and pH (, ) monitoring under simulated solar photolysis (open symbols) and photocatalysis with 0.2 g L-1 TiO2 (solid symbols). ..............65 Figure 3.3. Removal profiles of OTC (open symbols) and TOC (solid symbols) under different catalyst loads (,  - 0.1 g L-1 TiO2; ,  - 0.2 g L-1 TiO2; , ▲- 0.5 g L-1 TiO2). ............66 Figure 3.4. Removal profiles of OTC (open symbols) and TOC (closed symbols) with [TiO2] = 0.5 g L-1 and different initial pH values (,  - pH = 3; ,  - free pH; , ▲ - pH = 9; ,  - pH = 11). ............................................................................................67 Figure 3.5. Evolution profiles of Biodegradability () and Inhibition percentage () under 0.5 g L1 TiO2 and free initial pH against OTC () and TOC (▲) removal profiles. .....................68 Figure 3.6. LC-MS-ESI (-) mass spectra of OTC and its degradation by-products at 30 minutes of irradiation ([TiO2] = 0.5 g L-1, free pH, Solarbox experiment) ...................................69 Figure 3.7. Proposed scheme of OTC degradation pathways ([TiO2] = 0.5 g L-1, free pH, Solarbox experiment) .........................................................................................69 Figure 3.8. Removal profiles of OTC and TOC under simulated (,) and real (,) solar photolysis and under simulated (,) and real (▲,) solar photocatalysis with 0.5 g L-1 TiO2 and free initial pH. ............................................................................70 Figure 4.1. a) Normalized absorbance spectra of OXA (blue dotted line) and OTC (black dashed line) at pH = 7.5; solar UV spectrum (yellow solid line) adapted from Malato et al. (2002); b) OXA speciation diagram as a function of pH, including schematics of dissociation equilibrium (pKa value from Jiménez-Lozano et al. (2002). Ionic strength = 0 M, T = 25°C); c) OTC speciation diagram as a function of pH and d) OTC dissociation equilibrium (pKa values from Qiang and Adams (2004), Ionic strength = 0 M, T = 23°C)........................................79 Figure 4.2. Solar photolysis (open symbols) and photocatalysis with 0.5 g L-1 of TiO2 (solid symbols) of 20 mg L-1 OXA and OTC solutions at pH = 7.5: a) OXA concentration (,) and DOC (, ); b) OTC concentration (, ) and DOC (,). ..............................80 Figure 4.3. Removal profiles of OXA () and OTC () concentrations and DOC () evolution in the combined antibiotic solar photocatalytic experiment with 0.5 g L-1 of TiO2 and pH = 7.5 (C0 = 20 mg L-1 each). ..............................................................................81 Figure 4.4. Evolution profiles of OXA concentration (), DOC (), sum of low-molecular-weight carboxylate anions as mg C L-1 (LMWCA - ), total nitrogen (dotted line) and ammonium concentrations as mg N L-1 () and Normalized E. coli growth () under solar photocatalysis with 0.5 g L-1 of TiO2, pH = 7.5 ([OXA]0 = 40 mg L-1). ...........................................83 xviii Figure 4.5. Evolution profiles of OTC concentration (), DOC (), sum of low-molecular-weight carboxylate anions as mg C L-1 (LMWCA - ), total nitrogen concentration (dotted line), ammonium concentration as mg N L-1 () and Normalized E. coli growth () under solar photocatalysis with 0.5 g L-1 of TiO2, pH = 7.5 ([OTC]0 = 40 mg L-1). ...........................84 Figure 4.6. Removal profiles of a) OXA () and b) OTC (), alone and in the presence of 1 g L-1 of Cl- (), SO42- (), NO3-▼, NH4+ (), PO43- (), 0.1 g L-1 HCO3- () under simulated solar photocatalysis with 0.5 g L-1 TiO2, pH = 7.5 ([OXA]0 = [OTC]0 = 20 mg L-1). ...................85 Figure 4.7. Removal profiles of a) OXA () and b) OTC (), alone and in the presence of 10 mM NaN3 () and 50 mM D-mannitol () under simulated solar photocatalysis with 0.5 g L-1 TiO2, pH = 7.5 ([OXA]0 = [OTC]0 = 20 mg L-1). ..................................................87 Figure 5.1. a) Amoxicillin UV absorbance spectrum (dashed line) and solar UV spectrum (solid line) adapted from Malato et al. (2002); b) Antibiotic speciation diagram as a function of pH and c) Schematics of dissociation equilibrium (pKa values from Andreozzi et al. (2005); Ionic strength = 0.1 M, T = 25°C ). .......................................................................97 Figure 5.2. Solar photolysis (open symbols) and photocatalysis with 0.5 g L-1 of TiO2 (solid symbols) of AMX solutions with 20 mg L-1 at pH = 7.5: dimensionless AMX concentration (, ) and DOC (, ). ..........................................................................98 Figure 5.3. Solar photocatalysis ([TiO2] = 0.5 g L-1) of AMX solution with 40 mg L-1: dimensionless AMX concentration (), DOC (), sum of low-molecular-weight carboxylate anions as mg C L-1 (LMWCA, ), sulfate as mg SO42--S L-1 (), Total nitrogen (--) and ammonium as mg N L-1 (). .......................................................................99 Figure 5.4. Escherichia coli (grey columns) and Staphylococcus aureus (white columns) normalized bacterial growth at different phototreatment times under solar photocatalysis with 0.5 g L-1 of TiO2 at pH 7.5 ([AMX]0 = 40 mg L-1), compared to the respective positive control, Cont. (+). .......................................................................................... 101 Figure 5.5. Removal profiles of 20 mg L-1 of AMX alone () and a) in the presence of 1 g L-1 of Cl- (), SO42- (), NO3- (▼), NH4+ (), PO43- (), 0.1 g L-1 HCO3- () and b) in the presence of 10 mM NaN3 () and 50 mM D-mannitol () under simulated solar photocatalysis with 0.5 g L-1 TiO2 and pH = 7.5. ................................................. 102 Figure 6.1. a) Structural formula and b) dissociation equilibrium diagram of OTC (pKa1 = 3.6, pKa2 = 7.58, pKa3 = 9.03, Ionic strength = 0 M, T = 23 °C (Qiang and Adams, 2004)); c) Absorption spectra of 20 mg L-1 OTC: OTC alone (──); OTC + 9.5 mg L-1 oxalic acid (──); OTC + 9.5 mg L-1 oxalic acid + 2 mg L-1 Fe3+ (──); OTC + 2 mg L-1 Fe3+ (─ ─ ─); solar spectrum (──); xenon lamp spectrum (──). .................................................... 114 Figure 6.2. a) Effect of initial pH (■ - pH = 3.0; ● - pH = 4.0; ▲ - pH = 5.0) on the degradation of OTC (C0 = 20 mg L-1) using conventional solar photo-Fenton process mediated by 2 mg L-1 Fe (II). Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m-2, total added H2O2 = 90 mg L-1; b) Speciation diagrams for iron (III) as a function of pH in a solution containing 20 mg L-1 of OTC and 3.58 × 10-2 mM (2 mg L-1) of Fe (III) without accounting (left) or accounting (right) for 1.07 × 10-1 mM (9.5 mg L-1) oxalic acid. Ionic strength = 4 mM. The speciation software MINEQL+ was used to calculate the data. ....................................................... 115 Figure 6.3. Effect of Fe (III) concentration (▼ – 1.0 mg L-1; ▲ – 2.0 mg L-1; ■ – 5.0 mg L-1) on the degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by ferrioxalate (1:3 iron/oxalate molar ratio). Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m2, initial pH unadjusted and total added H2O2 = 90 mg L-1 ....................................... 118 xix Figure 6.4. a) Effect of initial pH (● – pH0 ~ 4.0; ■ - pH0 = 5.0; ▼ - pH0 = 6.0) on the degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1.b) Speciation diagram for iron (III) as a function of pH in a solution containing 1.07 × 10-1 mM (9.5 mg L-1) oxalic acid and 3.58 × 10-2 mM (2 mg L-1) of Fe (III) without accounting (left) or accounting (right) for 10 mM (1 g L-1) SO42-. Ionic strength = 4 mM (left), Ionic strength = 30 mM (right). The speciation software MINEQL+ was used to calculate the data. ....................................................... 121 Figure 6.5. a) Effect of initial pH (■ - pH0 ~ 3.6, ● – pH0 = 5.0) on the degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:1 iron/citrate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1. b) Speciation diagram for iron (III) as a function of pH in a solution containing 3.58×10-2 mM (6.8 mg L-1) citric acid and 3.58×10-2 mM (2 mg L-1) of Fe (III), accounting (left) or not accounting (right) with 20 mg L-1 of OTC. Ionic strength = 4 mM. The speciation software MINEQL+ was used to calculate the data. ................................. 123 Figure 6.6. Degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: pH0 = 5.0 and total added H2O2 = 90 mg L-1. Effect of: a) Initial temperature (● – T = 12 °C; ■ - T = 25 °C; ▲ – T = 35 °C), I = 44 WUV m-2. ................................................................. 124 Figure 6.7. Degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: pH0 = 5.0 and total added H2O2 = 90 mg L-1. Effect of radiation intensity (♦ - I = 24.7 WUV m-2; ▲ – I = 37 WUV m-2; ■ - I = 44 WUV m-2), T = 25 °C. ........................................................ 125 Figure 6.8. Follow-up of OTC degradation (C0 = 20 mg L-1), DOC removal, H2O2 consumption, total dissolved iron and pH in the absence (■) and in the presence of 1 g L-1 of Cl- (●), SO42- (), NO3- (♦), 0.1 g L-1 of HCO3- (▲), 5 mg C L-1 of HA (×) using the solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Process parameters: pH0 = 5.0, T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1. ....................... 126 Figure 6.9. a) Effect of the matrix on OTC degradation (C0 = 20 mg L-1), DOC removal, H2O2 consumption and total dissolved iron using the solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio, performed at pH0 = 4.0 (closed symbols) and pH0 = 5.0 (open symbols). Matrixes: DW (■, □), TF (, ) WW (●, ○). Process parameters: T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1; b) Speciation diagram for iron (III) as a function of pH in WW (left) and TF (right) effluents, solution containing 1.07 × 10-1 mM (9.5 mg L-1) oxalic acid and 3.58 × 10-2 mM (2 mg L-1) of Fe (III). Ionic strength = 3 mM. The speciation software MINEQL+ was used to calculate the data. ............................ 128 Figure 6.10. a) Evolution profiles of OTC degradation (C0 = 20 mg L-1), DOC removal, H2O2 consumption, total dissolved iron and pH using the solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio, performed in the lab-scale photoreactor (■) and in the pilot-plant (●). The sum of LMWCA in the pilot-plant experiment is shown with () and without () initial oxalate concentration. Common process parameters: pH0 = 5.0 and total added H2O2 = 90 mg L-1. Lab-scale experiment process parameters: T = 25 °C, I = 44 WUV m-2; pilot-plant average process parameters: = 26 °C, G = 16 WUV m-2; b) Evolution profiles of detected low-molecular-weigth carboxylate anions during the solar pilot-plant experiment: oxalate (■), oxamate () tartronate (▲), acetate (▼), malonate (♦), maleate () and formate (); c) Normalized biomass yield of E. coli DSM 1103 grown in the presence of different concentrations of OTC standards (upper, grey bars), and in the presence of samples taken at different photo-treatment periods (lower, white bars) of the experiment performed in the pilot-plant. Values represent means and standard deviation (n = 3). A – OTC only, B – OTC and added oxalic acid, C – OTC and added oxalic acid and Fe (III). ...................... 129 xx Figure 7.1. a) Chromatograms obtained by HPLC-UV/Vis analysis at 230 nm. Retention times of the compounds are: AMX: 12.1 min; TP1: 8.2 min and TP2: 6.9 min; b) MS/MS spectrum of Amoxicillin [m/z = 366]; c) MS/MS of Amoxicilloic acid [m/z = 384]. ........................ 141 Figure 7.2. Normalized growth rate in samples taken in the Bio-photo-Fenton combined process in NaCl matrix, at pH0 = 5.0. Values represent means and standard deviation (n = 3). Control (-) and Time 0.0 h are representative of the beginning and the end of the biological step. Time 0.5 h and onwards represent the photo-treatment period. ............................................ 149 Figure 7.3. Follow-up of the Bio-TiO2 combined process on the degradation of AMX (square), its resulting transformation products (TP1 - circle; TP2 - triangle; TP3 – star; TP4 - diamond) and DOC (pentagon), using: a) EM (black symbols) or Buffer (red symbols); b) WW (blue symbols) or Cl (orange symbols). [AMX]0 = 20 mg L-1, Photocatalytic process parameters: [TiO2] = 0.2 g L-1, pH0 = 5.5, T = 25 °C, I = 44 WUV m-2. ....................................... 150 Figure 7.4. Speciation diagrams for iron(III) species as a function of pH in: a) NaCl matrix: without accounting (left) or accounting (right) for 1.07×10-1 mM (9.5 mg L-1) oxalic acid; [Fe (III)] = 3.58×10-2 mM (2 mg L-1), Ionic strength = 0.15 M; and b) WW matrix: without accounting (left) or accounting (right) for 3.22×10-1 mM (29 mg L-1) oxalic acid. [Fe (III)] = 3.58×10-2 mM (2 mg L-1). Ionic strength = 3.3 mM. The speciation software MINEQL+ was used to calculate the data. ................................................................................. 153 Figure 7.5. Evolution profiles of AMX (square) and its transformation products (TP1 - circle; TP2 - triangle; TP3 - diamond) during the Bio-Fe3+/Oxalate/H2O2/UV-Vis combined process performed in a) NaCl matrix, and b) WW matrix. The pH in the photocatalytic step was adjusted to 4.0 (closed symbols) or to 5.0 (open symbols). ...................................... 154 Figure 7.6. Follow-up of DOC removal (square), sum of LMWCA (star), pH (circle), total dissolved iron (triangle) and H2O2 consumption (diamond) during the photocatalytic stage of the Bio--Fe3+/Oxalate/H2O2/UV-Vis combined process performed in: a) NaCl matrix, and b) WW matrix, at pH = 4.0 (closed symbols) or pH = 5.0 (open symbols). Process parameters: [Fe (III)] = 2 mg L-1, initial 1:3 (NaCl) or 1:9 (WW) iron/oxalate molar ratio, total added H2O2 = 90 mg L-1, T = 25°C, I = 44 WUV m-2. +Ox represents extra additions of oxalic acid. ........ 155 xxi List of Tables Table 1.1. Examples of reported levels of some antibiotics in different aquatic media in various countries. ............................................................... 8 Table 1.2. Fundamental TiO2/UV photocatalytic parameters and respective effect on reaction rates. Adapted from Malato et al. (2009). .............................. 14 Table 1.3. Fundamental photo-Fenton process parameters and respective effect on reaction rates. Adapted from Malato et al. (2009). .............................. 16 Table 1.4. Recent applications of solar-driven photocatalytic processes towards the removal of antibiotics from different aquatic media. ........................... 22 Table 2.1. Physico-chemical properties of OTC, OXA and AMX antibiotics. ............ 43 Table 2.2. Pump program for HPLC gradient runs. ......................................... 54 Table 2.3. Analytical parameters of working calibration curves of OTC, OXA and AMX antibiotics. ............................................................................ 54 Table 2.4. Main characteristics of the used matrices (Chapter 7) and tested effluents (Chapter 6, 7). .................................................................. 57 Table 3.1. Elemental compositions and exact mass measurements of OTC and its degradation by-products ([TiO2] = 0.5 g L-1, free pH, Solarbox experiment), using HPLC-MS-ESI(-). ..................................................................... 68 Table 3.2. Kinetic constant values for Solarbox and CPC photolysis and photocatalysis ([TiO2] = 0.5 g L-1, free pH) experiments. ............................ 71 Table 4.1. Pseudo-first-order kinetic parameters for solar photocatalytic degradation experiments of OXA and OTC, [TiO2] = 0.5 g L-1; pH = 7.5. ........... 82 Table 4.2. Pseudo-first-order kinetic parameters simulated solar photocatalytic degradation experiments of OXA and OTC, alone or with (+) inorganic ions and scavengers; [TiO2] = 0.5 g L-1; pH = 7.5; ([OXA]0 = [OTC]0 = 20 mg L-1). ........... 86 Table 5.1. Pseudo-first order kinetic constant values for AMX degradation under solar TiO2-assisted photocatalytic system: [TiO2] = 0.5 g L-1; pH = 7.5. ........... 98 Table 5.2. Pseudo-first order kinetic constant values for AMX degradation, alone or with (+) inorganic ions and scavengers, under simulated solar TiO2-assisted photocatalytic systems: [TiO2] = 0.5 g L-1; pH = 7.5. ............................... 103 Table 6.1. Pseudo-first-order kinetic parameters for the Fe3+/Oxalate/H2O2/UV-Vis process on the degradation of OTC (C0 = 20 mg L-1). Iron/oxalate molar ratio: 1:3. Overall conditions: total added H2O2 = 90 ppm; T = 25 °C; I = 44 WUV m-2. 119 Table 6.2. Main characteristics of the tested effluents, before the OTC-spike step.127 Table 7.1. Main characteristics of the used aqueous matrices before MC inoculation. ............................................................................... 143 Table 7.2. Identification of bacterial strains recovered from the AMX-enriched culture (MC). ............................................................................. 147 Table 7.3. Zero-order kinetic parameters for AMX depletion by the MC in different aqueous matrices. [AMX]0 = 0.02 g L-1; Incubation T = 30 °C; Continuous shaking at 120 rpm; V0 = 1.2 L. All experiments were performed at nearneutral pH. ............................................................................... 148 xxii xxiii Notation Acronyms AA Antibacterial activity AMX Amoxicillin AOP(s) Advanced oxidation process(es) BOD5 Biological oxygen demand (5 days) CAS Conventional activated sludge COD Chemical oxygen demand CPC(s) Compound parabolic pollector DAD Diode array detector DBE Double bond equivalent DOC Dissolved organic carbon EM Enrichment medium ESI Electrospray ionization FBR Fixed-bed reactor HPLC High performance liquid chromatography LC Liquid chromatography LMWCA Low-molecular-weight carboxylate anions LOD Limit of detection LOQ Limit of quantification MBR Membrane bioreactor MC Mixed culture MS Mass spectrometry OTC Oxytetracycline OXA Oxolinic acid PCR Polymerase chain reaction PPCPs Pharmaceuticals and personal care products rRNA Ribosomal ribonucleic acid TiO2 Titanium dioxide TP Transformation products UV Ultraviolet Vis Visible YE Yeast extract WW Wastewater matrix WWTP Wastewater treatment plant Variables Ar Illuminated area (m2) ACO Collector area per order(m2 m-3-order) C Concentration (mg L-1 or mM) i0 Initial value of compound/species/acronym i [ i ] Concentration of compound/species i (mg L-1 or mM) I Irradiance (W m-2) QUV Accumulated UV energy per litre of solution (kJUV L-1) T Temperature (°C) t Time (s, min or h) G UV Average solar ultraviolet irradiance (W m-2) V Volume (L) Chapter 1 6 1.2 The problem of antibiotic residues in the environment Human and veterinary pharmaceutical chemicals are widely used for diagnosis, treatment, alteration or prevention of disease and other health conditions, and include a broad class of substances such as analgesics, antibiotics, lipid lowering agents, hormones and other endocrine disrupting compounds. They are distinguished by their functionalities, physico-chemical and biological properties (Kümmerer, 2001), and are designed to perform a certain biological activity on human beings, animals, bacteria or other organisms (Halling-Sørensen et al., 1998). The contamination of the environment by Pharmaceuticals and Personal Care Products (PPCPs) as a result of metabolic excretion, improper disposal and/or industrial waste has been the subject of special attention over recent years, as reflected by the increasing literature regarding their sources, fate and biological effects, as well as studies on their removal and degradation from different environmental matrices. Such works range from general reviews, approaches to environmental risk assessment, measurement and detection in effluents from wastewater treatment plants (WWTPs) and assorted environmental compartments, development of analytical methods of extraction and identification and, finally, different processes of removal and degradation of PPCPs (Halling-Sørensen et al., 1998; Andreozzi et al., 1999; Adams et al., 2002; Heberer, 2002; Kolpin et al., 2002; Fent et al., 2006; Seifrtová et al., 2009). Amongst all PPCPs, antibiotics are a diverse group of medical substances whose antibacterial, anti-fungal or anti-parasitical properties play a major role in modern medicine, both human and veterinary. They can be either derived from certain microorganisms or obtained by chemical synthesis. A general classification is made grouping antibiotics by chemical structure or by mechanism of action (Kümmerer, 2009). For instance, β-lactams inhibit cell wall synthesis, Tetracyclines and Macrolides inhibit protein synthesis, Quinolones interfere with nucleic acid metabolism and Sulfonamides act as competitive inhibitors of growth factors and other metabolites (Son et al., 2009). Depending on the mechanism of action, antibiotics can be bactericidal (they kill bacteria) or bacteriostatic (they inhibit bacterial reproduction or growth). Antibiotic usage for human and veterinary consumption varies significantly by groups and by country (Mölstad et al., 2002; Sarmah et al., 2006), but information on the values of production and usage is still not widely available. A few international comparisons have been made so far (Cars et al., 2001). Goossens and co-workers (Goossens et al., 2007), under the European Surveillance of Antimicrobial Consumption (ESAC) project, have shown that the outpatient use of Tetracyclines, Macrolides and Quinolones antibiotics in the United States of America is Chapter 1 7 higher than in any country of Europe, where the β-lactams make up the largest share. Regarding veterinary antibiotics, a review by Sarmah and co-workers (2006) has collected values of use and production on a global scale, showing Ionophores, a veterinary-only class of antibiotics, to represent the largest class by reported use in the United States and New Zealand. On a global level, human antibiotics are consumed domestically, in clinics or in hospitals, whereas veterinary antibiotics are generally used in livestock therapeutics and fish feeds. As they are only partially metabolized in the organism (Hirsch et al., 1999), they are excreted (together with their metabolites), to receiving WWTPs and, ultimately, to other environmental compartments (Hektoen et al., 1995; Thiele-Bruhn, 2003; Jones et al., 2005b). The degree to which antibiotics are metabolized in human and animal bodies varies substantially, but when the amounts used are multiplied by excretion rates, even those with high metabolization rates can be important (Kümmerer, 2009). On a local level, effluents from drug production facilities are also considered an important source of antibiotics (Larsson et al., 2007; Li et al., 2008a; Li et al., 2008b). Sorption of non-biodegradable antibiotics in the biomass of WWTPs may also become an issue when desorption conditions change during land application of biosolids, increasing antibiotic bioavailability (Kim et al., 2005). Thus, with their release to the environment (Figure 1.1) and with the improvement of detection methods, antibiotics have been measured on different matrices all over the world (Table 1.1). Figure 1.1. Sources and possible routes of exposure of antibiotics in the environment. Adapted from Kemper (2008). Concentrations found are generally low (Table 1.1) but nevertheless, a continuous introduction can offset antibiotic natural transformation and removal rates (Jones et al., 2005a), raising concerns about more subtle changes. On the one hand, acute environmental impacts of most studied substances may be unlikely (Baguer et al., 2000; Boxall et al., 2003), but on the other hand, several studies have shown the development of resistance in environmental bacteria Veterinary use Human use Animal Farms Aquaculture WWTP Sediments Soils Surface waterGround water Manufacturers Landfills Ind. WWTP Propagation of antibiotic resistance genes and ecotoxicological effects Chapter 1 8 (McKeon et al., 1995; Morris and Masterton, 2002; Kümmerer, 2004; Baquero et al., 2008; Dantas et al., 2008; Martinez, 2009), effects on activated sludge bacteria (Halling-Sørensen et al., 2002; Cunningham et al., 2006; Alighardashi et al., 2009) and on algal communities (Halling-Sørensen, 2000; Wilson et al., 2003). Finally, a possible cumulative exposure to human beings in drinking water is especially worrisome (Webb et al., 2003; Collier, 2007; Vaz-Moreira et al., 2011). The importance of developing practical and effective wastewater treatment processes to avert antibiotic pollution is thus of the utmost importance to assure the quality of aquatic environments. Table 1.1. Examples of reported levels of some antibiotics in different aquatic media in various countries. Group Antibiotic Concentration (μg L-1) Compartment Location References β-Lactams Amoxicillin 0.280a 0.030b Domestic WWTP Australia Watkinson et al. (2007) Penicillin G 0.153a 0.002b Production facility WWTP China Li et al. (2008b) Macrolides Clarithromycin 0.059 – 1.433a 0.012 - 0.232b Domestic WWTP Taiwan Lin et al. (2009) Erythromycin 0.113a 0.290b Domestic WWTP U.K. Roberts and Thomas (2006) 0.070 Surface water Quinolones Ciprofloxacin 0.017 – 2.500a 0.022 – 0.620b Domestic WWTP Canada Guerra et al. (2014) Norfloxacin 0.059a 0.013b Domestic WWTP China Li et al. (2009) Oxolinic Acid 10 - 2500 Shrimp pond water Vietnam Le and Munekage (2004) Sulfonamides Sulfamethoxazole 0.013 - 0.155a 0.004 - 0.039b Domestic WWTP Luxembourg Pailler et al. (2009) 0.001 - 0.022 Surface water Sulfamethazine 0.29a 0.036b Domestic WWTP U.S.A. Karthikeyan and Meyer (2006) Tetracyclines Tetracycline 42.2 – 158 23.2 – 29.2a Hospital wastewater Domestic WTTP Portugal Pena et al. (2010) Oxytetracycline 0.07 – 1.34 Surface water U.S.A. Lindsey et al. (2001) 19.5 – 920 (×103)b 235 – 484 OTC production facility WWTP Receiving river China Li et al. (2008a) aInfluent, bEffluent 1.3 Removal of antibiotics by conventional and advanced treatments In conventional WWTPs, biological degradation is the main process associated to the elimination of micropollutants such as antibiotics. Physical steps such as sorption on sludge or particulate matter, filtration or stripping simply alter the phase in which antibiotics are present Chapter 1 9 (Larsen et al., 2004; Kim et al., 2005). Still, in accordance to studies that show how many of these substances fail to be readily biodegradable in simulated conditions, it is expected that certain antibiotics persist through conventional wastewater treatment systems conditions (Kümmerer et al., 2000; Ingerslev et al., 2001; Drillia et al., 2005). Several authors also note the importance of studying transformation products (TPs) resulting from biological conversion of parent compounds, which may be more stable in the environment (Lamm et al., 2009; Tambosi et al., 2010; Pérez-Parada et al., 2011). Le-Minh et al. (2010) dedicated a review to summarize the most important factors affecting the reported varying efficiencies of the removal by conventional and advanced treatments methods of the different classes of antibiotics. For instance, β-lactam antibiotics are highly susceptible to chemical and biochemical hydrolysis of the β-lactam ring during biological treatments while the capacity of activated carbon to adsorb particular compounds depends on the hydrophobic (nonpolar) or hydrophilic (polar) nature of the chemical. The removal efficiencies may even vary between antibiotics belonging to the same class, i.e., presenting similar molecular structure and physic-chemical properties, as seen on a work by García Galán et al. (2012) reporting on the removal of several sulfonamide antibiotics upon conventional activated sludge (CAS) and advanced membrane bioreactor (MBR). Miège and co-workers (2009) published a database on the fate of PPCPs in WWTPs. For the seven most cited antibiotics over 117 papers involving 184 molecules, mean removal efficiencies ranged from 18 % (Trimethoprim) to 80 % (Norfloxacin). In a reported study (Göbel et al., 2007), the elimination of detected antibiotics in the raw wastewater of two WWTPs (two sulfonamides, four macrolides and trimethoprim) was studied through CAS systems coupled with a fixed-bed reactor (FBR) and MBR, respectively. Removal in primary treatment (sand filter) was generally low and considered as not significant for all antibiotics. As to the secondary effluents of CAS systems and FBR, the two sulfonamides inconsistently showed either high positive or negative elimination values, suggesting a possible retransformation between their main metabolites. Trimethoprim showed only a slight elimination of up to 20 %, and varying results, including negative values, were obtained for the studied macrolides (-20 to 20 %). This is in contrast to the results of the CAS system coupled with the MBR, which not only showed no increase in the load of any antibiotic but also a higher tendency of elimination. Despite this, full removal was not obtained for any molecule. Camacho-Muñoz et al. (2012) compared the effectiveness of conventional suspended wastewater treatments (activated sludge and oxidation ditches) and low-cost treatments (trickling filters, anaerobic lagoons and constructed wetlands) on the removal of several PPCPs, including the antibiotics Sulfamethoxazole and Trimethoprim, detected in influents from 11 Chapter 1 10 urban WWTPs. Reported removal rates of these antibiotics were as high as 99 %, while mean removal rates of other PPCPs were 64 and 55 % for conventional and low-cost techniques, respectively. Nevertheless, most compounds were still detected in effluent wastewater. Adams et al. (2002) determined the effectiveness of common drinking water treatment processes in the removal of seven common antibiotics. Powdered activated carbon, reverse osmosis and oxidation with chlorine and ozone were shown to be effective in removing over 90% of each compound from both distilled and river water. In contrast, coagulation, flocculation and sedimentation with alum and iron salts, excess lime/soda ash softening, ultraviolet (UV) irradiation at disinfection dosages and ion exchange were not. Moreover, Rizzo and co-workers (2013) remarked that the common UV disinfection process is inappropriate for assorted antibiotic removal, given that not all antibiotics present UV absorption spectra which overlap with the UV lamp emission (peak at 254 nm). Choi et al. (2008) also evaluated the treatment of seven Tetracycline antibiotics from raw waters by coagulation (poly-aluminum chloride as coagulant) and adsorption (granular activated carbon (GAC) filter). Efficiency of coagulation removal ranged from 43 to 94%, depending on the type of tetracycline, at optimum conditions, from synthetic water, but it slightly decreased (44~67%) in river water due to organic interference, notwithstanding an insignificant difference between removal efficiencies. On the other hand, GAC filtration showed to be more effective, removing more than 68% of incoming tetracyclines, with general removal efficiencies above 90%. A coupling of both techniques was suggested to improve tetracycline removal. Two wide-ranging studies by Rivera-Utrilla discussed the removal from water of nitroimidazoles (2009) and tetracycline antibiotics (2013a) by adsorption/biosorption on activated carbons and sludge-derived adsorbents. They addressed the role of the chemical properties of the different activated carbons and of the solution pH and also the influence of the presence of electrolytes and bacteria, matrix effects and different regimes on adsorption rates. In 2011, the same authors published an overview on activated carbon modifications to enhance their water treatment applications (Rivera-Utrilla et al., 2011). Regardless of the advantage of not generating toxic nor pharmaceutically active products, the major drawback of activated carbon adsorption is that concerning the transference of the contaminants to a new phase, concentrating them (Daghrir and Drogui, 2013; Rivera-Utrilla et al., 2013b). Chamberlain and Adams (2006) reported the application of free chlorination and monochloramination for the oxidation of antibiotics (Carbadox and sulfonamides, macrolides) in surface waters in laboratory under conditions similar to drinking water treatment. Chlorination readily removed sulfonamides at near neutral pH levels, whereas for macrolides, Chapter 1 11 only partial removal was obtained. Little removal of both antibiotic classes was observed with monochloramination at any condition. However, for Carbadox, both processes showed fast reactions and a near complete removal of the parent compound would be expected. However, specific oxidation byproducts were not analyzed. Several authors contend the application of chlorination processes to treat micropollutants, since thorough information regarding the formation and fate of harmful chlorinated byproducts is still lacking (Le-Minh et al., 2010; Oncu and Balcioglu, 2013). Tambosi et al. (2010) underlined the role of biodegradation in the removal of three antibiotics, amongst other PPCPs, in two different MBR set-ups, compared to sludge sorption or physical retention in the membranes. The antibacterial properties of each compound are suggested to account for differences in antibiotic removal efficiency. Conversely, Radjenovic et al. (2007) and Radjenović et al. (2009) found varying removal rates of PPCPs (including the antibiotics Erythromycin, Sulfamethoxazole, Ofloxacin and Trimethoprim) when comparing CAS system to MBR treatment. As in the abovementioned studies, these authors bring up the fact that MBR processes would not completely halt discharge of micropollutants and reckon that, although a promising technology, MBR processes still require optimization of design and operational conditions to overcome the incomplete removal of antibiotics. An overview of removal of pharmaceuticals with MBRs technology by Sipma and co-workers (2010) also supports this conclusion, while Larsen et al. (2004) also specified that MBR have high material costs and energy demands, albeit having the advantage of smaller space requirements and of increasing solids retention time. Rejection of trace pollutants by ultrafiltration (UF), nanofiltration (NF) and/or reverse osmosis (RO) membrane systems have been showing good overall removal results (Li et al., 2004; Košutić et al., 2007; Snyder et al., 2007; Yoon et al., 2007; Koyuncu et al., 2008; Radjenović et al., 2008), but major drawbacks result from expensive membrane disposal and substitution, high energy and operation requirements and possible greater toxicity levels in the brine compared to the influent water (Snyder et al., 2007). Busetti and Heitz (2011) provides an example of the efficiency of a microfiltration-reverse osmosis treatment integrated in a full scale operational water reclamation plant in the removal of nitroimidaozle, sulfonamide, lincosamide and macrolide antibiotics detected in secondary treated effluents. The reported estimated membrane rejection was generally higher than 91%. In recent years, the application of Advanced Oxidation Processes (AOPs) to treat wastewaters contaminated with components that have high chemical stability and/or low biodegradability, such as pesticides or pharmaceuticals, has been subject to intensive research. For this reason, the following section will deal entirely with AOPs. Chapter 1 12 1.4 Removal of antibiotics by Advanced Oxidation Processes AOPs comprise different processes of generating the highly reactive and non-selective hydroxyl radicals (•OH). Malato et al. (2003) concisely enunciates that •OH radicals are the second strongest known oxidant after fluoride (Eº(•OH/H2O) = 2.80 V/SHE), and that kinetic rate constants for most reactions fall in the order of 106 to 109 M-1 s-1. The reactions through which they attack organic molecules can be hydrogen abstraction, electrophilic addition, electron transfer and also radical-radical reactions, citing Legrini et al. (1993) in the abovementioned work. A recent work by Wols and Hofman-Caris (2012) points out the very high •OH radical rate constants for a wide range of organic micropollutants, including antibiotics. The classification of AOPs can be divided as follows: photochemical (UV/O3, UV/H2O2, UV/H2O2/O3), photocatalytic (TiO2/UV, Photo-Fenton) or chemical oxidation processes (O3, O3/H2O2, H2O2/Fe2+) (Poyatos et al., 2009). The characteristics and mechanisms of these processes will not be discussed here, as there are plenty of detailed studies regarding general and particular aspects of each technique or combination thereof (Andreozzi et al., 1999; Huber et al., 2003; Malato et al., 2003; Gogate and Pandit, 2004a; b; Pignatello et al., 2006; Poyatos et al., 2009). Comprehensive reviews covering the application of AOPs to aqueous pharmaceuticals are also available (Ikehata et al., 2006; Dalrymple et al., 2007; Esplugas et al., 2007; Kanakaraju et al., 2013), as well as specific reviews dealing with antibiotic removal exclusively (Homem and Santos, 2011; Michael et al., 2013b; Oncu and Balcioglu, 2013). Fatta-Kassinos et al. (2011) published a pertinent review on the often neglected subject of the significance of the resulting transformation by-products. It compiles information concerning the identification of TPs formed during the application of natural photolytic and AOPs and the respective potential biological effects. It also presents a critical view on the discrepancies and differences between published experimental configurations for photo-driven (advanced oxidation) processes, which prevent an otherwise uniform comparison of data and information relevant to real environmental conditions. Of special importance is the ability of AOPs to achieve the complete mineralization of this kind of pollutants, yielding CO2, water and inorganic compounds, or at least a partial decomposition to more biodegradable and/or less harmful intermediates. The later would allow for a useful and cost efficient combination with biological processes (Marco et al., 1997; Schaar et al., 2010; Oller et al., 2011). Considering the current limitations of implementing these processes in existing WWTP (high flow rates, capital and reactant costs, catalyst separation step, for example), the combination of membrane processes with AOPs have also been proposed to Chapter 1 13 optimize wastewater treatment (Westerhoff et al., 2009; Senta et al., 2011; Liu et al., 2014). At last, with respect to high energetic costs of implementing UV lamp-based wastewater treatments (Rosenfeldt et al., 2006), the AOPs relying on solar irradiation, such as heterogeneous photocatalysis mediated by TiO2/UV and the Photo-Fenton reaction, are considered the most promising and environmental friendly technologies (Muñoz et al., 2006). The use of Compound Parabolic Collectors (CPCs) greatly enhances the efficiency of these solar photocatalytic processes as it increases the amount of incident solar UV photons, both direct and diffuse, that can be used to degrade target substances (Rodríguez et al., 2004; ColinaMárquez et al., 2010). Bandala and Estrada (2007) performed a comparative study between four types of solar collectors using oxalic acid and the pesticide carbaryl as model contaminants, employing TiO2 as photocatalyst. Compound parabolic collector geometry demonstrated the highest turnover rate in the photocatalytic degradation of both target compounds, followed by V-shaped trough collector, parabolic concentrator and tubular collector. An in-depth review put forth by Malato and co-workers (2009) expounds on the use of sunlight to produce •OH radicals by means of these two solar-driven processes, describing the influence of fundamental parameters, the analytical and toxicological tools, the necessary hardware, photocatalyst enhancement techniques, treatment integration with other AOPs and/or biodegradation. Both solar heterogeneous photocatalysis mediated by TiO2/UV and solar Photo-Fenton process will be briefly addressed in the following sub-sections. 1.4.1 Solar TiO2/UV photocatalysis Heterogeneous photocatalysis using suspended TiO2 is of special interest due to the chemical stability of the photocatalyst, low cost and ability of using the small percentage of the ultraviolet radiation coming from the sun. Monteiro et al. (2014) described the mechanism as follows. The absorption by the the semiconductor (TiO2) of incident photons of energy matching or exceeding the semiconductor band-gap energy produces conduction-band electrons cb (TiO2) and valenceband holes vb (TiO2), i.e. electron-hole pairs (Eq. 1.01). Once at the surface of the semiconductor, the presence of as suitable acceptor (for cb ) and donor (for vb ) will avoid the near instantaneous and undesirable generated recombination (Linsebigler et al., 1995; Furube et al., 2001). Hydroxyl anions and water molecules adsorbed on TiO2 surface act as electron donors, while molecular oxygen acts as electron acceptor, leading to the formation of hydroxyl (•OH) and superoxide (O2 ) radicals (Peral and Ollis, 1992; Pelizzetti and Chapter 1 14 Minero, 1993; Augugliaro et al., 1999) (see eq. 1.02-1.04). When an organic molecule (RH) is adsorbed onto semiconductor surface, the reaction with hydroxyl radical occur, followed by structural breakdown into several intermediates until, eventually, total mineralization (see eq. 1.05) (Hoffmann et al., 1995; Kolen'ko et al., 2005). Due to their high oxidation potential, the photogenerated holes can also participate in the direct oxidation of the organic pollutants (eq. 1.06) (Cermenati et al., 1997; Benoit-Marquié et al., 2000). A Peroxide (HOO ) radical can also be generated from the protonation of O2 radical and subsequently form hydrogen peroxide (see eq. 1.07-1.08).     222 TiOhTiOehTiO VBCB    (1.01)     HOHOHTiOh adsVB 22 (1.02)   OHOHTiOh adsVB   2 (1.03)     )(2)(22 adsadsCB OOTiOe (1.04)   HRRHRHHO (1.05)   HRRHRHTiOhVB )( 2 (1.06)   HOOHO2 (1.07) 2222 OOHHHOOO  (10.8) Table 1.2 summarizes the parameters and their influence on the photocatalytic rate kinetics, based on the abovementioned review by Malato et al. (2009). Table 1.2. Fundamental TiO2/UV photocatalytic parameters and respective effect on reaction rates. Adapted from Malato et al. (2009). Parameter Influence or effect on reaction rates Initial pollutant concentration (C0) Most reactions follow pseudo-first order kinetics (C = C0×e-kt), so maximum efficiency would be attained at saturation level of the catalyst surface. Catalyst load ([TiO2]) Reaction rates increase with increasing catalyst load, up until a point in which, depending on reactor geometry and experimental conditions, additional catalyst particles block the penetration of incident UV light. pH Influences the pollutant adsorption onto the catalyst surface and the catalyst particles aggregation. Temperature Irrelevant influence in the range of 20 to 80 °C. Irradiance Only wavelengths up to 390 nm are useful (~ 5% solar spectrum). Reaction rate is proportional to the radiant flux (Φ), but high values should be avoided because electron-hole recombination is favored. O2 concentration No mineralization is possible without O2, while reaction rates increase with increasing dissolved oxygen concentration (up to a certain level). Chapter 1 15 1.4.2 Solar photo-Fenton process Photo-Fenton comprises the combination of ferrous iron (Fe2+) with hydrogen peroxide (H2O2) and (solar) UV-Vis radiation resulting in the production of two moles of OH per mole of hydrogen peroxide (Eq. 1.09 and 1.10), as simplified by Gogate and Pandit (2004b):   HOOHFeOHFe 3 22 2 (1.09)   OHFehOHFe   2 2 (1.10) Pignatello et al. (2006) summarizes the reasons for the optimum operational pH value of the (photo-) Fenton process around 3 as follows: first, the solubility of Fe3+-hydroxy complexes decreases for pH values above 3; second, [Fe(OH)]2+, the most photoactive species (with absorption bands between 290 and 400 nm), reaches its maximum molar fraction around the aforementioned pH. Consequently, there is a limit in the application of this process in industrial scale due to the costs associated with pH corrections (initial acidification and final neutralization). The formation of complexes between Fe (III) and carboxylate ions is pointed out as the most viable way to overcome this liability. In this way, the photo-Fenton process is improved by extending the solubility of iron to higher and more practical pH values, by presenting stronger radiation absorption at wavelengths until 580 nm and by increasing the quantum yield of Fe2+ production according to Eq. 1.11 (Jeong and Yoon, 2005; Pignatello et al., 2006).        2 42 2 42 2 23 42 3)1( OCOCnFehOCFe n n (1.11) Ferricarboxylate-mediated solar photo-Fenton has already been successfully applied to treat different wastewaters and specific pollutants, whereby carboxylate ions such as oxalate, citrate and EDDS (ethylenediamine-N, N’-disuccinic acid) were used to form complexes with Fe3+ (Silva et al., 2007; Prato-Garcia et al., 2009; Rodríguez et al., 2009; Huang et al., 2012; Monteagudo et al., 2012). For a second time, the work by Malato and co-workers (2009) will be based upon to summarize the main photo-Fenton process parameters and their influence on photocatalytic rate kinetics (Table 1.3). Chapter 1 22 Table 1.4. Recent applications of solar-driven photocatalytic processes towards the removal of antibiotics from different aquatic media. Antibiotics Process Operating conditions Results References Oxytetracycline (OTC) UV, UV/TiO2 [OTC] = 5 - 40 mg L-1 Matrix: Ultrapure water NF-doped TiO2 film pH = 2.0 - 11.0; Broadband light intensity: Visible: 0.399 mW cm-2, Solar: 0.475 mW cm-2; Radiation blocked below 420 nm. OTC degradation largely influenced by the solution pH, which determines the different electric charge state of OTC species. With increasing pH, the light absorption of OTC exhibits red shift to the visible light while the degradation rate of OTC by photolysis under solar/visible light is significantly accelerated. Photocatalytic degradation suggests 5 pathways: direct photolytic degradation, UV/Vis light-induced photocatalytic oxidation and reduction, and visible light-induced OTC self-photosensitized oxidation and reduction. Zhao et al. (2013) Tetracycline (TC) UV/TiO2 [TC] = 40 mg L-1 Matrix: Deionized water [TiO2] = 0.5 g L-1 Light intensity (360 nm): UV lamp: 1.210 mW cm-2 Solarium: 1.980 mW cm-2 UV-A lamp: 0.059 mW cm-2 Antibiotic half-life: 10 (UV lamp), 20 (solarium) and 120 min (UV-A lamp) TOC removal after 120 min: 90%, 75%, 12%, respectively. BOD5/COD ratio from 0.45 to 0.85 after 120 min (solarium device). Full antibacterial activity inhibition at 55 and 70 min (solarium, UV-lamp); only 15% reduction after 120 min (UV-A lamp), Reyes et al. (2006) Tetracycline (TC) UV/TiO2, UV/ZnO [TC] = 20 mg L-1 Matrix: Deionized water [TiO2] = 0.5 - 1.5 g L-1; pH = 3 - 10 [ZnO] = 0.2 - 1.5 g L-1; pH = 6 - 11 Light intensity (300–800 nm): Xe lamp: 250 W m-2 Optimal oxidation conditions: [TiO2] = 1.5 g L-1, pH = 8.7; [ZnO] = 1.0 g L-1; pH = 11. Complete antibiotic removal after 15 min with both photocatalysts. TOC removal after 60 min: 70 % (TiO2), 100 % (ZnO). Full antibacterial activity inhibition at 30 min with TiO2. Palominos et al. (2009) Tetracycline (TC) Photo-Fenton [TC] = 24 mg L-1 Matrix: Ultrapure water, Surface water (SW), Effluent from WWTP (WW) [FeOx], [Fe(NO3)3] = 0.20 mmol L-1 [H2O2] = 1 - 10 mmol L-1 pH = 2.5 Broadband light intensity: Black-light: 19 W m-2 Solar: 15 to 20 W m-2 Black-light results: Complete TC degradation: 1-1.5 min (Fe(NO3)3), 8 min (FeOx). Similar residual TOC concentration after 60 min (4 mg L−1). Similar efficiency in SW matrix, substantial loss of efficiency in WW matrix Solar light results: Complete antibiotic removal: 0.5 min/0.054 J cm-2 UV dose with FeOx against 3 min (0.270 J cm-2 UV dose) with Fe(NO3)3. Similar residual TOC concentration (2 mg L-1) after 60 min/5.870 J cm-2 UV dose No interference of SW or WWTP matrices. Bautitz and Nogueira (2007) Chapter 1 23 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References Oxolinic Acid (OXA) UV/TiO2 [OXA] = 20 mg L-1 Matrix: Ultrapure water [TiO2] = 0.2 - 1.5 g L-1; pH = 7.5 - 11 Light intensity (max at 365 nm): Black light: 14 W m-2 Optimal oxidation conditions: [TiO2] = 1.0 g L-1, pH = 7.5 Complete antibiotic removal after 30 min. TOC and COD removal after 60 min: 50% Full antibacterial activity inhibition at 60 min. Initial toxicity (V. fischeri) reduced by ~70% after 60 min. Giraldo et al. (2010) UV/TiO2 [OXA] = 18 mg L-1 Matrix: Ultrapure water Immobilized TiO2 in sintered glass cylinders pH = 9 Light intensity (max at 365 nm): Black light: 14 W m-2 Complete antibiotic removal after 120 min. Similar COD removal profile. TOC removal after 100 min: 50% Average oxidation state (AOS): -1.5 (0 min), +2 (40 min), +3 (120 min) Full antibacterial activity inhibition at 60 min. Palominos et al. (2008) Mixture of Amoxicilin (AMX), Ampicillin (AMP), Cloxacillin (CLX) UV/TiO2 [AMX], [AMP], [CLX] ~100 mg L-1 Matrix: Deionized water [TiO2] = 0.5 – 2.0 g L-1; pH = 3 - 11 [H2O2] = 0 – 300 mg L-1 UVA lamp (365 nm) Nominal power: 6W With [TiO2] = 1.0 g L-1, pH = 11, [H2O2] = 0 mg L-1, after 300 min: 70.9 % AMX, 91.4 % AMP, 100 % CLX removal 11.2 % COD and 3% DOC removal With [TiO2] = 1.0 g L-1, pH = 5, [H2O2] = 100 mg L-1 Complete AMX and CLX removal after 20 mins Complete AMP removal after 30 min 26.3 % COD and 13.9% DOC removal after 300 min Elmolla et al. (2010) Amoxicillin (AMX) UV/TiO2 [AMX] = 1 -100 mg L-1 Matrix: Deionized water [TiO2] = 1.0 g L-1 Carbonand iron-doped titania pH = 3 - 9 Irradiance: UV-lamp (365 nm): 0.5 mW cm-2 Artificial daylight : 0.6 mW cm-2 Solar radiation: 16 mW cm-2 UV-lamp results after 6 h: Best pH = 6.0; AMX conversion rates increase from 1 to 25 mg L-1, decrease afterwards due to mass transfer limitation; COD removal between 10 and 40 %. Artificial daylight results after 6 h (AMX0 = 25 mg L-1, pH = 6.0): Regular TiO2: ~80 % AMX removal Carbon-doped titania (37 at.% C, 200 °C): 30 % AMX removal Iron-doped titania (0.89 at.% Fe, 200 °C): 25 % AMX removal COD removal between 10 and 30 % After 2 h with solar radiation (AMX0 = 25 mg L-1, pH = 6.0): Regular TiO2: > 80 % AMX removal Release of N: 1.5 % and of S: 14 % COD removal between 10 and 30 % Carbon-doped titania (37 at.% C, 200 °C): ~75 % AMX removal Iron-doped titania (2.2 at.% Fe, 200 °C): ~75 % AMX removal Klauson et al. (2010) Chapter 1 24 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References Amoxicillin (AMX) UV/TiO2 [AMX] = 2.5 -30 mg L-1 Matrix: Ultrapure water WWTP effluent [TiO2] = 0.1 – 0.75 g L-1 pH = 5 or 7.5 UVA lamp (350 - 400 nm) Photon flux: 8 × 10-4 E L-1 min -1 Effect of catalyst loading (gTiO2 L-1) after 20 min (AMX0 =10 mg L-1, pH = 5.0): AMX removal: 100 %, except at 0.1 g L-1 (65 %) DOC removal: 43, 60, 65 and 70 % at 0.1, 0.25, 0.5 and 0.75 gTiO2 L-1, respectively Effect of initial pH (AMX0 = 10 mg L-1, 0.5 gTiO2 L-1): AMX removal after 25 min: 100 % with both pH; TOC removal after 90 min: 93 versus 75 % with pH = 5 and pH = 7.5, respectively. Effect of water matrix (AMX0 = 10 mg L-1, 0.5 gTiO2 L-1, pH = 7.5): Ultrapure water: AMX removed after 25 min WWTP effluent: AMX removed after60 min Antibacterial properties: Depends on AMX concentration and the test bacteria in question; Escherichia coli and Klebsiella pneumonia are affected only by AMX, while Enterococcus faecalis is also partially affected by its reaction by-products Dimitrikapoulou et al. (2012) Amoxicillin (AMX) Fenton [AMX] = 10 -200 mg L-1 Matrix: Ultrapure water [Fe(II)] = 0 – 50 mg L-1 [H2O2] = 10 – 500 mg L-1 pH = 3.5 Optimized H2O2/Fe/AMX ratio: 255/25/105 mg L-1 AMX removal after 2.5 min: 100% TOC removal after 15 min: 37% Ay and Kargi (2010) Amoxicillin (AMX) Photo-Fenton [AMX] = 10 -200 mg L-1 Matrix: Ultrapure water [Fe(II)] = 0 – 50 mg L-1 [H2O2] = 10 – 500 mg L-1 pH = 3.5 UV lamp (254 nm): Light intensity: 4.98 × 10-6 E s-1 Optimized H2O2/Fe/AMX ratio: 100/40/105 mg L-1 AMX removal after 2.5 min: 100% TOC removal after 60 min: 53% Ay and Kargi (2011) Chapter 1 25 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References Mixture of AMX with other pharmaceuticals UV H2O2/FeII(/UV) H2O2/FeIII(/UV) UV/H2O2 UV/TiO2 O3(+UV), O3/H2O2(+UV) O3/TiO2(+UV) [AMX] = 1 µM ~ 0.36 mg L-1 Matrix: Ultrapure water, Surface water Groundwater Effluents 1, 2 from WWTP pH = 3 (UV, Fenton-based trials) = 7 (reminder) pH = 8.1 (GW), 7.7 (SW), = 7.6 (WW1), 8.2 (WW2) [TiO2] = 0.001 g L-1 [FeII] and [FeIII] = 10 µM [H2O2] = 10 µM O3 flow rate: 16 mg h-1 UV lamp (254 nm) Light intensity: 1.81 × 10-6 E s-1 Pseudo-first-order rates constants in UP (min-1): pH = 3 UV: 0.109, H2O2/FeII = 0.117, H2O2/FeIII = 0.028 UV/H2O2 = 0.237, UV/TiO2 = 0.114, O3 = 0.559 FeII/UV = 0.151, FeIII/UV = 0.125 H2O2/FeII/UV = 0.265, H2O2/FeIII/UV = 0.230 pH = 7 O3 = 0.564, UV/O3 = 1.172, O3/H2O2 = 1.385, O3/H2O2/UV = 1.393, O3/TiO2 = 1.058, O3/TiO2/UV = 1.203 Pseudo-first-order rates constants in real water matrices (min-1): SW: UV: 0.041, UV/TiO2 = 0.049, H2O2/FeII = 0.025, H2O2/FeII/UV = 0.049 Gw: UV: 0.093, UV/TiO2 = 0.102, H2O2/FeII = 0.037, H2O2/FeII/UV = 0.095 WW1: O3 = 0.498, O3/H2O2 = 1.177 WW2: O3 = 0.458, O3/H2O2 = 1.059 Benitez et al. (2011) Amoxicillin (AMX) Photo-Fenton [AMX] = 50 mg L-1 Matrix: Deionized water [FeOx], [FeSO4] = 0.05 mM [H2O2] = 120 mg L-1 pH = 2.5-2.8 Light intensity (300–800 nm): Xe lamp: 250 W m-2 FeOx results: Complete AMX removal after 5 min TOC removal after 240 min: 81 % 100% N release as ammonium after 240 min Inhibition towards: V. fischeri: maintained around 30% D. magna: between 70-95 % due to oxalate. Fe SO4 results: Complete AMX removal after 15 min TOC removal after 240 min: 73 % 100% N release as ammonium after 240 min Inhibition towards: V. fischeri: maintained around 30% D. magna: 65 % (0 min), 5 % (90 min), 100 % (150 min), 45 % (240 min) Trovó et al. (2011) Chapter 1 26 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References Lincomycin (LCM) UV/TiO2 [LCM] = 10 – 75 µM (4 -30 mg L-1) Matrix: Deionized water [TiO2] = 0.2 g L-1 pH = 6.3 Solar light Complete removal of LCM in every concentration value with less than 2 Einstein (cumulative photon energy). Full TOC removal depends on the initial concentration of LCM The presence of membrane reactors allowed catalyst separation and the operation in continuous mode, as the membranes rejection for LCM and its oxidation products was quite high. Augugliaro et al. (2005) Lincomycin (LCM) UV/TiO2 [LCM] = 10, 20 and 50 mg L-1 Matrix: Deionized water [TiO2] = 0 and 0.4 g L-1 pH ranged between 5.6 – 6.5 Medium pressure Hg lamp (125 W): Photon flux: 10 mW cm-2 [TiO2] = 0 g L-1 results: LCM removal after 5 h: 70 (10 mg L-1), 40 (20 mg L-1) and 20 % (50 mg L-1) No mineralization [TiO2] = 0.4 g L-1 results: Complete LCM removal after: 0.5 (10 mg L-1), 0.75 (20 mg L-1) and 2 h (50 mg L-1). Complete TOC removal after 5 (10 mg L-1) and 8 h(20 mg L-1) Nearly 95% TOC removal after 10 h (50 mg L-1) Recovery of 70 % of sulfur as sulfate after 9 h Recovery of 75 % of N as ammonium and 25 % as nitrate Di Paola et al. (2006) Sulfamethazine (SMT) UV/TiO2, UV/ZnO [SMT] = 50 mg L-1 Matrix: Deionized water [TiO2] = [ZnO] = 1 g L-1 [H2O2] = 0 – 800 mg L-1 pH = 4.8 UVA lamp (max 366 nm): Photon flux: 2.02 × 10-4 E L-1 min-1 Initial apparent photonic efficiencies (ξ) in SMT and TOC removal: TiO2: 1.25 ± 0.08; 3.10 ± 0.08; 1.25 ± 0.08; ZnO: 2.15 ± 0.06; 3.58 ± 0.21; 1.25 ± 0.08; TiO2 + 100 mg L-1 H2O2: 2.79 ± 0.08; 3.54 ± 0.07 ZnO + 100 mg L-1 H2O2: 1.02 ± 0.10; 2.47 ± 0.09 Kaniou et al. (2005). Sulfamethoxazole (SMX) UV/TiO2 [SMS] = 100 mg L-1 Matrix: Ultrapure water [TiO2] = 0 – 2.0 g L-1 pH = 2 – 11 Light intensity (300–800 nm): Xe lamp: 250 W m-2 After 6 h with optimal oxidation conditions ([TiO2] = 0.5g L-1, pH = 5): SMX removal: 82 %, TOC removal: 23 % Recovery of 40 % of sulfur as sulfate Recovery of 24 % of N as ammonium Slow increase of BOD5/COD ratio from 0.0 to nearly 0.12, related to low decrease of the specific ultraviolet absorbance (aromaticity content). Abellán et al. (2007) Chapter 1 27 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References Sulfamethoxazole (SMX) Photo-Fenton [SMX] = 200 mg L-1 Matrix: Deionized water [Fe2+] = 10 mg L-1 [H2O2] = 300 or 400 mg L-1 pH = 2.8 3 black-light blue lamps (8 W each) Results with 300 mg L-1 of H2O2 after 67 min: Complete SMX removal, 30 % TOC removal Recovery of 26 % of N in the form of ammonium Final BOD5/COD ratio of 0.18 Results with 400 mg L-1 of H2O2 after 88 min: Complete SMX removal, 53 % TOC removal Recovery of 31 % of N in the form of ammonium Final BOD5/COD ratio 0.26 Both experiments presented neither inhibition towards activated sluge nor toxicity towards V. fischeri. González et al. (2009) Mixture of SMX with 8 other pharmaceutical compounds UV/TiO2 Photo-Fenton [SMX] = 0.1 mg L-1 Matrix: Deionized water Freshwater (FW1) Freshwater w/o HCO3 (FW2) [TiO2] = 5 mg L-1 [Fe2+] = 5 mg L-1 [H2O2] = 50 mg L-1 pH = 2.8, unadjusted Solar light Irradiation time necessary to remove SMX DW: TiO2: 145 min (t30 W) Photo-Fenton (pH = 2.8): 20 min (t30 W) Photo-Fenton (unadjusted pH): 13 min (t30 W) FW1: Photo-Fenton (unadjusted pH): 14 min (t30 W) FW2: Photo-Fenton (unadjusted pH): 50 min (t30 W) Klamerth et al. (2009) Chloramphenicol (CAP) Photo-Fenton [CAP] = 200 mg L-1 Matrix: Ultrapure water [Fe2+] = 5 - 15 mg L-1 [H2O2] = 50 – 500 mg L-1 pH = 2.8 Photon flux of UVA lamp (400 W): 295-290 nm: 6.0×10-7 E s-1 295-710 nm: 3.3×10-6 E s-1 Average UVA lamp irradiance: 1100 W m-2 Average solar UVA irradiance: 320-400 nm: 41.5 ± 1.2 W m-2 Chosen conditions: [Fe2+] = 10 mg L-1, and [H2O2] = 400 mg L-1 (single addition). Results in lab-scale experiment after 60 min (QUV = 31.7 kJ L-1): Complete CAP removal after 16 min TOC and COD removal: 90 and 93 %, respectively. AOS: from +0.48 to +3.39 H2O2 consumption rate: 6.7 mg L-1 min -1 Results in solar pilot-plant after 60 min (QUV = 5.3 kJ L-1): Dark-Fenton removed 98% of CAP during mixing time. TOC and COD removal: 90 and 93 %, respectively. AOS: from +1.4 to +2.9 H2O2 consumption rate: 12.5 mg L-1 min -1 Trovó et al. (2013) Chapter 1 28 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References Ofloxacin (OFX) UV/TiO2 Photo-Fenton [OFX] = 10 mg L-1 Matrix: Effluent from WWTP [TiO2] = 0.25 - 4 g L-1 pH = 2, 8 and 10 [Fe2+] = 1 - 5 mg L-1 [H2O2] = 1.357 – 8.142 mM (46 – 277 mg L-1) pH = 2 – 4.5 Xenon lamp (1000 W): Irradiance: 273 W m-2 Chosen Photo-Fenton conditions: [Fe2+] = 5 mg L-1, [H2O2] = 2.714 mM (92 mg L-1) and pH = 3 Removal of OFX after 30 min: 100 % TOC removal after 120 min: 50 % Profile of toxicity towards D. magna (48h) at 0, 15, 30 60 and 120 min: 20, 40, 90, 90 and 50 % (respectively). Chosen TiO2 conditions: [TiO2] = 3 g L-1, pH = 8 Removal of OFX after 120 min: 60 % TOC removal after 120 min: 10 % Profile of toxicity towards D. magna (48h) at 0, 15, 30 60 and 120 min: 10, 20, 20, 10 and 5 % (respectively). Michael et al. (2010) Ofloxacin (OFX) Photo-Fenton [OFX] = 10 mg L-1 Matrix: Demineralized water Simulated natural freshwater Simulated municipal wasteawater Real effluent from WWTP [Fe2+] = 2 mg L-1 pH ~ 2.8 [H2O2] = 2.5 mg L-1 (dose addition after total consumption) Solar energy Achieved TOC removal after the consumption of 12 mg L-1 of H2O2: Demineralized water: 78.1 % Simulated natural freshwater: 58,3 % Simulated municipal wastewater (SWW): 40.5 % Real effluent from WWTP (RE): 35.8 % Evaluation of toxicity towards V. fischeri (30 min): SWW: 33 % (0 min), 7 % (after 30 min of illumination) RE: 33 % (0 mg L-1 of H2O2), 64 % (12 mg L-1 of H2O2) Michael et al. (2013) Chapter 1 29 Table 1.4. (Continued) Antibiotics Process Operating conditions Results References 8 antibiotics, amongst 32 micropollutants, detected in real a secondary effluent UV254/H2O2, Photo-Fenton (UV254 or UVVIS) Σmicropollutants (ng L-1) = 29.506, from which: [Azithromycin (ATM)] = 295, [Ciprofloxacin (CFX)] = 129 [Clarithromycin (CTM] = 518 [Metronidazole (MNZ)] = 456 [Norfloxacin (NFX)] = 27 [Ofloxacin (OFX)] = 41 [Sulfamethoxazole (SM)] = 578 [Trimethoprim (TMP)] = 131 Matrix: Effluent from WWTP [Fe2+] = 5 mg L-1 (1.48 mg L-1 already present in the effluent) [H2O2] = 10 – 25, 50 mg L-1 Natural pH of the effluent (7.4) UVA lamp (254 nm): Photon flux: 1.5×10-6 E s-1 UV-VIS lamp (290 – 800 nm) Irradiance: 550 W m-2 Removal % ([Fe2+] = 5 mg L-1, [H2O2] = 50 mg L-1): De la Cruz et al. (2012) UV254/H2O2 (30 min) Fenton (30 min) P-F (UV254) (30 min) P-F (UV-VIS) (90 min) ATM 100 23 23 32 CFX 100 60 60 100 CTM 100 10 10 15 MNZ 100 55 55 88 NFX 100 100 100 100 OFX 100 49 49 100 SMX 100 0 0 0 TMP 100 20 20 43 4 antibiotics, amongst 62 micropollutants, detected in real secondary effluents Photo-Fenton Average detected concentrations (ng L-1, n = 10): [Ciprofloxacin (CFX)] = 705 [Ofloxacin (OFX)] = 1082 [Sulfamethoxazole (SMX)] = 844 [Trimethoprim (TMP)] = 332 [Sulfapyridine (SPD)] = 241 Matrix: Effluents collected from WWTP [Fe2+] = 5 mg L-1 [H2O2] kept at 50 mg L-1 pH = 3: classic photo-Fenton pH = 6.5: modified photo-Fenton with 10 mg L-1 Humic acids (HA) or 0.2 mM of EDDS Solar energy Results of photo-Fenton experiments performed with: Klamerth et al. 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Comparison of the efficiency of ·OH radical formation during ozonation and the advanced oxidation processes O3/H2O2 and UV/H2O2. Water Res. 40, 3695-3704. Sarmah, A.K., Meyer, M.T., Boxall, A.B.A., 2006. A global perspective on the use, sales, exposure pathways, occurrence, fate and effects of veterinary antibiotics (VAs) in the environment. Chemosphere 65, 725-759. Schaar, H., Clara, M., Gans, O., Kreuzinger, N., 2010. Micropollutant removal during biological wastewater treatment and a subsequent ozonation step. Environ. Pollut. 158, 1399-1404. Seifrtová, M., Nováková, L., Lino, C., Pena, A., Solich, P., 2009. An overview of analytical methodologies for the determination of antibiotics in environmental waters. Anal. Chim. Acta 649, 158-179. Senta, I., Matošić, M., Jakopović, H.K., Terzic, S., Ćurko, J., Mijatović, I., Ahel, M., 2011. Removal of antimicrobials using advanced wastewater treatment. J. Hazard. Mater. 192, 319328. Silva, M.R.A., Trovó, A.G., Nogueira, R.F.P., 2007. Degradation of the herbicide tebuthiuron using solar photo-Fenton process and ferric citrate complex at circumneutral pH. J. Photochem. Photobiol. A: Chem. 191, 187-192. Sipma, J., Osuna, B., Collado, N., Monclús, H., Ferrero, G., Comas, J., Rodriguez-Roda, I., 2010. Comparison of removal of pharmaceuticals in MBR and activated sludge systems. Desalination 250, 653-659. Snyder, S.A., Adham, S., Redding, A.M., Cannon, F.S., DeCarolis, J., Oppenheimer, J., Wert, E.C., Yoon, Y., 2007. Role of membranes and activated carbon in the removal of endocrine disruptors and pharmaceuticals. Desalination 202, 156-181. Son, H.S., Ko, G., Zoh, K.D., 2009. Kinetics and mechanism of photolysis and TiO2 photocatalysis of triclosan. J. Hazard. Mater. 166, 954-960. Tambosi, J.L., de Sena, R.F., Favier, M., Gebhardt, W., José, H.J., Schröder, H.F., de Fatima Peralta Muniz Moreira, R., 2010. Removal of pharmaceutical compounds in membrane bioreactors (MBR) applying submerged membranes. Desalination 261, 148-156. Thiele-Bruhn, S., 2003. Pharmaceutical antibiotic compounds in soils - A review. J. Plant Nutr. Soil Sci. 166, 145-167. Trovó, A.G., de Paiva, V.A.B., Machado, A.E.H., de Oliveira, C.A., Santos, R.O., 2013. Degradation of the antibiotic chloramphenicol by photo-Fenton process at lab-scale and solar pilot plant: Kinetic, toxicity and inactivation assessment. Sol. Energy 97, 596-604. Trovó, A.G., Pupo Nogueira, R.F., Agüera, A., Fernandez-Alba, A.R., Malato, S., 2011. Degradation of the antibiotic amoxicillin by photo-Fenton process - Chemical and toxicological assessment. Water Res. 45, 1394-1402. Chapter 1 39 Vaz-Moreira, I., Nunes, O.C., Manaia, C.M., 2011. Diversity and antibiotic resistance patterns of Sphingomonadaceae isolates from drinking water. Appl Environ Microbiol 77, 5697-5706. Watkinson, A.J., Murby, E.J., Costanzo, S.D., 2007. Removal of antibiotics in conventional and advanced wastewater treatment: Implications for environmental discharge and wastewater recycling. Water Res. 41, 4164-4176. Webb, S., Ternes, T., Gibert, M., Olejniczak, K., 2003. Indirect human exposure to pharmaceuticals via drinking water. Toxicol. Lett. 142, 157-167. Westerhoff, P., Moon, H., Minakata, D., Crittenden, J., 2009. Oxidation of organics in retentates from reverse osmosis wastewater reuse facilities. Water Res. 43, 3992-3998. Wilson, B.A., Smith, V.H., Denoyelles Jr, F., Larive, C.K., 2003. Effects of three pharmaceutical and personal care products on natural freshwater algal assemblages. Environ. Sci. Technol. 37, 1713-1719. Wols, B.A., Hofman-Caris, C.H.M., 2012. Review of photochemical reaction constants of organic micropollutants required for UV advanced oxidation processes in water. Water Res. 46, 2815-2827. Yoon, Y., Westerhoff, P., Snyder, S.A., Wert, E.C., Yoon, J., 2007. Removal of endocrine disrupting compounds and pharmaceuticals by nanofiltration and ultrafiltration membranes. Desalination 202, 16-23. Zhao, C., Pelaez, M., Duan, X., Deng, H., O'Shea, K., Fatta-Kassinos, D., Dionysiou, D.D., 2013. Role of pH on photolytic and photocatalytic degradation of antibiotic oxytetracycline in aqueous solution under visible/solar light: Kinetics and mechanism studies. Appl. Catal., B 134–135, 83-92. Chapter 1 40 41 2 Materials and methods This chapter contains an overview of all the chemicals and reagents used in this thesis, a detailed description of the experimental units used to perform all the photocatalytic trials and the corresponding experimental procedures. The employed analytical methods are also herein described. Chapter 2 42 Chapter 2 43 2.1 Chemicals and Reagents Oxytetracycline hydrochloride (OTC.HCl, CAS# 2058-46-0), Oxolinic acid (OXA, CAS# 14698-29-4) and Amoxicillin (AMX, CAS# 26787-78-0) were purchased from Sigma-Aldrich. Their molecular structure can be seen in Figure 2.1, while Table 2.1 presents a short summary of documented physico-chemical properties of the three compounds. a) b) c) Figure 2.1. Molecular structures of a) OTC, b) OXA, c) AMX antibiotics. Table 2.1. Physico-chemical properties of OTC, OXA and AMX antibiotics. Property OTC OXA AMX Molecular formulaa C22H24N2O9 (.HCl) C13H11NO5 C16H19N3O5S Molecular Weight (g mol-1)a 460.43 (496.89) 261.2 365.40 HPLC puritya 97 % 99 % 100 % Specific gravityb 1.63 1.55 n.a. Solubility (pH 7) - in water (20 °C)b 1000 mg L-1 3.2 mg L-1 3430 mg L-1 Log KOW at pH 7 (20 °C)b -1.22 1.67 0.87 pKa1 3.22c 6.92d 2.68e pKa2 7.46c - 7.49e pKa3 8.94c - 8.94e n.a. – not available; aCertificate of analysis (Sigma-Aldrich); bVSDB (2013); cQiang and Adams (2004): Ionic strength = 0 M; T = 23 °C ; dJiménez-Lozano et al. (2002): Ionic strength = 0 M; T = 25 °C; eAndreozzi et al. (2005): Ionic strength = 0.1 M; T = 25 °C Antibiotics were stored at 4 °C and solutions were prepared daily by weighting the appropriate mass, taking into consideration HPLC purity and molecular weight (as in the case of Oxytetracycline (OTC) against OTC.HCl). Solutions of OTC and AMX were easily prepared, whereas solutions of OXA, given its very limited solubility in neutral pH levels, required an initial pH adjustment. According to the Sigma-Aldrich product information sheet (Product number O0877), OXA solutions were prepared in 0.05 M NaOH (pH ~ 11), which required a subsequent neutralization step. Heterogeneous photocatalytic experiments used Degussa P-25 (80 % anatase and 20 % rutile) Titanium Dioxide (TiO2). Photo-Fenton experiments were performed using hydrogen peroxide (Quimitécnica, S.A., 50 % (w/v), 1.10 g cm-3), iron (II) sulphate heptahydrate (Panreac), Chapter 2 44 iron (III) hexahydrate (Merck), oxalic acid dihydrate (VWR Prolabo, purity  98 %) and citric acid monohydrate (VWR; 100 %). Ultrapure and deionized water necessary for analysis or antibiotic solutions were obtained using a millipore system (Direct-Q model) and reverse osmosis system (Panice®), respectively. Pure or diluted solutions of sulfuric acid (Pronalab, 96 %, 1.84 g/cm3) and sodium hydroxide (Merck) were used for pH adjustment. NaCl, MgSO4.7H2O, NaHCO3, KNO3, NH4Cl, K3PO4.3H2O, D-Mannitol and NaN3 were all analytical grade, while Humic acids were Alfa Aesar (CAS# 1415-93-6). For HPLC-DAD analysis, gradient-grade acetonitrile and methanol were obtained from Merck, while oxalic acid dehydrate (100 %) was from VWR Prolabo. Chapter 2 45 2.2 Experimental units and procedure 2.2.1 SOLARBOX lab-scale photoreactor 2.2.1.1 Description The first set of experiments described in Chapter 3 was performed in the Department of Chemical Engineering at the Faculty of Chemistry of the University of Barcelona. The SOLARBOX lab-scale photocatalytic apparatus comprises: i) a solar radiation simulator (Solarbox, Co.fo.me.gra 220 V 50 Hz) ii) a tubular reactor (illuminated volume, Vi = 0.078 L, Schott-Duran type 3.3, Germany, cut-off at 280 nm, internal diameter 21.1 mm, length 223 mm and thickness 2.11 mm) in the axis of a parabolic mirror; iii) one glass vessel (capacity of 1.0 L) with a cooling jacket, coupled to a refrigerated thermostatic bath (Haake K10) to ensure a constant temperature during the experiment; iv) a magnetic stirrer (OVAN) to ensure complete homogenization of the solution inside the glass vessel; v) one peristaltic pump (Ismatec, model Ecoline VC-280 II) to promote the water recirculation between the photoreactor and the glass vessel; vi) pH meter (GLP 22). All connections and pipes employed were made of Teflon and/or glass material to avoid losses by adsorption. A scheme and general views of the installation can be seen in Figure 2.2. 2.2.1.2 Experimental procedure The stirred reservoir tank was filled with a 20 mg L-1 Oxytetracycline solution. The solution was continuously pumped through the Duran tubular reactor (illuminated volume, Vi = 0.078 L) placed at the bottom of the Solarbox, in the axis of a parabolic mirror, and recirculated to the reservoir tank at a flow rate of 0.65 L min-1 (illuminated time, ti = 0.12 min; dark time, td = 1.42 min). Once the air of the system was purged and after some minutes of recirculation, a sample was taken. Then, TiO2 was added to the reservoir tank and the resulting suspension was left recirculating in the dark for 60 minutes to achieve adsorption equilibrium. After that, the Solarbox was turned on and a Xe-OP lamp (Phillips 1 kW) placed inside started to irradiate the tubular reactor. Radiant power entering the reactor was determined to be 3.55 J s-1 (between 290 and 400 nm). It was calculated by uranyl oxalate actinometry (Kuhn et al., 2004) taking into account the transmittance of Duran glass, the reactor’s geometry, the actinometric system quantum yield and the useful wavelengths for TiO2. The temperature of the solution was kept at 25 ºC by controlling the temperature of the jacket in the reservoir tank through the ultrathermostat bath. All samples were pre-filtrated through 0.2 μm Nylon VWR membrane filters Chapter 2 46 before analysis. Initial pH values were unbuffered and adjusted when needed with diluted phosphoric acid or sodium hydroxide solutions. Figure 2.2. SOLARBOX lab-scale experimental set-up: a) schematic representation (adapted from Méndez-Arriaga et al. (2008)); b) and c) views of the thermostatic bath, reservoir tank, peristaltic pump, sunlight simulator and photoreactor equipped with a parabolic reflector a) b) c) Chapter 2 47 2.2.2 SUNTEST lab-scale photoreactor 2.2.2.1 Description The lab-scale experiments described in Chapters 4, 5, 6 and 7 were performed at the Chemical Engineering Department of the Faculty of Engineering of the University of Porto (FEUP). The lab-scale photocatalytic apparatus (SUNTEST) comprises: i) a solar radiation simulator (ATLAS, model SUNTEST XLS+) with 1100 cm2 of exposition area, a 1700 W air-cooled xenon arc lamp, a daylight filter and quartz filter with infrared (IR) coating; ii) a compound parabolic collector with 0.025 m2 of illuminated area with electropolished anodized aluminium reflectors and borosilicate tube (Schott-Duran type 3.3, Germany, cut-off at 280 nm, internal diameter 46.4 mm, length 160 mm and thickness 1.8 mm); iii) one glass vessel (capacity of 1.5 L) with a cooling jacket, coupled to a refrigerated thermostatic bath (Lab. Companion, model RW-0525G) to ensure a constant temperature during the experiment; iv) a magnetic stirrer (Velp Scientifica, model ARE) to ensure complete homogenization of the solution inside the glass vessel; v) one peristaltic pump (Ismatec, model Ecoline VC-380 II, at a flow rate of 0.63 L min-1) to promote the water recirculation between the photoreactor and the glass vessel; vi) pH and temperature meter (VWR sympHony SB90M5). Vi = 0.270 L; Vi/Vt = 0.23 (initial Vt of 1.2 L); ti = 0.43 min; tdark = 1.16 min. All the systems are connected using Teflon tubing. The photoreactor has two polypropylene caps with four equidistant inlets and outlets to ensure a better distribution of the feed stream throughout the reactor. A schematic representation of the SUNTEST installation can be seen in Figure 2.3. 2.2.2.2 Experimental procedure In all the experiments, a solution of 1.2 L of 20 mg L-1 of OTC (4.34 × 10-5 M), OXA (7.66 × 10-5 M) or AMX (5.47 × 10-5 M), was added to the recirculation glass vessel and homogenized by stirring in darkness. The temperature set-point of the refrigerated thermostatic bath was controlled to keep the antibiotic solution at 25 °C. After 15 minutes, a sample was taken to confirm the initial antibiotic and DOC concentrations. The experimental procedure then varied according to the employed photocatalytic processes: a) TiO2/UV; b) Fe2+/H2O2/UV-Vis and c) Fe3+/Oxalate or Citrate/H2O2/UV-Vis. Chapter 2 54 Table 2.2. Pump program for HPLC gradient runs. Time (min) Mobile phase – A: B: C (%) OTC OXA AMX 0.0 10 10 10 20 10 70 10 5 80 3.0 15 10 15 25 10 65 15 5 75 5.0 20 10 20 30 10 60 20 5 70 7.0 10 10 10 20 10 70 10 5 80 14.0 10 10 10 20 10 70 10 5 80 The corresponding calibration curves of OTC, OXA and AMX were constructed with n = 5 concentration levels, ranging from 1.0 to 20 mg L-1, using chromatographic peak areas as a function of their concentration, and limits of detection and quantification were also calculated. Table 2.3 presents a summary of the analytical parameters of the calibration curves for each antibiotic. The Limit of Quantification (LOQ) and Limit of Detection (LOD) values were obtained with the following calculations (adapted from Miller and Miller (1984)): LOQ: bLOQL bxa  OQ y , where b sb  10yLOQ ; LOD: bLOD bxa  LOD y , where b sb  3yLOD , where a is the slope and sb is the residual standard deviation of the intercept (b). Table 2.3. Analytical parameters of working calibration curves of OTC, OXA and AMX antibiotics. Antibiotic λ (nm) Regression equation parameters Slope (a ± sa) Intercept (b ± sb) R2a sa/ab LOQc LODd OTC 354 162923 ± 4×102 -22762 ± 5×103 0.999 0.265 0.33 0.10 OXA 260 842813 ± 2×104 22548 ± 2×104 0.999 0.281 0.29 0.09 AMX 230 112963 ± 4×102 1681 ± 4×103 0.999 0.311 0.38 0.11 aCorrelation coefficient, bRelative standard deviation of a, c, dLimits of Quantification/Detection (mg L-1) 2.3.2 Dissolved Organic Carbon (DOC) To assess the degree of mineralization of antibiotics in the photolytic/photocatalytic experiments, the dissolved organic carbon (DOC) was determined by means of a TC-TOC-TN analyzer, equipped with an ASI-V autosampler (Shimadzu, model TOC-VCSN) and provided with a NDIR detector. Calibrations were performed with standard solutions of potassium hydrogen phthalate (total carbon) and a mixture of sodium hydrogen carbonate/carbonate (inorganic carbon). Total nitrogen was measured in the same TC-TOC-TN analyzer coupled with a TNM-1 unit (Shimadzu, model TOC-VCSN) calibrated with standard solutions of potassium nitrate, through thermal decomposition and NO detection by chemiluminescence method. Chapter 2 55 2.3.3 Inorganic ions and low-molecular-weight carboxylate anions Ion chromatography was used to monitor the release or formation of inorganic compounds and low-molecular-weight carboxylate anions (LMWCA) along the photocatalytic experiments, as well as to characterize the tested effluents used in chapter in Chapters 6 and 7. Anions (chloride, nitrate, nitrite, phosphate, sulphate) and LMWCA (acetate, propionate, formate, pyruvate, valerate, malonate, maleate, oxalate, phthalate and citrate) were monitored with a Dionex ICS-2100 system, equipped with an AS 11-HC 4 × 250 mm column and a ASRS®300 4 mm suppressor. Cations (ammonium) were monitored with a Dionex DX-120 system, equipped with a CS12A 4 × 250 mm column and a CSRS®300 4 mm suppressor. Isocratic elution was performed using 30 mM NaOH/20 mM methane sulfonic acid, at a flow rate of 1.5 and 1.0 mL min-1, for anion and cation analysis, respectively. The gradient program used for the quantification of LMWCA comprised a pre-run for 8 min with 1 mM NaOH, 20 min with 30 mM NaOH and 10 min with 60 mM NaOH, at a flow rate of 1.5 mL min-1, using an eluent generator cartridge (Dionex, RFICTM). The chromatography software was Chromeleon client, version 6.80 DU10a (Dionex corporation, 1994-2010). In chapters 6 and 7, the same HPLC-DAD apparatus, using a RezexTM ROA-Organic Acid H+ (8 %), LC Column 300 × 7.8 mm), was also used to monitor oxalate, oxamate, tartronate, acetate, malonate, maleate and formate. The isocratic method used 0.005 N H2SO4 delivered at a flow rate of 0.5 mL min-1. Run time was 50 min, injection volume 10 µL and the wavelength of the detector was set at 210 nm. 2.3.4 Hydrogen peroxide and dissolved iron concentration Evaluation of H2O2 concentration during experiments was performed by the metavanadate method, based on the reaction of H2O2 with ammonium metavanadate in acidic medium, which results in the formation of a red-orange color peroxovanadium cation, with maximum absorbance at 450 nm (Nogueira et al., 2005). Dissolved iron concentration was determined by colorimetry with 1,10-phenantroline according to ISO 6332. 2.3.5 UV spectra and photometric measurements The spectrophotometric measurements to obtain the antibiotics’ UV absorption spectra were carried out with a UNICAM HEλIOS α spectrophotometer. The absorbance measurements to determine the concentration of dissolved iron and of H2O2 were carried out with a Spectroquant® Pharo 100 (Merck) spectrophotometer. Chapter 2 56 2.3.6 Ecotoxicity and Antimicrobial activity In Chapter 3, ecotoxicity of the samples was measured as acute toxicity using the Microtox® test, where the inhibition percentage of the bioluminiscence of Vibrio fischeri after 15 min of contact time was determined. For intermediates identification, samples were analyzed using an electrospray/mass spectrometer (ion spray) ESI-MS, and a LC-MSD-TOF (2006) mass spectrometer. The antimicrobial activity (AA) was assessed by the biomass yield of strains Escherichia coli DSM 1103 (Chapters 4, 5) and Staphylococcus aureus DSM 1104 (Chapter 5) grown for 24 h in filter sterilized phototreated samples. The samples were supplemented with 2 g L-1 yeast extract, and AA was measured by changes in optical density (λ = 610 nm) and normalized by that of a positive control grown in mineral medium B (Barreiros et al., 2003) supplemented with the same concentration of yeast extract. All the assays were performed in triplicate.In Chapter 6, the AA assays were carried out in 96-well microtiter plates using a Synergy HT Multi-Mode Microplate Reader (Biotek Instruments, USA) and E. coli DSM 1103 as test strain. Samples collected over time were supplemented with 2 g L-1 of yeast extract and inoculated with the test strain at an initial optical density at 610 nm of 0.08. Cultures were grown at 30 °C and 300 rpm for 20 h. Controls with different OTC concentrations (0 to 20 mg L-1) were grown in parallel under the same conditions. Normalized biomass yield was obtained by dividing the biomass yield of the culture grown in the sample by the biomass yield obtained in the absence of OTC. 2.3.7 BOD5 and COD analysis In Chapter 3, biochemical oxygen demand at 5 days (BOD5) determinations were carried out according to the Standard Methods (5120D) by means of the OxiTop® system, while for chemical oxygen demand (COD), the Standard Methods (5220D) procedures were followed (APHA, 1985). 2.4 Kinetic modelling A pseudo-first-order mathematical model was fitted to the experimental data obtained from the kinetic studies by a non-linear regression method (software Fig.P for Windows from Fig.P Software Incorporated). The model parameters were obtained by minimizing the sum of the squared deviations between experimental and predicted values. Model goodness was evaluated through the calculation of relative standard deviations (σi), regression coefficients (R2) and residual variance (SR2). Chapter 2 57 2.5 Matrix characterization Table 2.4 presents the main characteristics of the four matrices in which AMX biodegradation trials were performed in chapter 8, as well as of the two effluents into which OTC was spiked in Chapter 6. Table 2.4. Main characteristics of the used matrices (Chapter 7) and tested effluents (Chapter 6, 7). Matrix pH DOC / IC (mg L-1) Average anion concentrations (g L-1) [PO43-] [Cl-] [SO42-] EM 7.2 421 / 4.24 2.56 0.34 0.38 Buffer 7.2 Residual 4.75 NaCl 6.5 Residual 5.15 WW 6.8 5.6 / 4.9 0.01 0.06 0.04 Effluent pH DOC / IC (mg L-1) Average anion concentrations (mg L-1) [PO43-] [Cl-] [SO42-] WW 6.5 5.5 / 5.1 13.5 61.5 42.9 TF 6.7 3.9 / 1.8 0.95 Chapter 2 58 2.6 References Andreozzi, R., Canterino, M., Marotta, R., Paxeus, N., 2005. Antibiotic removal from wastewaters: The ozonation of amoxicillin. J. Hazard. Mater. 122, 243-250. APHA, 1985. Standard Methods for the Examination of Water and Wastewater (sixteenth ed.). American Public Health Association/American Water Works Association/Water Pollution Control Federation, Washington DC, USA (1985). Barreiros, L., Nogales, B., Manaia, C.M., Silva Ferreira, A.C., Pieper, D.H., Reis, M.A., Nunes, O.C., 2003. A novel pathway for mineralization of the thiocarbamate herbicide molinate by a defined bacterial mixed culture. Environ. Microbiol. 5, 944-953. Delépée, R., Maume, D., Le Bizec, B., Pouliquen, H., 2000. Preliminary assays to elucidate the structure of oxytetracycline’s degradation products in sediments: Determination of natural tetracyclines by high-performance liquid chromatography–fast atom bombardment mass spectrometry. J. Chrom. B Biomed. Sci. Appl. 748, 369-381. Hirakawa, K., Mori, M., Yoshida, M., Oikawa, S., Kawanishi, S., 2004. Photo-irradiated titanium dioxide catalyzes site specific DNA damage via generation of hydrogen peroxide. Free Radical Res. 38, 439-447. Jiménez-Lozano, E., Marqués, I., Barrón, D., Beltrán, J.L., Barbosa, J., 2002. Determination of pKa values of quinolones from mobility and spectroscopic data obtained by capillary electrophoresis and a diode array detector. Anal. Chim. Acta 464, 37-45. Klamerth, N., Rizzo, L., Malato, S., Maldonado, M.I., Agüera, A., Fernández-Alba, A.R., 2010. Degradation of fifteen emerging contaminants at μg L-1 initial concentrations by mild solar photo-Fenton in MWTP effluents. Water Res. 44, 545-554. Kuhn, H.J., Braslavsky, S.E., Schmidt, R., 2004. Chemical actinometry (IUPAC technical report). Pure Appl. Chem. 76, 2105-2146. Lindsey, M.E., Tarr, M.A., 2000. Inhibition of hydroxyl radical reaction with aromatics by dissolved natural organic matter. Environ. Sci. Technol. 34, 444-449. Méndez-Arriaga, F., Esplugas, S., Giménez, J., 2008. Photocatalytic degradation of nonsteroidal anti-inflammatory drugs with TiO2 and simulated solar irradiation. Water Res. 42, 585594. Miller, J.C., Miller, J.N., 1984. Statistics for Analytical Chemistry. Wiley, New York. Miralles-Cuevas, S., Arqués, A., Maldonado, M.I., Sánchez-Pérez, J.A., Malato Rodríguez, S., 2013. Combined nanofiltration and photo-Fenton treatment of water containing micropollutants. Chem. Eng. J. 224, 89-95. Nogueira, R.F.P., Oliveira, M.C., Paterlini, W.C., 2005. Simple and fast spectrophotometric determination of H2O2 in photo-Fenton reactions using metavanadate. Talanta 66, 86-91. Oka, H., Ito, Y., Matsumoto, H., 2000. Chromatographic analysis of tetracycline antibiotics in foods. J. Chromatogr. A 882, 109-133. Pouliquen, H., Delépée, R., Larhantec-Verdier, M., Morvan, M.L., Le Bris, H., 2007. Comparative hydrolysis and photolysis of four antibacterial agents (oxytetracycline oxolinic Chapter 2 59 acid, flumequine and florfenicol) in deionised water, freshwater and seawater under abiotic conditions. Aquaculture 262, 23-28. Qiang, Z., Adams, C., 2004. Potentiometric determination of acid dissociation constants (pKa) for human and veterinary antibiotics. Water Res. 38, 2874-2890. Raja, P., Bozzi, A., Mansilla, H., Kiwi, J., 2005. Evidence for superoxide-radical anion, singlet oxygen and OH-radical intervention during the degradation of the lignin model compound (3-methoxy-4-hydroxyphenylmethylcarbinol). J. Photochem. Photobiol. A: Chem. 169, 271278. VSDB, The Veterinary Substance Database, in, developed by the Agriculture & Environment Research Unit (AERU), University of Hertfordshire 2011-2013, 2013. Chapter 2 60 61 3 Photocatalytic degradation of oxytetracycline using TiO2 under natural and simulated solar radiation The main objective of the present study was to assess the photocatalytic degradation mediated by suspended TiO2 of the antibiotic Oxytetracycline (OTC), under simulated solar irradiation. The influence of the photocatalyst concentration and initial pH was tested, and optimal parameters to remove OTC and enhance its mineralization were found to be 0.5 g L-1 of TiO2 with no initial pH adjustment (pH0 ~ 4.4). Under optimal conditions, biodegradability was enhanced, while its toxicity was decreased. A scheme of possible degradation pathways of OTC is presented. The same experiment performed in pilot-plant equipped with Compound Parabolic Collectors (CPCs), using natural solar radiation, demonstrates that CPC reactors were more effective in the usage of accumulated UV energy to completely degrade and almost completely mineralize OTC. This Chapter is based on the research article “Pereira, J.H.O.S., Vilar, V.J.P., Borges, M.T., González, O., Esplugas, S., Boaventura, R.A.R. Photocatalytic degradation of oxytetracycline using TiO2 under natural and simulated solar radiation. Sol. Energy (2011) 85, 2732-2740”. Chapter 3 62 Chapter 3 63 3.1 Introduction The contamination of the environment by human and veterinary pharmaceuticals as a result of metabolic excretion, improper disposal and/or industrial waste has been subject to special attention over recent years (Halling-Sørensen et al., 1998). The main concern lies on their continual release and subsequent widespread presence in the environment because even at trace concentrations (ng L-1 to µg L-1), these drugs present a whole spectrum of possible and still mostly unknown physiological effects on non-target species (Daughton and Ternes, 1999). Among these substances, antibiotics are worrisome due to the risk of aquatic toxicity and the development of resistant bacterial strains (Kemper, 2008; Kümmerer, 2009). Conventional wastewater treatment plants (WWTPs) are not designed to remove antibiotics present at trace levels (Okuda et al., 2008; Miège et al., 2009). Different approaches to this problem may rely on source control measures or on the improvement of WWTPs with new endof-pipe technologies (Larsen et al., 2004), including advanced treatments such as membrane filtration, activated carbon adsorption and Advanced Oxidation Processes (AOPs). AOPs are recommended when wastewater constituents, such as pesticides or pharmaceuticals, have a high chemical stability and/or low degradability. Considering the ability of AOPs to achieve the complete mineralization of pollutants to CO2, water and inorganic compounds, or at least their partial oxidation to more biodegradable and/or less harmful intermediates, they allow a more useful and cost efficient combination with biological processes (Marco et al., 1997; Ikehata et al., 2006). AOPs comprise different processes that involve the generation of hydroxyl radicals (•OH), which are very reactive and non-selective, and can be divided in photochemical (UV/O3, UV/H2O2), photocatalytic (TiO2/UV, Photo-Fenton) or chemical oxidation processes (O3, O3/H2O2, H2O2/Fe2+) (Gogate and Pandit, 2004a; b; Poyatos et al., 2009). These processes are rather costly and energy intensive (Rosenfeldt et al., 2006), therefore, research is being focused on the application of AOPs assisted by solar irradiation, such as heterogeneous photocatalysis and Photo-Fenton reaction (Bahnemann, 2004; Muñoz et al., 2005). Oxytetracycline (OTC, Figure 3.1) is a widely used broad spectrum antibiotic, especially employed in veterinary medicine (Halling-Sørensen et al., 1998; Sarmah et al., 2006). Studies on OTC photolytic and ozonation degradation (Pouliquen et al., 2007; Jiao et al., 2008; Li et al., 2008; Zhao et al., 2009; Zhao et al., 2010) and on TiO2-supported zeolites using UV sterilization lamps (Zhao and Deng, 2009; Zhao et al., 2009; Zhao et al., 2010) have been reported in the literature, but none, to the best of our knowledge, used both suspended TiO2 and UV radiation in the solar range. Chapter 3 70 3.3.2 Solar CPC pilot plant experiments Using the same initial conditions ([OTC] = 20 mg L-1, [TiO2] = 0.5 g L-1 and free pH), a photolytic and photocatalytic set of experiments were carried out in the CPC pilot plant, during summer days. Both antibiotic degradation and mineralization are plotted in terms of accumulated UV energy (QUV,n kJ L-1) in Figure 3.8. Solar photocatalysis needed almost 10 times less energy (1.6 kJ L-1) than solar photolysis (12 kJ L-1) to completely remove OTC from solution. After 9 kJ L-1 of accumulated energy, 80% of mineralization was reached for solar photocatalysis and, as expected, almost negligible mineralization with solar photolysis. It must be taken into consideration that the CPC pilot plant has no means to maintain a constant temperature, and, in spite of the fact that it generally fails to play a significant role in the photocatalytic process (Malato et al., 2009), temperature of the aqueous solution during solar exposure may have influenced OTC removal/mineralization (Doi and Stoskopf, 2000). Although similar amounts of accumulated UV energy are expected for the same degrees of OTC degradation and mineralization when using natural and simulated solar radiation, our experiments show that this was not the case, as it can be seen in Figure 3.8, which also presents the Solarbox results under the same conditions, but expressed in terms of QUV (kJ L-1). Figure 3.8. Removal profiles of OTC and TOC under simulated (,) and real (,) solar photolysis and under simulated (,) and real (▲,) solar photocatalysis with 0.5 g L-1 TiO2 and free initial pH. Hence, the kinetics of both photolytic and photocatalytic degradation of OTC for each experimental set-up were studied in order to compare their efficiency and are presented in Table 3.2. The fitting of the results showed that all reactions approximately followed pseudo-first order kinetics. The photolytic rate constants were calculated as 0.033 ± 0.001 L kJ-1 and 0.205 ± 0.005 L kJ-1, for Solarbox and CPC experiments respectively, whereas photocalytic rate constants were 0.28 ± 0.03 L kJ-1 and 2.63 ± 0.03 L kJ-1, for Solarbox and CPC experiments 0 5 10 15 20 25 30 35 40 0.0 0.2 0.4 0.6 0.8 1.0 RAD-ON Oxytetracycline (C/C0) QUV (kJ/L) 0.0 0.2 0.4 0.6 0.8 1.0 <LOD TOC/TOC0 Chapter 3 71 respectively. The somewhat proportionally increase of rate constants from photolytic to photocatalytic experiments are relatively similar for both experimental set-ups, but the major difference lies on the increase of rate constants for each reaction from the Solarbox to the CPC experiments, corroborated with the initial rate of reactions, presented in Table 3.2 Table 3.2. Kinetic constant values for Solarbox and CPC photolysis and photocatalysis ([TiO2] = 0.5 g L-1, free pH) experiments. Experiment k (L kJ-1)a r0 (mg kJ-1)b R2 SR2 (mg2 L-2) Solarbox photolysis 0.033 ± 0.001 0.73 ± 0.02 0.995 0.069 CPC photolysis 0.205 ± 0.005 2.6 ± 0.2 0.999 0.083 Solarbox TiO2 0.28 ± 0.03 4.3 ± 0.5 0.988 1.165 CPC TiO2 2.63 ± 0.03 36.2 ± 0.4 0.999 0.009 a – pseudo-first order kinetic rate, b - initial reaction rate The relationship between the amount of radiation that enters a reactor and the radiation absorbed by the catalyst is the most important difference between different scale/set-up experiments (Wiebe and Moore, 1977), and it depends on radiation source and its spectral distribution, catalyst and contaminant concentration, and, finally, reactor geometry. As most of these parameters are the same, or fairly equal between the two experimental setups used, despite both systems presenting different residence times and illuminated volume to total volume ratios, we may assume that the smaller diameter of the Solarbox reactor accounts significantly for this difference, as the light pathway length is much smaller, and so there are more photons failing to be absorbed by TiO2 particles. 3.4 Conclusions Simulated solar photocatalysis using suspended TiO2 showed a good performance for the elimination of the antibiotic Oxytetracycline from aqueous solutions, not only enhancing its mineralization and biodegradability, but also decreasing its toxicity. OTC degradation conditions depend considerably on catalyst concentration and initial pH levels showed to be [TiO2] = 0.5 g L-1 and free initial pH. OTC molecule degradation pathways were also proposed. The major difference between experiments using the best conditions in lab-scale using simulated solar radiation and in pilot plant scale using natural solar radiation regarded reactor geometry, demonstrating that CPC reactors were more effective in the usage of accumulated UV energy to completely degrade and almost completely mineralize OTC. Chapter 3 72 3.5 References Abellán, M.N., Bayarri, B., Giménez, J., Costa, J., 2007. Photocatalytic degradation of sulfamethoxazole in aqueous suspension of TiO2. Appl. Catal., B 74, 233-241. Alexy, R., Kümpel, T., Kümmerer, K., 2004. Assessment of degradation of 18 antibiotics in the Closed Bottle Test. Chemosphere 57, 505-512. Backhaus, T., Grimme, L.H., 1999. The toxicity of antibiotic agents to the luminescent bacterium Vibrio fischeri. Chemosphere 38, 3291-3301. Bahnemann, D., 2004. Photocatalytic water treatment: Solar energy applications. Sol. Energy 77, 445-459. Daughton, C.G., Ternes, T.A., 1999. Pharmaceuticals and personal care products in the environment: Agents of subtle change? Environ. Health Perspect. 107, 907-938. Doi, A.M., Stoskopf, M.K., 2000. The kinetics of oxytetracycline degradation in deionized water under varying temperature, pH, light, substrate, and organic matter. J. Aquat. Anim. Health 12, 246-253. Gogate, P.R., Pandit, A.B., 2004a. A review of imperative technologies for wastewater treatment I: Oxidation technologies at ambient conditions. Adv. Environ. Res. 8, 501-551. Gogate, P.R., Pandit, A.B., 2004b. A review of imperative technologies for wastewater treatment II: Hybrid methods. Adv. Environ. Res. 8, 553-597. González, O., Sans, C., Esplugas, S., 2007. Sulfamethoxazole abatement by photo-Fenton. Toxicity, inhibition and biodegradability assessment of intermediates. J. Hazard. Mater. 146, 459-464. Halling-Sørensen, B., 2001. Inhibition of aerobic growth and nitrification of bacteria in sewage sludge by antibacterial agents. Arch. Environ. Contam. Toxicol. 40, 451-460. Halling-Sørensen, B., Nors Nielsen, S., Lanzky, P.F., Ingerslev, F., Holten Lützhøft, H.C., Jørgensen, S.E., 1998. Occurrence, fate and effects of pharmaceutical substances in the environment - A review. Chemosphere 36, 357-393. Ikehata, K., Jodeiri Naghashkar, N., Gamal El-Din, M., 2006. Degradation of aqueous pharmaceuticals by ozonation and advanced oxidation processes: A review. Ozone Sci. Eng. 28, 353-414. Jeong, J., Song, W., Cooper, W.J., Jung, J., Greaves, J., 2010. Degradation of tetracycline antibiotics: Mechanisms and kinetic studies for advanced oxidation/reduction processes. Chemosphere 78, 533-540. Jiao, S., Zheng, S., Yin, D., Wang, L., Chen, L., 2008. Aqueous oxytetracycline degradation and the toxicity change of degradation compounds in photoirradiation process. J. Environ. Sci. 20, 806-813. Kemper, N., 2008. Veterinary antibiotics in the aquatic and terrestrial environment. Ecol. Indicators 8, 1-13. Kümmerer, K., 2009. Antibiotics in the aquatic environment - A review - Part I. Chemosphere 75, 417-434. Chapter 3 73 Kümmerer, K., Al-Ahmad, A., Mersch-Sundermann, V., 2000. Biodegradability of some antibiotics, elimination of the genotoxicity and affection of wastewater bacteria in a simple test. Chemosphere 40, 701-710. Larsen, T.A., Lienert, J., Joss, A., Siegrist, H., 2004. How to avoid pharmaceuticals in the aquatic environment. J. Biotechnol. 113, 295-304. Li, K., Yediler, A., Yang, M., Schulte-Hostede, S., Wong, M.H., 2008. Ozonation of oxytetracycline and toxicological assessment of its oxidation by-products. Chemosphere 72, 473-478. Loftin, K.A., Adams, C.D., Meyer, M.T., Surampalli, R., 2008. Effects of ionic strength, temperature, and pH on degradation of selected antibiotics. J. Environ. Qual. 37, 378-386. Malato, S., Blanco, J., Vidal, A., Alarcón, D., Maldonado, M.I., Cáceres, J., Gernjak, W., 2003. Applied studies in solar photocatalytic detoxification: An overview. Solar Energy 75, 329-336. Malato, S., Fernández-Ibáñez, P., Maldonado, M.I., Blanco, J., Gernjak, W., 2009. Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catal. Today 147, 1-59. Marco, A., Esplugas, S., Saum, G., 1997. How and why combine chemical and biological processes for wastewater treatment. Water Sci. Technol. 35, 321-327. Miège, C., Choubert, J.M., Ribeiro, L., Eusèbe, M., Coquery, M., 2009. Fate of pharmaceuticals and personal care products in wastewater treatment plants - Conception of a database and first results. Environ. Pollut. 157, 1721-1726. Muñoz, I., Rieradevall, J., Torrades, F., Peral, J., Domènech, X., 2005. Environmental assessment of different solar driven advanced oxidation processes. Sol Energy 79, 369-375. Okuda, T., Kobayashi, Y., Nagao, R., Yamashita, N., Tanaka, H., Tanaka, S., Fujii, S., Konishi, C., Houwa, I., 2008. Removal efficiency of 66 pharmaceuticals during wastewater treatment process in Japan Water Sci. Technol. 57, 65-71. Pouliquen, H., Delépée, R., Larhantec-Verdier, M., Morvan, M.L., Le Bris, H., 2007. Comparative hydrolysis and photolysis of four antibacterial agents (oxytetracycline oxolinic acid, flumequine and florfenicol) in deionised water, freshwater and seawater under abiotic conditions. Aquaculture 262, 23-28. Poyatos, J.M., Muñio, M.M., Almecija, M.C., Torres, J.C., Hontoria, E., Osorio, F., 2009. Advanced Oxidation Processes for Wastewater Treatment: State of the Art. Water, Air, Soil Pollut. 1-18. Rodríguez, S.M., Gálvez, J.B., Rubio, M.I.M., Ibáñez, P.F., Padilla, D.A., Pereira, M.C., Mendes, J.F., De Oliveira, J.C., 2004. Engineering of solar photocatalytic collectors. Sol. Energy 77, 513-524. Rosenfeldt, E.J., Linden, K.G., Canonica, S., von Gunten, U., 2006. Comparison of the efficiency of ·OH radical formation during ozonation and the advanced oxidation processes O3/H2O2 and UV/H2O2. Water Res. 40, 3695-3704. Chapter 3 74 Sarmah, A.K., Meyer, M.T., Boxall, A.B.A., 2006. A global perspective on the use, sales, exposure pathways, occurrence, fate and effects of veterinary antibiotics (VAs) in the environment. Chemosphere 65, 725-759. van der Grinten, E., Pikkemaat, M.G., van den Brandhof, E.J., Stroomberg, G.J., Kraak, M.H.S., 2010. Comparing the sensitivity of algal, cyanobacterial and bacterial bioassays to different groups of antibiotics. Chemosphere 80, 1-6. Wiebe, J.A., Moore, D.E., 1977. Oxidation photosensitized by tetracyclines. J. Pharm. Sci. 66, 186-189. Zhao, C., Deng, H., 2009. Removal of oxytetracycline in water by UV/Hydrophobic Zeolite loaded with TiO2. J. Tongji Univ. 37, 1360-1365. Zhao, C., Deng, H., Li, Y., Liu, Z., 2010. Photodegradation of oxytetracycline in aqueous by 5A and 13X loaded with TiO2 under UV irradiation. J. Hazard. Mater. 176, 884-892. Zhao, C., Deng, H.P., Shang, R., 2009. Removal of oxytetracycline in water by improved UV disinfection process. J Civ. Arch Env. Eng. 31, 152-156. 75 4 Insights into Solar TiO2-Assisted Photocatalytic Oxidation of Two Antibiotics Employed in Aquatic Animal Production, Oxolinic acid and Oxytetracycline In this study, solar driven, TiO2-assisted heterogeneous photocatalytic experiments in a pilot-plant with compound parabolic collectors (CPCs) were carried out to study the degradation of two authorized veterinary antibiotics with particular relevance in finfish aquaculture, Oxolinic Acid (OXA) and Oxytetracycline (OTC), using pure solutions of individual or mixed antibiotics. Firstly, the influence of natural solar photolysis was assessed for each antibiotic. Secondly, photocatalytic degradation kinetic rate constants for individual and mixed antibiotics were compared, using a catalyst load of 0.5 g L-1 and an initial pH around 7.5. Thirdly, for individually photocatalytic-treated OXA and OTC in the same conditions, the growth inhibition of Escherichia coli DSM 1103 was followed, and the mineralization extent was assessed by the residual dissolved organic carbon (DOC), lowmolecular-weight carboxylate anions and inorganic ions concentration. Finally, the effect of inorganic ions, such as chlorides, sulfates, nitrates, phosphates, ammonium and bicarbonates, on the photocatalytic degradation of individual solutions of OXA and OTC was also evaluated and the formation of different reactive oxygen species were probed using selective scavengers. This Chapter is based on the research article “João H.O.S. Pereira, Ana C. Reis, Daniel Queirós, Maria T. Borges, Olga C. Nunes, Vítor J. P. Vilar, Rui A. R. Boaventura. Insights into Solar TiO2-Assisted Photocatalytic Oxidation of Two Antibiotics Employed in Aquatic Animal Production, Oxolinic acid and Oxytetracycline. Sci Tot Environ (2013) 274-283”. Chapter 4 76 Chapter 4 77 4.1 Introduction The resulting environmental contamination by the use of antibiotics in a wide range of human activities has been receiving special attention in recent years (Sarmah et al., 2006; Li et al., 2008; Verlicchi et al., 2010; Michael et al., 2013). A particular case of major concern lies on the use of these substances as a means to control infectious diseases in animal production. Aquaculture, the controlled production of aquatic organisms, is an important food industry that represents a special situation as, in contrast to terrestrial farms, therapeutic agents are most often directly added to the water (Burka et al., 1997; Cabello, 2006; Noga, 2010). Freshwater or marine water finfish aquaculture practices encompass open or closed systems, and according to the suitability of each case, antibiotics may be administered by medicated feed pellets, bath treatments or individual injection (Rigos and Troisi, 2005). Individual injection of large amounts of fish is mostly impractical, while oral administration is considered the main route of fish medication (Samuelsen, 2006). As most of these systems are interconnected with surrounding water bodies, continuous or intermittent wastewater discharges containing unconsumed food and animal faeces, may result in their contamination (Rico et al., 2012), contributing to the development of antibiotic (multi)resistant bacteria resistant strains of bacteria and potential drug accumulation and toxicity to aquatic fauna and flora (Sapkota et al., 2008; Buschmann et al., 2012; Henríquez-Núñez et al., 2012). In the particular case of bath treatments, recommended for recirculating aquaculture systems (due to biofilter bacteria drug sensitivity) or small sized facilities, fishes are moved to smaller containers (20 L, for instance (Samuelsen and Ervik, 2001)) and put directly in contact with antibiotic solution during the prescribed treatment time. Antibiotic concentrations used may range from 10 to 200 mg L-1 (Samuelsen, 2006) and the resulting effluents are either commonly diluted until residues reach acceptable disposal levels or filtered through activated carbon treatment units (Aitcheson et al., 2000; 2001; Noga, 2010), which are not destructive methods. Therefore, alternative and proper means of disposal have to be proposed before safely discharging these effluents into surrounding aquatic systems. The use of Advanced Oxidation Processes (AOPs) to degrade antibiotics has been showing promising results in recent years (Uslu and Balcioglu, 2008; González et al., 2009; Rozas et al., 2010). AOPs are characterized by the production of •OH radicals, which are very reactive and non-selective, leading to largely satisfactory results in the mineralization of pollutants to CO2, water and inorganic compounds, or in their partial degradation to less harmful and/or more biodegradable compounds (Malato et al., 2009; Oller et al., 2011). Chapter 4 78 Heterogeneous photocatalysis using suspended TiO2 is one AOP of special interest due to the chemical stability of the photocatalyst, low cost, and ability of using the small percentage of the ultraviolet radiation coming from the sun, which may also greatly reduce the operation costs. The use of Compound Parabolic Collectors (CPC) greatly enhances the efficiency of this process (Rodríguez et al., 2004; Colina-Márquez et al., 2010) as it increases the amount of incident photons that can be used to degrade target substances. In this work, the degradation of Oxolinic Acid and Oxytetracycline (two of the most extensively used antibiotics in aquaculture (Rigos and Troisi, 2005; Alday-Sanz et al., 2012)), in aqueous solution, was studied, individually and in a mixture, using heterogeneous photocatalysis with suspended TiO2 at pilot-plant scale under natural solar radiation. The antibiotics’ UV normalized absorbance spectra, speciation diagrams as a function of pH and schemes of dissociation equilibrium are shown in Figure 4.1. Despite the fact that some photocatalytic studies (optimization of process parameters such as catalyst load, pH, etc.) have already been made with these substances (Palominos et al., 2008; Giraldo et al., 2010; Zhao et al., 2010; Pereira et al., 2011), none so far has been carried out at this scale and in a situation of mutual presence, a relevant issue given the natural co-occurrence of different antibiotic residues whose relevance results from the natural occurrence and/or application of antibiotic mixtures (Avisar et al., 2010). Moreover, for each individual antibiotic, the influence of natural solar photolysis was also assessed, the antibacterial activity followed, and the mineralization extent quantified (in terms of Dissolved Organic Carbon (DOC), low-molecular-weight carboxylate ions and inorganic ions present in the solutions). Furthermore, the influence of commonly occurring inorganic ions in fresh and marine fish farms waters, and of reactive oxygen species scavengers was also determined in a lab-scale CPC photoreactor using artificial sunlight. The use of a figure-of-merit has been provided for the direct comparison of the solar energy efficiency of the photocatalytic process independent of the nature of the system (Bolton et al., 2001). Chapter 4 79 Figure 4.1. a) Normalized absorbance spectra of OXA (blue dotted line) and OTC (black dashed line) at pH = 7.5; solar UV spectrum (yellow solid line) adapted from Malato et al. (2002); b) OXA speciation diagram as a function of pH, including schematics of dissociation equilibrium (pKa value from Jiménez-Lozano et al. (2002). Ionic strength = 0 M, T = 25°C); c) OTC speciation diagram as a function of pH and d) OTC dissociation equilibrium (pKa values from Qiang and Adams (2004), Ionic strength = 0 M, T = 23°C). 4.2 Materials and Methods All the chemicals and reagents used in this work, the detailed description of the lab-scale and pilot-plant scale experimental units, along with the corresponding experimental procedures followed, and, finally, the employed analytical methods can be consulted in Chapter 2. 240 280 320 360 400 0.0 0.5 1.0 1.5 2.0 Normalized absorbance (A. u.) Wavelenght (nm) 0.0 0.1 0.2 0.3 0.4 0.5 0.6 Solar UV spectrum (W m-2 nm-1) a) pKa1 = 6.92 HOXA OXAH+ H3OTC+ H2OTC HOTCOTC2pKa1 = 3.22 pKa2 = 7.46 pKa3 = 8.94 H3OTC+ H2OTC HOTCOTC2b) c) d) H+ H+ H+ Chapter 4 86 the presence of PO43-, which strongly adsorbs on the surface of TiO2 (Chen et al., 2003) and whose negative charges results in a repulsion effect between TiO2 and both antibiotics. Table 4.2. Pseudo-first-order kinetic parameters simulated solar photocatalytic degradation experiments of OXA and OTC, alone or with (+) inorganic ions and scavengers; [TiO2] = 0.5 g L-1; pH = 7.5; ([OXA]0 = [OTC]0 = 20 mg L-1). Antibiotic Experiment Lab-scale photoreactor (I = 44 W m-2) k (L kJ-1)a r0 (mg kJ-1)b R2 SR2 (mg2 L-2) OXA individual 1.09 ± 0.07 18 ± 1 0.989 0.425 + 1.0 g L-1 Cl1.33 ± 0.08 22 ± 1 0.995 0.285 + 1.0 g L-1 NO31.29 ± 0.04 21.5 ± 0.7 0.998 0.068 + 1.0 g L-1 HCO31.16 ± 0.03 20.7 ± 0.5 0.998 0.088 + 1.0 g L-1 SO421.04 ± 0.03 19.0 ± 0.5 0.998 0.121 + 1.0 g L-1 NH4+ 1.5 ± 0.1 25 ± 2 0.985 0.696 + 1.0 g L-1 PO430.83 ± 0.02 14.9 ± 0.4 0.997 0.117 + 10 mM NaN3 0.375 ± 0.008 6.7 ± 0.1 0.996 0.099 + 50 mM D-Mannitol 0.25 ± 0.03 4.2 ± 0.5 0.926 1.147 OTC individual 4.3 ± 0.4 59 ± 6 0.993 0.293 + 1.0 g L-1 Cl2.74 ± 0.08 44 ± 1 0.999 0.067 + 1.0 g L-1 NO32.48 ± 0.02 36.1 ± 0.3 0.999 0.006 + 1.0 g L-1 HCO34.11 ± 0.08 68 ± 1 0.999 0.022 + 1.0 g L-1 SO421.93 ± 0.03 32.0 ± 0.5 0.999 0.041 + 1.0 g L-1 NH4+ 3.1 ± 0.1 47 ± 2 0.998 0.075 + 1.0 g L-1 PO431.25 ± 0.01 21.2 ± 0.1 0.999 0.006 + 10 mM NaN3 0.97 ± 0.07 19 ± 1 0.969 1.551 + 50 mM D-Mannitol 0.59 ± 0.03 10.2 ±0.5 0.970 0.467 a – pseudo-first order kinetic rate, b - initial reaction rate Selective scavengers D-Mannitol and sodium azide (NaN3) were used to assess the role of reactive oxygen species, hydroxyl radical (•OH) and singlet oxygen (1O2), on the photocatalytic degradation of OXA and OTC under the same tested conditions. From the analysis of Figure 4.7 (a) and (b), it can be seen that, compared to the absence of reactive species scavengers, OXA and OTC degradation is mainly attributed to the •OH attack, while 1O2 plays a secondary, but non-negligible role in OXA and OTC self-sensitization under UV solar light. The remainder of the degradation may be explained by the participation of the abovementioned photolysis effect or by the participation of direct oxidation by photo-generated holes of molecules adsorbed on the catalyst surface (Giraldo et al., 2010; Zhao et al., 2013). Chapter 4 87 Figure 4.7. Removal profiles of a) OXA () and b) OTC (), alone and in the presence of 10 mM NaN3 () and 50 mM D-mannitol () under simulated solar photocatalysis with 0.5 g L-1 TiO2, pH = 7.5 ([OXA]0 = [OTC]0 = 20 mg L-1). 4.3.5 Solar photocatalytic efficiency index Valuable figures-of-merit have been recommended by the IUPAC to allow for a direct comparison between the electricor solarenergy efficiency of different AOPs, independently of the nature of the used system. In the case of solar-based AOPs, where the cost of incident solar radiation is zero, the collector area is considered as the main capital cost. As the capital cost of the solar collectors are proportionally related to their area, it is suitable to use figures of merit based on the solar collector area. Therefore, for a low pollutant concentration range, the appropriate figure-of-merit is the collector area per order ( CO A ) (Bolton et al., 2001). CO A is the collector area required to reduce the concentration of a contaminant (C) in polluted water in a unit of volume by one order of magnitude in a time ( 0 t = 1 h) when the standardized incident solar irradiance ( 0 S E ) is 1000 W m−2. The CO A (m2 m-3-order), in batch operation, can be calculated from Eq. 3.01:            0 0 0log C C VtE tUVA A tS G r CO (3.01) where Ar is the illuminated collector surface area (m2), G UV is the average solar ultraviolet irradiance (W m-2) over the period t of the treatment (h), t V the total reactor volume (m3), and C and 0 C are the final and initial antibiotic concentrations (mg L-1, or mM in the case of the antibiotic mixture). This figure-of-merit is inversely proportional to fundamental efficiency factors, e.g., when CO A values increase, there is a loss in the system efficiency. In the case of the two individual OXA photocatalytic experiments, at 20 and 40 mg L-1, the CO A index varied from 0.24 to 1.15 m2 m-3-order, respectively. Even though there was a near-5 fold decrease in 0.0 0.5 1.0 1.5 2.0 2.5 3.0 0.0 0.2 0.4 0.6 0.8 1.0 a) Oxolinic Acid (C/C0) QUV (kJ L-1) RAD-ON 0.0 0.5 1.0 1.5 2.0 0.0 0.2 0.4 0.6 0.8 1.0 b) Oxytetracycline (C/C0) QUV (kJ L-1) RAD-ON Chapter 4 88 efficiency, it should be mentioned that G UV values were nearly doubled between experiments (22.7 and 42.1 W m-2, respectively) and so it was the necessary treatment time (0.5 and 1 h). In turn, the CO A index for the two individual OTC photocatalytic experiments were 0.15 and 0.31 m2 m-3-order for the individual 20 and 40 mg L-1 experiments. The G UV values were very similar (30.2 and 35.5 W m-2, respectively), while the necessary time was doubled (0.33 and 0.5 h). In the combined antibiotic experiment, after adding up the OXA and OTC initial and final concentrations in terms of mM for the calculation of 0 C and C , the CO A index was 0.49 m2 m-3order ( G UV = 22.3 W m-2 and t = 1 h). A similar CO A index, 0.4 m2 m-3-order, was obtained by Sousa et al. (2013) in the TiO2-assisted photocatalytic oxidation of the anxiolytic drug lorazepam with the same pilot-plant, but using only 0.2 g L-1 TiO2. These results bring into consideration the viability of performing this treatment on these compounds (individually or mixed) on different times of the year without substantial conversion of efficiency losses. Chapter 4 89 4.4 Conclusions Solar photocatalysis with TiO2 was shown to be 100% efficient in the degradation of individual OXA and OTC solutions (C0 = 20 mg L-1) and mixtures, being necessary low amounts of accumulated UV energy per liter of solution (QUV ~ 1 kJUV L-1), and also in the extensive mineralization of the remaining DOC content, as compared to solar photolysis alone. According to the pseudo-first-order kinetic rate constants, the individual removal of OTC fares better than that of OXA. Nevertheless, when both antibiotics are mixed, they disappear from solution quite simultaneously and reach a proper amount of mineralization shortly afterwards. Furthermore, within the operational conditions studied, the efficiency of the solar photocatalytic CPC reactor has shown to be consistent. A detailed assessment of the antibacterial activity and formation of carboxylic acids and inorganic ions showed that after the point of complete antibiotic abatement, the remaining organics do not inhibit reference bacterial growth and consist largely of biodegradable substances. Although the OXA and OTC degradation was mainly attributed to hydroxyl radicals, singlet oxygen also plays an important role in antibiotics self-photosensitization under UV/visible solar light. The presence of PO43considerably affects the photocatalytic efficiency and must be taken into consideration when applying TiO2 photocatalysis in complex water matrixes (Raja et al., 2005), while other inorganic ions (Cl-, SO42-, NO3-, NH4+ and HCO3-) do not substantially hinder complete OXA and OTC removal. This treatment could be proposed as a viable alternative to treat effluents resulting from high antibiotic-using activities such as bath treatments in fish farms. Effluents could be treated locally, requiring only short periods of solar exposure, and safely discharged after a proper catalyst recovery step. Chapter 4 90 4.5 References Abdullah, M., 1990. Effects of common inorganic anions on rates of photocatalytic oxidation of organic carbon over illuminated titanium dioxide. J. Phys. Chem. 94, 6820-6825. Ahmed, S., Rasul, M.G., Martens, W.N., Brown, R., Hashib, M.A., 2011. Advances in heterogeneous photocatalytic degradation of phenols and dyes in wastewater: A review. Water, Air, Soil Pollut. 215, 3-29. Barreiros, L., Nogales, B., Manaia, C.M., Silva Ferreira, A.C., Pieper, D.H., Reis, M.A., Nunes, O.C., 2003. A novel pathway for mineralization of the thiocarbamate herbicide molinate by a defined bacterial mixed culture. Environ. Microbiol. 5, 944-953. Bayarri, B., Abellán, M.N., Giménez, J., Esplugas, S., 2007. Study of the wavelength effect in the photolysis and heterogeneous photocatalysis. Catal. Today 129, 231-239. Boaventura, R., Pedro, A.M., Coimbra, J., Lencastre, E., 1997. Trout farm effluents: Characterization and impact on the receiving streams. Environ. Pollut. 95, 379-387. Bolton, J.R., Bircher, K.G., Tumas, W., Tolman, C.A., 2001. Figures-of-merit for the technical development and application of advanced oxidation technologies for both electricand solardriven systems. Pure Appl. Chem. 73, 627-637. Boreen, A.L., Arnold, W.A., McNeill, K., 2004. Photochemical fate of sulfa drugs in then aquatic environment: Sulfa drugs containing five-membered heterocyclic groups. Environ. Sci. Technol. 38, 3933-3940. Calza, P., Pelizzetti, E., Minero, C., 2005. The fate of organic nitrogen in photocatalysis: An overview. J. Appl. Electrochem. 35, 665-673. Chen, F., Zhao, J., Hidaka, H., 2003. Adsorption factor and photocatalytic degradation of dyeconstituent aromatics on the surface of TiO2 in the presence of phosphate anions. Res. Chem. Intermed. 29, 733-748. Colina-Márquez, J., MacHuca-Martínez, F., Puma, G.L., 2010. Radiation absorption and optimization of solar photocatalytic reactors for environmental applications. Environ. Sci. Technol. 44, 5112-5120. Faria, P.C.C., Órfão, J.J.M., Pereira, M.F.R., 2008. Activated carbon catalytic ozonation of oxamic and oxalic acids. App. Cat. B: Environ. 79, 237-243. Fernández-Ibáñez, P., Blanco, J., Malato, S., De Las Nieves, F.J., 2003. Application of the colloidal stability of TiO2 particles for recovery and reuse in solar photocatalysis. Water Res. 37, 3180-3188. Garcia-Segura, S., Brillas, E., 2011. Mineralization of the recalcitrant oxalic and oxamic acids by electrochemical advanced oxidation processes using a boron-doped diamond anode. Water Res. 45, 2975-2984. Giraldo, A.L., Peñuela, G.A., Torres-Palma, R.A., Pino, N.J., Palominos, R.A., Mansilla, H.D., 2010. Degradation of the antibiotic oxolinic acid by photocatalysis with TiO2 in suspension. Water Res. 44, 5158-5167. Guillard, C., Lachheb, H., Houas, A., Ksibi, M., Elaloui, E., Herrmann, J.M., 2003. Influence of chemical structure of dyes, of pH and of inorganic salts on their photocatalytic degradation by Chapter 4 91 TiO2 comparison of the efficiency of powder and supported TiO2. J. Photochem. Photobiol. A: Chem. 158, 27-36. Guillard, C., Puzenat, E., Lachheb, H., Houas, A., Herrmann, J.M., 2005. Why inorganic salts decrease the TiO2 photocatalytic efficiency. Int. J. of Photoener 7, 1-9. Han, S.K., Hwang, T.M., Yoon, Y., Kang, J.W., 2011. Evidence of singlet oxygen and hydroxyl radical formation in aqueous goethite suspension using spin-trapping electron paramagnetic resonance (EPR). Chemosphere 84, 1095-1101. Hirakawa, K., Mori, M., Yoshida, M., Oikawa, S., Kawanishi, S., 2004. Photo-irradiated titanium dioxide catalyzes site specific DNA damage via generation of hydrogen peroxide. Free Radical Res. 38, 439-447. Jiao, S., Zheng, S., Yin, D., Wang, L., Chen, L., 2008. Aqueous photolysis of tetracycline and toxicity of photolytic products to luminescent bacteria. Chemosphere 73, 377-382. Jiménez-Lozano, E., Marqués, I., Barrón, D., Beltrán, J.L., Barbosa, J., 2002. Determination of pKa values of quinolones from mobility and spectroscopic data obtained by capillary electrophoresis and a diode array detector. Anal. Chim. Acta 464, 37-45. Malato, S., Fernández-Ibáñez, P., Maldonado, M.I., Blanco, J., Gernjak, W., 2009. Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catal. Today 147, 1-59. Miller, R.A., Walker, R.D., Carson, J., Coles, M., Coyne, R., Dalsgaard, I., Gieseker, C., Hsu, H.M., Mathers, J.J., Papapetropoulou, M., Petty, B., Teitzel, C., Reimschuessel, R., 2005. Standardization of a broth microdilution susceptibility testing method to determine minimum inhibitory concentrations of aquatic bacteria. Dis. Aquat. Org. 64, 211-222. Miralles-Cuevas, S., Arqués, A., Maldonado, M.I., Sánchez-Pérez, J.A., Malato Rodríguez, S., 2013. Combined nanofiltration and photo-Fenton treatment of water containing micropollutants. Chem. Eng. J. 224, 89-95. Oller, I., Malato, S., Sánchez-Pérez, J.A., 2011. Combination of Advanced Oxidation Processes and biological treatments for wastewater decontamination - A review. Sci. Tot. Environ. 409: 4141-4166. Palominos, R.A., Mora, A., Mondaca, M.A., Pérez-Moya, M., Mansilla, H.D., 2008. Oxolinic acid photo-oxidation using immobilized TiO2. J. Hazard. Mater. 158, 460-464. Pereira, J.H.O.S., Vilar, V.J.P., Borges, M.T., González, O., Esplugas, S., Boaventura, R.A.R., 2011. Photocatalytic degradation of oxytetracycline using TiO2 under natural and simulated solar radiation. Sol. Energy 85, 2732-2740. Pouliquen, H., Delépée, R., Larhantec-Verdier, M., Morvan, M.L., Le Bris, H., 2007. Comparative hydrolysis and photolysis of four antibacterial agents (oxytetracycline oxolinic acid, flumequine and florfenicol) in deionised water, freshwater and seawater under abiotic conditions. Aquaculture 262, 23-28. Qiang, Z., Adams, C., 2004. Potentiometric determination of acid dissociation constants (pKa) for human and veterinary antibiotics. Water Res. 38, 2874-2890. Raja, P., Bozzi, A., Mansilla, H., Kiwi, J., 2005. Evidence for superoxide-radical anion, singlet oxygen and OH-radical intervention during the degradation of the lignin model compound (3methoxy-4-hydroxyphenylmethylcarbinol). J. Photochem. Photobiol. A: Chem. 169, 271-278. Chapter 4 92 Sirtori, C., Zapata, A., Malato, S., Gernjak, W., Fernández-Alba, A.R., Agüera, A., 2009. Solar photocatalytic treatment of quinolones: Intermediates and toxicity evaluation. Photochemical and Photobiological Sciences 8, 644-651. Soares, P.A., Silva, T.F.C.V., Manenti, D.R., Souza, S.M.A.G.U., Boaventura, R.A.R., Vilar, V.J.P., 2013. Insights into real cotton-textile dyeing wastewater treatment using solar advanced oxidation processes. Enviro. Sci. Pollut. Res. 1-14. Sousa, M.A., Gonçalves, C., Pereira, J.H.O.S., Vilar, V.J.P., Boaventura, R.A.R., Alpendurada, M.F., 2013. Photolytic and TiO2-assisted photocatalytic oxidation of the anxiolytic drug lorazepam (Lorenin® pills) under artificial UV light and natural sunlight: A comparative and comprehensive study. Sol. Energy 87, 219-228. Toral, M.I., Orellana, S.L., Soto, C.A., Richter, P., 2011. Extraction and Determination of Oxytetracycline Hydrochloride and Oxolinic Acid in Fish Feed by Derivative Spectrophotometry of First Order. Food Anal. Method 1-8. Yahiat, S., Fourcade, F., Brosillon, S., Amrane, A., 2011. Removal of antibiotics by an integrated process coupling photocatalysis and biological treatment - Case of tetracycline and tylosin. Int. Biodeterior. Biodegrad. 65, 997-1003. Zhao, C., Pelaez, M., Duan, X., Deng, H., O'Shea, K., Fatta-Kassinos, D., Dionysiou, D.D., 2013. Role of pH on photolytic and photocatalytic degradation of antibiotic oxytetracycline in aqueous solution under visible/solar light: Kinetics and mechanism studies. App. Cat. B Environ. 134–135, 83-92. 93 5 Assessment of Solar Driven TiO2-Assisted Photocatalysis Efficiency on Amoxicillin Degradation The objective of this work was to evaluate the efficiency of a solar TiO2-assisted photocatalytic process on Amoxicillin (AMX) degradation. Firstly, solar photolysis of AMX was compared with solar photocatalysis in a CPC pilot scale photoreactor. Another experiment was also carried out to accurately follow the antibacterial activity against Escherichia coli DSM 1103 and Staphylococcus aureus DSM 1104 and mineralization of AMX by tracing the contents of dissolved organic carbon (DOC), low-molecular-weight carboxylate anions and inorganic anions. Finally, the influence of individual inorganic ions on AMX photocatalytic degradation efficiency and the involvement of some reactive oxygen species were also assessed. This Chapter is based on the research article “João H.O.S. Pereira, Ana C. Reis, Olga C. Nunes, Maria T. Borges, Vítor J. P. Vilar, Rui A. R. Boaventura. Assessment of Solar Driven TiO2-Assisted Photocatalysis Efficiency on Amoxicillin Degradation. Environ Sci Pollut Res (2014) 21:1292-1303”. Chapter 5 94 Chapter 5 95 5.1 Introduction The most widely used group of antibiotics in human medicine in Europe is the penicillin’s, which have showed an increasing trend of prescription over the last decade (Versporten et al. 2011). In 2009 they represented 47% of total outpatient use, from which the use of amoxicillin (AMX) alone, or combined with β-lactamase inhibitors, represented 83.9% (Versporten et al. 2011). The combination of this high level of prescription and the fact that amoxicillin has a low metabolic rate in humans, leading to excretion rates of 80-90% (Hirsch et al. 1999), accounts for an ubiquitous presence of this pharmaceutical in domestic and hospital wastewaters (Andreozzi et al. 2004; Längin et al. 2009; Watkinson et al. 2009; Leung et al. 2012). Adding to the varying efficiencies reported for the removal of antibiotics in conventional wastewater treatment plants (WWTP), which were not specifically designed to remove them (Leung et al. 2012), amoxicillin, as well as other antibacterial drugs, are frequently found in rivers and other water bodies receiving treated WWTP effluents (Kasprzyk-Hordern et al. 2008). Another source of antibiotic contamination may result from animal husbandry use, as is the case of fish farming activities (Lalumera et al. 2004; Rigos and Troisi 2005), where effluents resulting from bath immersion treatments using amoxicillin concentrations as high as 200 mg L-1 (Mitchell and Rodger 2011) require special consideration before disposal. These are often diluted in aquatic receiving bodies or often treated with non-destructive methods (Noga 2010). Despite the fact that some of these substances do not persist due to natural degradation processes (Jones et al. 2005), their continual release to the environment raises several concerns regarding their ecotoxicological potential to humans and animals (Kim and Aga 2007; Martinez 2009; Santos et al. 2010; Escher et al. 2011). Several alternative treatments have been proposed in recent years to tackle this situation, which are summarized in the revision performed by Homem and Santos (2011) and Michael et al. (2013). Among them, Advanced Oxidation Processes (AOPs) comprise different processes of generating hydroxyl radicals (•OH), which are very reactive and not highly selective. AOPs can be divided in photochemical (UV, UV/O3, UV/H2O2), photocatalytic (TiO2/UV, Fe2+/H2O2/UV-Vis) or chemical oxidation processes (O3, O3/H2O2, H2O2/Fe2+) (Poyatos et al. 2009). Their operation for full chemical degradation can be rather costly, so research is being focused on the application of AOPs that rely on solar irradiation as the source of UV radiation (such as heterogeneous photocatalysis and photo-Fenton reaction) and the possible combination with a pre or post biological treatment step (Malato et al. 2009; Oller et al. 2011). Chapter 5 102 Figure 5.5. Removal profiles of 20 mg L-1 of AMX alone () and a) in the presence of 1 g L-1 of Cl- (), SO42- (), NO3- (▼), NH4+ (), PO43- (), 0.1 g L-1 HCO3- () and b) in the presence of 10 mM NaN3 () and 50 mM D-mannitol () under simulated solar photocatalysis with 0.5 g L-1 TiO2 and pH = 7.5. Consequently, the aforementioned effect of the presence of most inorganic ions did not substantially affect the AMX photocatalytic kinetic rates (Table 5.1). The only considerable exception is the case of PO43-, whose presence not only promoted the initial dark adsorption of AMX onto the catalyst surface, but consequently enhanced the oxidation rate of AMX compared to its absence (Figure 5.5a). Chen et al. (2003) have reported a similar phenomenon of increased adsorption of certain dye-constituent aromatics on the surface of TiO2 in the presence of phosphate anions. The negatively charged phosphate anions strongly adsorbed onto the TiO2 surface may favor not only the adsorption of AMX species bearing a positive (Chen et al. 2003) or neutral (Guillard et al. 2003) charge on the –NH2 group, but also the formation of free hydroxyl radicals via the enhancement of the separation of the photogenerated hole and electron facilitated by an inner-sphere surface complex, as suggested by Zhao and co-workers (2008). The same authors thus concluded that phosphate modification accelerates the degradation of pollutants either more prone to hydroxyl radical attack or with weak adsorption on pure TiO2 particles. To evaluate the role of reactive species, such as hydroxyl radical (•OH) and singlet oxygen (1O2), in the AMX photocatalytic degradation, selective scavengers D-Mannitol and sodium azide (NaN3), respectively, were used. The insignificant photocatalytic degradation of the antibiotic during the phototreatment period in the presence of D-Mannitol (Figure 5.5b and Table 5.2), purports the role of •OH radicals as the major responsible for AMX degradation, which is in agreement with the results presented by Song et al. (2008). However, according to Figure 5.5 b, the 1O2 reactive species also plays an important role in AMX selfphotosensitization under UV/visible solar light, as reported by Zhao et al. (2013). 0.0 0.5 1.0 1.5 2.0 2.5 3.0 0.0 0.2 0.4 0.6 0.8 1.0 a) Amoxicillin (C/C0) QUV (kJ L-1) RAD-ON 0.0 0.5 1.0 1.5 2.0 2.5 3.0 0.0 0.2 0.4 0.6 0.8 1.0 b) Amoxicillin (C/C0) QUV (kJ L-1) RAD-ON Chapter 5 103 Table 5.2. Pseudo-first order kinetic constant values for AMX degradation, alone or with (+) inorganic ions and scavengers, under simulated solar TiO2-assisted photocatalytic systems: [TiO2] = 0.5 g L-1; pH = 7.5. Lab-scale Photoreactor (I = 44 WUV m-2) Experiment k (L kJ-1)a r0 (mg kJ-1)b R2 SR2 (mg2 L-2) AMX individual 0.47 ± 0.01 8.4 ± 0.2 0.996 0.108 + 1.0 g L-1 Cl0.47 ± 0.01 8.3 ± 0.2 0.997 0.119 + 1.0 g L-1 NO30.57 ± 0.02 10.5 ± 0.4 0.996 0.171 + 0.1 g L-1 HCO30.47 ± 0.01 8.9 ± 0.2 0.997 0.095 + 1.0 g L-1 SO420.50 ± 0.01 9 ± 2 0.993 0.167 + 1.0 g L-1 NH4+ 0.55 ± 0.03 10.3 ± 0.6 0.985 0.513 + 1.0 g L-1 PO430.669 ± 0.004 10.76 ± 0.06 0.999 0.005 + 10 mM NaN3 0.193 ± 0.004 3.52 ± 0.01 0.831 0.09 + 50 mM D-Mannitol 0.032 ± 0.004 0.59 ± 0.07 0.831 0.09 a – pseudo-first order kinetic rate, b - initial reaction rate 5.3.4 Solar photocatalytic efficiency index In the case of AOPs based on free solar radiation, the main capital cost is proportionally related to the area of the collectors. As such, it is suitable to use a figure-of-merit based on the solar collector area recommended by the IUPAC to allow for a direct comparison between the solarenergy efficiency of different AOPs, independently of the nature of the used system. Therefore, for a low pollutant concentration range, the appropriate figure-of-merit is the collector area per order (ACO) (Bolton et al. 2001). CO A is the collector area required to reduce the concentration of a contaminant (C) in polluted water in a unit of volume by one order of magnitude in a time ( 0 t = 1 h) when the standardized incident solar irradiance ( 0 S E ) is 1000 W m−2. The CO A (m2 m-3-order), in batch operation, can be calculated from Eq. 5.01:            0 0 0log C C VtE tUVA A tS G r CO (5.01) where Ar is the illuminated collector surface area (m2), G UV is the average solar ultraviolet irradiance (W m-2, as measured in the 300-400 nm range, the useful UV wavelength range for solar photocatalysis with TiO2) over the period t of the treatment (h), t V the total reactor volume (m3), and C and 0 C are the final and initial antibiotic concentrations (mg L-1). This figure-of-merit indicates a loss in the system efficiency when ACO values increase, in an inverse proportion. As previously mentioned, CPC geometry has been shown to be highly efficient when applied to solar photocatalytic treatments. Chapter 5 104 This remark has been emphasized by Bandala and Estrada (2007), when they compared the same figure-of-merit using different solar reactors (which use direct and diffuse UV irradiation). Within this study, for the two different initial antibiotic concentrations, 20 and 40 mg L-1, 0.85 ( G UV = 25.2 W m-2) and 1.65 ( G UV = 32.18 W m-2) hours were required to achieve a decrease of the initial AMX concentrations in one order of magnitude, resulting in an ACO index of 0.65 to 1.21 m2 m-3-order, respectively. It should be mentioned that the 30% solar UV irradiance increase between experiments (given that these were performed in different days), was compensated by a decrease of the useful solar exposure time needed to reduce the AMX concentration by an order of magnitude. Therefore, according to the results obtained, there is no substantial conversion efficiency loss in the studied antibiotic concentration range. As expected, a 2 fold increase of pollutant concentration only doubled the ACO index. A similar system efficiency was obtained by Sousa et al. (2013) in the solar photocatalytic oxidation of the anxiolytic drug lorazepam using the same pilot-plant but mediated by only 0.2 g L-1 TiO2 (ACO index = 0.4 m2 m-3-order). Chapter 5 105 5.4 Conclusions The use of CPC photoreactors for solar UV photons capture was shown to be effective for the TiO2-assisted photocatalytic ([TiO2] = 0.5 g L-1) degradation of the antibiotic Amoxicillin in aqueous solutions at neutral pH conditions (7.5). Solar UV radiation alone was unable to attack the antibiotic molecules during the same phototreatment period. The TiO2 solar photocatalysis was able to reduce the antibiotic concentration from 40 to 3.1 mg L-1 after 4.6 kJ of UV accumulated energy per litre of solution, leading to a considerable reduction of the antibacterial activity. At the end of the phototreatment period (11.7 kJ L-1 of UV energy accumulated in the system), 71% mineralization was achieved, being seventy percent of the residual DOC content in the form of low molecular weight carboxylate anions, mainly propionic and maleic acids. Even though the amount of sulfate detected at the end of the reaction corresponds to the stoichiometric conversion of sulfur contained in AMX, only 30% of the initial nitrogen was converted to ammonium, showing that the mineralization of nitrogenated by-products was less effective during the phototreatment period. Although the AMX degradation was mainly attributed to hydroxyl radicals, singlet oxygen also plays an important role in AMX self-photosensitization under UV/visible solar light. Screenings of individual inorganic ions effects have shown that the presence of phosphates at the studied pH level enhance AMX photocatalytic oxidation, while the presence of other inorganic ions (Cl-, SO42-, NO3-, NH4+ and HCO3-) did not considerably alter the reaction rate. The results obtained build on the potential for developing future treatments for wastewaters of various origins containing high concentrations of AMX with relatively short periods of solar irradiance, allowing their subsequent safe discharge due to the elimination of the risk of antibiotic resistance promotion. Chapter 5 106 5.5 References Abdullah, M., 1990. Effects of common inorganic anions on rates of photocatalytic oxidation of organic carbon over illuminated titanium dioxide. J. Phys. Chem. 94, 6820-6825. Ahmed, S., Rasul, M.G., Martens, W.N., Brown, R., Hashib, M.A., 2011. Advances in heterogeneous photocatalytic degradation of phenols and dyes in wastewater: A review. Water, Air, Soil Pollut. 215, 3-29. Barreiros, L., Nogales, B., Manaia, C.M., Silva Ferreira, A.C., Pieper, D.H., Reis, M.A., Nunes, O.C., 2003. A novel pathway for mineralization of the thiocarbamate herbicide molinate by a defined bacterial mixed culture. Environ. Microbiol. 5, 944-953. Bayarri, B., Abellán, M.N., Giménez, J., Esplugas, S., 2007. Study of the wavelength effect in the photolysis and heterogeneous photocatalysis. Catal. Today 129, 231-239. Boaventura, R., Pedro, A.M., Coimbra, J., Lencastre, E., 1997. Trout farm effluents: Characterization and impact on the receiving streams. Environ. Pollut. 95, 379-387. Bolton, J.R., Bircher, K.G., Tumas, W., Tolman, C.A., 2001. Figures-of-merit for the technical development and application of advanced oxidation technologies for both electricand solardriven systems. Pure Appl. Chem. 73, 627-637. Boreen, A.L., Arnold, W.A., McNeill, K., 2004. Photochemical fate of sulfa drugs in then aquatic environment: Sulfa drugs containing five-membered heterocyclic groups. Environ. Sci. Technol. 38, 3933-3940. Calza, P., Pelizzetti, E., Minero, C., 2005. The fate of organic nitrogen in photocatalysis: An overview. J. Appl. Electrochem. 35, 665-673. Chen, F., Zhao, J., Hidaka, H., 2003. Adsorption factor and photocatalytic degradation of dyeconstituent aromatics on the surface of TiO2 in the presence of phosphate anions. Res. Chem. Intermed. 29, 733-748. Colina-Márquez, J., MacHuca-Martínez, F., Puma, G.L., 2010. Radiation absorption and optimization of solar photocatalytic reactors for environmental applications. Environ. Sci. Technol. 44, 5112-5120. Faria, P.C.C., Órfão, J.J.M., Pereira, M.F.R., 2008. Activated carbon catalytic ozonation of oxamic and oxalic acids. App. Cat. B Environ. 79, 237-243. Fernández-Ibáñez, P., Blanco, J., Malato, S., De Las Nieves, F.J., 2003. Application of the colloidal stability of TiO2 particles for recovery and reuse in solar photocatalysis. Water Res. 37, 3180-3188. Fernández, P., Blanco, J., Sichel, C., Malato, S., 2005. Water disinfection by solar photocatalysis using compound parabolic collectors. Catal. Today 101, 345-352. Garcia-Segura, S., Brillas, E., 2011. Mineralization of the recalcitrant oxalic and oxamic acids by electrochemical advanced oxidation processes using a boron-doped diamond anode. Water Res. 45, 2975-2984. Giraldo, A.L., Peñuela, G.A., Torres-Palma, R.A., Pino, N.J., Palominos, R.A., Mansilla, H.D., 2010. Degradation of the antibiotic oxolinic acid by photocatalysis with TiO2 in suspension. Water Res. 44, 5158-5167. González, O., Sans, C., Esplugas, S., Malato, S. 2009. Application of solar advanced oxidation processes to the degradation of the antibiotic sulfamethoxazole. Photochem. Photobiol Sci. 8, 1032-1039 Chapter 5 107 Guillard, C., Lachheb, H., Houas, A., Ksibi, M., Elaloui, E., Herrmann, J.M., 2003. Influence of chemical structure of dyes, of pH and of inorganic salts on their photocatalytic degradation by TiO2 comparison of the efficiency of powder and supported TiO2. J. Photochem. Photobiol. A: Chem. 158, 27-36. Guillard, C., Puzenat, E., Lachheb, H., Houas, A., Herrmann, J.M., 2005. Why inorganic salts decrease the TiO2 photocatalytic efficiency. Int. J. Photoener. 7, 1-9. Han, S.K., Hwang, T.M., Yoon, Y., Kang, J.W., 2011. Evidence of singlet oxygen and hydroxyl radical formation in aqueous goethite suspension using spin-trapping electron paramagnetic resonance (EPR). Chemosphere 84, 1095-1101. Hirakawa, K., Mori, M., Yoshida, M., Oikawa, S., Kawanishi, S., 2004. Photo-irradiated titanium dioxide catalyzes site specific DNA damage via generation of hydrogen peroxide. Free Radical Res. 38, 439-447. Jiao, S., Zheng, S., Yin, D., Wang, L., Chen, L., 2008. Aqueous photolysis of tetracycline and toxicity of photolytic products to luminescent bacteria. Chemosphere 73, 377-382. Jiménez-Lozano, E., Marqués, I., Barrón, D., Beltrán, J.L., Barbosa, J., 2002. Determination of pKa values of quinolones from mobility and spectroscopic data obtained by capillary electrophoresis and a diode array detector. Anal. Chim. Acta 464, 37-45. Malato, S., Blanco J., Vidal, A., Richter, C., 2002. Photocatalysis with solar energy at a pilotplant scale: An overview. App. Cat. B: Environ. 37, 1-15. Malato, S., Fernández-Ibáñez, P., Maldonado, M.I., Blanco, J., Gernjak, W., 2009. Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catal. Today 147, 1-59. Miller, R.A., Walker, R.D., Carson, J., Coles, M., Coyne, R., Dalsgaard, I., Gieseker, C., Hsu, H.M., Mathers, J.J., Papapetropoulou, M., Petty, B., Teitzel, C., Reimschuessel, R., 2005. Standardization of a broth microdilution susceptibility testing method to determine minimum inhibitory concentrations of aquatic bacteria. Dis. Aquat. Org. 64, 211-222. Miralles-Cuevas, S., Arqués, A., Maldonado, M.I., Sánchez-Pérez, J.A., Malato Rodríguez, S., 2012. Combined nanofiltration and photo-Fenton treatment of water containing micropollutants. Chem. Eng. J. Palominos, R.A., Mora, A., Mondaca, M.A., Pérez-Moya, M., Mansilla, H.D., 2008. Oxolinic acid photo-oxidation using immobilized TiO2. J. Hazard. Mater. 158, 460-464. Pereira, J.H.O.S., Vilar, V.J.P., Borges, M.T., González, O., Esplugas, S., Boaventura, R.A.R., 2011. Photocatalytic degradation of oxytetracycline using TiO2 under natural and simulated solar radiation. Sol. Energy 85, 2732-2740. Pouliquen, H., Delépée, R., Larhantec-Verdier, M., Morvan, M.L., Le Bris, H., 2007. Comparative hydrolysis and photolysis of four antibacterial agents (oxytetracycline oxolinic acid, flumequine and florfenicol) in deionised water, freshwater and seawater under abiotic conditions. Aquaculture 262, 23-28. Qiang, Z., Adams, C., 2004. Potentiometric determination of acid dissociation constants (pKa) for human and veterinary antibiotics. Water Res. 38, 2874-2890. Raja, P., Bozzi, A., Mansilla, H., Kiwi, J., 2005. Evidence for superoxide-radical anion, singlet oxygen and OH-radical intervention during the degradation of the lignin model compound (3methoxy-4-hydroxyphenylmethylcarbinol). J. Photochem. Photobiol. A: Chem. 169, 271-278. Sirtori, C., Zapata, A., Malato, S., Gernjak, W., Fernández-Alba, A.R., Agüera, A., 2009. Solar photocatalytic treatment of quinolones: Intermediates and toxicity evaluation. Photochem. Photobiol. Sci. 8, 644-651. Chapter 5 108 Soares, P.A., Silva, T.F.C.V., Manenti, D.R., Souza, S.M.A.G.U., Boaventura, R.A.R., Vilar, V.J.P., 2013. Insights into real cotton-textile dyeing wastewater treatment using solar advanced oxidation processes. Environ. Sci. Pollut. R.1-14. Sousa, M.A., Gonçalves, C., Pereira, J.H.O.S., Vilar, V.J.P., Boaventura, R.A.R., Alpendurada, M.F., 2013. Photolytic and TiO2-assisted photocatalytic oxidation of the anxiolytic drug lorazepam (Lorenin® pills) under artificial UV light and natural sunlight: A comparative and comprehensive study. Sol. Energy 87, 219-228. Toral, M.I., Orellana, S.L., Soto, C.A., Richter, P., 2011. Extraction and Determination of Oxytetracycline Hydrochloride and Oxolinic Acid in Fish Feed by Derivative Spectrophotometry of First Order. Food Anal. Method 1-8. Vilar, V.J.P., Gomes, A.I.E., Ramos, V.M., Maldonado, M.I., Boaventura, R.A.R. 2009. Solar photocatalysis of a recalcitrant coloured effluent from a wastewater treatment plant. Photochem. Photobiol. Sci. 8, 691-698. Yahiat, S., Fourcade, F., Brosillon, S., Amrane, A., 2011. Removal of antibiotics by an integrated process coupling photocatalysis and biological treatment - Case of tetracycline and tylosin. Int. Biodeterior. Biodegrad. 65, 997-1003. Zhao, C., Pelaez, M., Duan, X., Deng, H., O'Shea, K., Fatta-Kassinos, D., Dionysiou, D.D., 2013. Role of pH on photolytic and photocatalytic degradation of antibiotic oxytetracycline in aqueous solution under visible/solar light: Kinetics and mechanism studies. App. Cat. B Environ. 134–135, 83-92. 109 6 Process Intensification at Near Neutral pH of a Homogeneous Photo-Fenton Reaction Using Ferricarboxylate Complexes: Application to Oxytetracycline Degradation This work demonstrates the application at near neutral pH of the photo-Fenton process mediated by ferricarboxylates on the treatment of aqueous solutions containing the antibiotic Oxytetracycline (OTC) under solar irradiation. The formation of a Fe:OTC complex after Fe2+ oxidation to Fe3+, in the presence of H2O2, showed the inconvenience of using the conventional Fe2+/H2O2/UV-Vis process at near neutral pH levels, as the Fe:OTC complex is retained in the filter. To overcome this, a Fe3+/Oxalate or Fe3+/Citrate/H2O2/UV-Vis process was proposed. Process efficiency was evaluated for different variables such as Fe3+ concentration, pH, temperature and irradiance, using a compound parabolic collector (CPC) photoreactor at lab-scale under simulated solar radiation. Reaction rates were compared in the presence of different inorganic anions and humic acids, and in two different real wastewater matrixes. This Chapter is based on the research article “João H.O.S. Pereira, Daniel B. Queirós, Ana C. Reis, Olga C. Nunes , Maria T. Borges, Rui A. R. Boaventura, Vítor J. P. Vilar. Process Enhancement at Near Neutral pH of a Homogeneous Photo-Fenton Reaction Using Ferricarboxylate Complexes: Application to Oxytetracycline Degradation. Chem Eng J (2014) 253, 217-228”. Chapter 6 110 Chapter 6 111 6.1 Introduction Antibiotics are a special group of pharmaceuticals used to control infection diseases in human and veterinary medicine. Residual concentration have been detected in various environmental compartments worldwide due to the fact that a large percentage of the consumed antibiotics are not completely metabolized (and thus are excreted as active substances) and that conventional wastewater treatment methods fail to completely remove them from solution (Kümmerer, 2009; Loos et al., 2009; Watkinson et al., 2009; Verlicchi et al., 2010). Even though direct cause and effect relationships are still to be established, it is widely recognized that antibiotic pollution contibutes to antibiotic resistance dissemination (Escher et al., 2011; Tamminen et al., 2011; Manaia et al., 2012). Thus, research in recent years has been focused on alternative ways to prevent the contamination of water supplies by this kind of pollutants. Urban wastewater treatment plants (WWTP) are one of its major sources (Michael et al., 2013), but highly-contaminated effluents resulting from fish farming activities may also be a special case to be considered (Rigos et al., 2006). Advanced oxidation processes (AOPs), a class of treatments involving different ways of generating the highly reactive and non-selective hydroxyl radical (OH) and other reactive oxygen species (Gogate and Pandit, 2004a; b), have been considered for the removal of these substances, especially those involving catalysis and solar irradiation such as the photo-Fenton process (Malato et al., 2009; Homem and Santos, 2011). Photo-Fenton comprises the combination of ferrous iron (Fe2+) with hydrogen peroxide (H2O2) and (solar) UV-Vis radiation resulting in the production of two moles of OH per mole of hydrogen peroxide (Eq. 6.01 and 6.02), as simplified by Gogate and Pandit (2004b):   HOOHFeOHFe 3 22 2 (6.01)   OHFehOHFe   2 2 (6.02) Pignatello et al. (2006) summarizes the reasons for the optimum operational pH value of the (photo-) Fenton process around 3 as follows: first, the solubility of Fe3+-hydroxy complexes decreases for pH values above 3; second, [Fe(OH)]2+, the most photoactive species (with absorption bands between 290 and 400 nm), reaches its maximum molar fraction around the aforementioned pH. As a consequence, there is a limited viability of applying this process in industrial scale due to the costs associated with pH corrections (initial acidification and final neutralization). Chapter 6 118 mg L-1), with an iron/oxalate molar ratio of 1:3, without initial pH adjustment. The degradation profiles of OTC were very similar for all tested Fe (III) concentrations and the calculated pseudo-first order kinetic rate constants are presented in Table 6.1. Figure 6.3. Effect of Fe (III) concentration (▼ – 1.0 mg L-1; ▲ – 2.0 mg L-1; ■ – 5.0 mg L-1) on the degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by ferrioxalate (1:3 iron/oxalate molar ratio). Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m-2, initial pH unadjusted and total added H2O2 = 90 mg L-1 A clear increase of reaction rates occurs with increasing iron amounts, but even with the lowest concentration tested, 1 mg L-1, OTC was no longer detected (LOD = 0.10 mg L-1) after 10 min of reaction (~0.6 kJUV L-1). By the end of the photo-treatment period (45 min, 2.75 kJUV L-1), 80% mineralization was achieved with both 2 and 5 mg L-1 of Fe (III), although a higher H2O2 consumption occurred with the highest Fe (III) concentration. In contrast, for 1 mg L-1 of Fe (III) only ~50% mineralization was achieved, with less H2O2 consumed than for the other Fe (III) concentrations. A direct relationship between the increases in hydroxyl radical formation and the initial Fe dosage is expected according to Eq. 6.01. The avoidance of the need of an iron removal step at the end of the treatment, associated with a satisfying mineralization degree, has shown 2 mg L-1 to be the concentration that represented the working option for photo-Fenton treatment with Fe (III)-oxalate complex. The pseudo-first order kinetic rate constant of the 2 mg L-1 experiment, 8.6 ± 0.5 L kJ-1, was the double of the rate constant reported in our previous work regarding the solar photocatalytic degradation of OTC in the presence of 0.5 g L-1 TiO2 at a neutral pH, carried out with the same experimental installation, 4.3 ± 0.4 L kJ-1 (Pereira et al., 2013). 0.0 0.2 0.4 0.6 0.8 1.0 0.0 0.5 1.0 1.5 2.0 010 20 30 40 50 0.0 1.0 2.0 3.0 4.0 5.0 010 20 30 40 50 0 4 8 12 16 OTC (C/C0) RAD-ON Consumed H2O2 (mM) RAD-ON Time (min) Total Dissolved Iron (mg Fe L-1) Time (min) DOC (mg C L-1) RAD-ON 010 20 30 40 50 3.0 4.0 5.0 6.0 7.0 pH Time (min) Chapter 6 119 Table 6.1. Pseudo-first-order kinetic parameters for the Fe3+/Oxalate/H2O2/UV-Vis process on the degradation of OTC (C0 = 20 mg L-1). Iron/oxalate molar ratio: 1:3. Overall conditions: total added H2O2 = 90 ppm; T = 25 °C; I = 44 WUV m-2. Influence of iron concentration (unadjusted pH) Experiment k (L kJ-1)a R2 SR2 (mg2 L-2) [Fe3+] = 1 mg L-1 7.6 ± 0.4 0.993 0.207 [Fe3+] = 2 mg L-1 8.6 ± 0.5 0.988 2.111 [Fe3+] = 5 mg L-1 15.8 ± 0.8 0.994 1.076 Influence of initial pH ([Fe3+] = 2 mg L-1) Experiment k (L kJ-1)a R2 SR2 (mg2 L-2) pH0 ~ 4.0 8.6 ± 0.5 0.988 2.111 pH0 = 5.0 6.3 ± 0.2 0.990 0.065 pH0 = 6.0 2.3 ± 0.1 0.990 0.146 Influence of Temperature and Irradiance ([Fe3+] = 2 mg L-1, pH0 = 5.0) Experiment k (L kJ-1)a R2 SR2 (mg2 L-2) T = 12 °C 3.4 ± 0.2 0.989 0.107 T = 35 °C 15.1 ± 0.6 0.998 0.012 I = 37 WUV m-2 7.5 ± 0.4 0.994 0.111 I = 24 WUV m-2 18 ± 2 0.987 0.513 Influence of Inorganic Ions and Humic Acids ([Fe3+] = 2 mg L-1, pH0 = 5.0) Experiment k (L kJ-1)a R2 SR2 (mg2 L-2) + 1.0 g L-1 Cl5.8 ± 0.4 0.971 0.823 + 1.0 g L-1 SO429.3 ± 0.9 0.974 0.541 + 0.1 g L-1 HCO37.9 ± 0.5 0.997 0.126 + 1.0 g L-1 NO37.2 ± 0.3 0.999 0.083 + 5.0 mg C L-1 HA 5.3 ± 0.2 0.998 0.066 Influence of the matrix ([Fe3+] = 2 mg L-1) Experiment k (L kJ-1)a R2 SR2 (mg2 L-2) WW effluent pH0 = 4.0 3.7 ± 0.3 0.991 0.417 pH0 = 5.0 1.5 ± 0.2 0.964 0.799 TF effluent pH0 = 4.0 9.59 ± 0.04 0.999 0.001 pH0 = 5.0 5.6 ± 0.4 0.997 0.277 a – pseudo-first order kinetic rate constant 6.3.2.2 Influence of initial solution pH After the previous experiments, the effect of the initial solution pH value was studied in order to find the highest possible working pH without compromising the reaction efficiency. The results, presented in Figure 6.4a, show that with unadjusted initial pH (~4.1), pH = 5.0 and pH = 6.0, OTC was no longer detected after 5, 10 and 15 min of illumination time (0.3, 0.6 and 0.9 kJ L-1 of accumulated UV energy, respectively). By the same order, DOC concentration was reduced Chapter 6 120 by 75, 51 and 37% of its initial value after 45 min of reaction (2.78 kJUV L-1). Despite the decreasing mineralization efficiency with increasing pH, there was not much difference between H2O2 consumption profiles. First of all, it must be noted that after the oxalate and iron addition steps and consequent pH adjustment, the initial concentration of OTC still decreases with increasing pH. Examining the pH range between 4 and 6 in the attempted iron speciation diagram (Figure 6.2b, right), the sum of the fractions of oxalate and OTC iron complexes remain somewhat constant, while the formation of an iron solid phase (Ferrihydrite, Fe(OH)3 (s)) is expected to form and rapidly increase after pH 5. Given the variation in both Fe(C2O4)2+ and Fe(C2O4)33fractions, as well as in 1:2 and 1:3 Fe-OTC complexes, the observed phenomenon could be attributed either to a diminishing strength in complexing iron by Fe(C2O4)33in favor of the 1:3 Fe-OTC complex, or to the slow formation of amorphous Fe(OH)3, confirmed by the lower initial dissolved iron concentration measured in the experiments at pH 5.0 and 6.0 (Figure 6.4a). As soon as irradiation started, a competition phenomenon between both ligands (oxalate and OTC) to form complexes with iron occurs, as seen by the erratic behaviour of DOC. A comparison with unfiltered DOC samples (data not shown) showed that in some sample points DOC is again retained in the filter. After hydroxyl radical attack of the original antibiotic molecule, some of the degradation products seems to retain iron (III)-chelating properties, whereby being retained in the filter at higher pH levels. However, after very low OTC concentrations, more iron is able to be chelated by oxalate, as a small increase in total dissolved iron concentration shows at 15 min. After this period, filtered and unfiltered DOC was found to be the same, so proper mineralization corresponds well with the marked increase in H2O2 consumption and possibly with the decarboxylation of ferrioxalate complexes, since iron levels begin to drop as well. Chapter 6 121 a) b) Figure 6.4. a) Effect of initial pH (● – pH0 ~ 4.0; ■ - pH0 = 5.0; ▼ - pH0 = 6.0) on the degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1.b) Speciation diagram for iron (III) as a function of pH in a solution containing 1.07 × 10-1 mM (9.5 mg L-1) oxalic acid and 3.58 × 10-2 mM (2 mg L-1) of Fe (III) without accounting (left) or accounting (right) for 10 mM (1 g L-1) SO42-. Ionic strength = 4 mM (left), Ionic strength = 30 mM (right). The speciation software MINEQL+ was used to calculate the data. Furthermore, to counter the lower amount of total dissolved iron along the pH = 6.0 experiment, an extra dose of oxalic acid was added after this same point. Although indeed iron concentration increased, it quickly began to drop, with no considerable increase in the mineralization rate. Regarding the influence of pH in the photochemistry itself, the role of the speciation of the main Fe (III)-oxalate complexes must be underlined. Faust and Zepp (1993) reported values of 1.0 and 0.6 at 436 nm for the two most photo-active Fe (III)-oxalate species, Fe(C2O4)2and Fe(C2O4)33-, respectively. Recently, Weller et al. (2013) also determined individual quantum yields for Fe(C2O4)2as 1.17 (366 nm) and 1.40 (436 nm), and for Fe(C2O4)33as 0.91 (366 nm) and 1.00 (436 nm). A speciation diagram discarding the Fe-OTC species, previously included for expository purposes only, has been calculated in the same manner (Figure 6.4b, left). Cumulatively, Fe(C2O4)2and Fe(C2O4)33are the main iron species in the studied pH range, albeit with changing proportions over pH. It can be observed that the fraction of Fe(C2O4)2lowers smoothly from 0.7 down to less than 0.1, whereas the fraction of Fe(C2O4)33evolves 0.0 0.2 0.4 0.6 0.8 1.0 OTC (C/C0) RAD-ON 0.0 0.5 1.0 1.5 2.0 Consumed H2O2 (mM) RAD-ON 010 20 30 40 50 0 4 8 12 16 DOC (mg C L-1) Time (min) +OA +OA RAD-ON 010 20 30 40 50 0.0 0.5 1.0 1.5 2.0 Total Dissolved Iron (mg Fe L-1) Time (min) RAD-ON 010 20 30 40 3.0 4.0 5.0 6.0 7.0 pH Time (min) 2 3 4 5 6 7 8 0.0 0.2 0.4 0.6 0.8 1.0 FeHC2O2+ 4 FeC2O+ 4 Fe(C2O4)- 2 Fe(C2O4)33 Fe(OH)3 (s) Molar fraction pH 2 3 4 5 6 7 8 0.0 0.2 0.4 0.6 0.8 1.0 Fe3+ Fe(OH)+ 2 FeSO+ 4 FeHC2O2+ 4 FeC2O+ 4 Fe(C2O4)- 2 Fe(C2O4)33 Fe(OH)3 (s) Molar fraction pH Chapter 6 122 from near 0.3, to a maximum of approximately 0.5 at pH = 5.0, and finally gets to less than 0.1 at pH = 6.0. Hence, by pH = 6.0, total iron precipitation would be expected. However, this precipitation occurs slowly, and Fe (III) may not be in equilibrium with the solid phase (Balmer and Sulzberger, 1999), so some oxalate is still expected to be found in solution, which may help to explain the decrease in the initial dissolved iron concentration shown at the maximum pH level tested. The decreasing reaction rate constants at pH ~ 4.1, pH = 5.0 and pH = 6.0 experiments, 8.6 ± 0.5, 6.3 ± 0.2 and 2.3 ± 0.1 L-1 kJ, respectively, appear to reflect this relationship between the molar fractions of the referred iron-oxalate complexes and their respective quantum yields (Table 6.1). Considering these results, the working pH of 5.0 was chosen for the subsequent set of experiments, since no extra dosage of oxalic acid was needed to obtain a satisfactory 49 % mineralization of the initial DOC content, OTC was no longer detected after 10 min of illumination time (QUV = 0.6 kJ L-1), and the final pH value of the solution approached the legal lower discharge pH value of 6.0 (Decree-Law n.º 236/98). Before proceeding, two more experiments were performed: Fe3+/Citrate/H2O2/UV-Vis system with 2 mg L-1 of Fe (III) and 1:1 iron/citrate molar ratio to compare the results with the previous unadjusted and adjusted initial pH = 5.0 Fe3+/Oxalate/H2O2/UV-Vis reactions. It can be seen in Figure 6.5a that for both initial pH values, the concentration of OTC after the citrate and ferric iron addition steps sharply decreases after only a minute of illumination time (QUV = 0.07 kJ L-1). The residual amount (~12 % of [OTC]0 in both cases) is slowly degraded until OTC is no longer detected after 5.0 and 7.5 min of reaction (0.3 and 0.4 kJUV L-1) for unadjusted pH0 and controlled pH0 = 5.0, respectively. In the same order, the respective DOC content was reduced to 31 and 54% of its initial value by the end of the illumination periods, while an average of 1.49 mM of H2O2 was consumed. The Fe3+/Citrate/H2O2/UV-Vis system presented remarkably higher reaction rate constants with unadjusted and adjusted initial pH = 5.0, k = 23 ± 3 L kJ-1 and k = 23 ± 4 L kJ-1, respectively. The speciation diagrams obtained with the MINEQL® software (Figure 6.5b) present significantly higher fractions of Fe (III) chelated with Citrate in the pH range of interest, compared with Oxalate. However, after the very quick initial OTC decay, total antibiotic removal seems to be hindered by the competitive effect of the progressive formation of not only intermediary byproducts resulting from OTC oxidation, but also from intermediates resulting from the ferricitrate decarboxylation (Rodríguez et al., 2009). Chapter 6 123 a) b) Figure 6.5. a) Effect of initial pH (■ - pH0 ~ 3.6, ● – pH0 = 5.0) on the degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:1 iron/citrate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1. b) Speciation diagram for iron (III) as a function of pH in a solution containing 3.58×10-2 mM (6.8 mg L-1) citric acid and 3.58×10-2 mM (2 mg L-1) of Fe (III), accounting (left) or not accounting (right) with 20 mg L-1 of OTC. Ionic strength = 4 mM. The speciation software MINEQL+ was used to calculate the data. Feng et al. (2012) reported good Tetracycline (0-40 μM) photodegradation results at near neutral pH values using Fe (III)-citrate with substantially higher iron/citrate molar ratios (from 1:10 up to 1:30). Nevertheless, this would imply a significant addition of extra dissolved carbon, in the order of dozens of mg C L-1. As a result, no further experiments with the Fe3+/Citrate/H2O2/UV-Vis system were considered. 6.3.2.3 Influence of temperature and irradiance When using photocatalytic systems under realistic conditions (solar pilot plants, for instance), special consideration must be given to two important process parameters, temperature and irradiance (Rodríguez et al., 2009; Zapata et al., 2009). There is a natural variability in solar irradiance depending on time of day, season of the year, atmospheric conditions, and even latitude, while the process usually takes place at ambient temperature (thus also depending on some of the abovementioned factors). In this way, further Fe3+/Oxalate/H2O2/UV-Vis 0.0 0.2 0.4 0.6 0.8 1.0 OTC (C/C0) RAD-ON 0.0 0.5 1.0 1.5 2.0 Consumed H2O2 (mM) RAD-ON 010 20 30 40 50 0 4 8 12 16 DOC (mg C L-1) Time (min) RAD-ON 010 20 30 40 50 0.0 0.5 1.0 1.5 2.0 Total dissolved iron (mg Fe L-1) Time (min) RAD-ON 010 20 30 40 50 3.0 4.0 5.0 6.0 7.0 pH Time (min) 2 3 4 5 6 7 8 9 0.0 0.2 0.4 0.6 0.8 1.0 Fe3+ FeOH2+ Fe(OH)+ 2 FeHCIT FeCIT FeOHCIT Fe(OTC)2 Fe(OTC)3 Fe(OH)3 (s) Molar fraction pH 2 3 4 5 6 7 8 9 0.0 0.2 0.4 0.6 0.8 1.0 Fe3+ FeOH2+ Fe(OH)+ 2 FeHCIT FeCIT FeOHCIT Fe(OH)3 (s) Molar fraction pH Chapter 6 124 experiments were performed under the chosen iron concentration, 2 mg L-1 of Fe (III), an iron/oxalate molar ratio of 1:3, and initial pH adjusted to 5.0, to screen the effects of different temperature and irradiance values on the decontamination of OTC in the lab-scale photoreactor. Two typical water temperature values commonly achieved in summer and winter conditions in our CPC pilot-plant installation were tested, 12 and 35°C (Figure 6.6), and compared to the default set temperature (25 °C). The lower reaction rate constant achieved with 12 °C (3.4 ± 0.2 L kJ-1) and the higher with 35 °C (15.1 ± 0.6 L kJ-1), while OTC was no longer quantifiable after 10 and 3 min, respectively, reflect a positive role of temperature in the photoFenton process. Figure 6.6. Degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: pH0 = 5.0 and total added H2O2 = 90 mg L-1. Effect of: a) Initial temperature (● – T = 12 °C; ■ - T = 25 °C; ▲ – T = 35 °C), I = 44 WUV m-2. The quantum yield of ferrioxalate is independent of temperature (Nicodem and Aquilera, 1983), but higher temperatures promote the regeneration rate of ferrous iron from ferric iron, thereby increasing hydroxyl radical production (Pignatello et al., 2006), which, in this case, seems to be analogous to a likely faster consumption of oxalic acid (there is a steeper reduction in the DOC content) reflected in the earlier and more accentuated decrease in total dissolved iron concentration. Higher temperatures were not tested, because they may promote an inefficient H2O2 decomposition into H2O and O2 (Monteagudo et al., 2013). Lower irradiances, 24 and 37 WUV m-2, on the other hand, did not result in major differences compared to the default value (I = 44 WUV m-2), regarding the required illumination time to fully degrade OTC, nor in H2O2 consumption rates (Figure 6.7). 0.0 0.2 0.4 0.6 0.8 1.0 OTC (C/C0) RAD-ON 0.0 0.5 1.0 1.5 2.0 Consumed H2O2 (mM) RAD-ON 010 20 30 40 50 0 4 8 12 16 DOC (mg C L-1) Time (min) +OA RAD-ON 010 20 30 40 50 0.0 0.5 1.0 1.5 2.0 2.5 Total Dissolved Iron (mg Fe L-1) Time (min) +OA RAD-ON 010 20 30 40 50 3.0 4.0 5.0 6.0 7.0 pH Time (min) Chapter 6 125 Figure 6.7. Degradation of OTC (C0 = 20 mg L-1) using solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Follow-up of OTC degradation, DOC removal, H2O2 consumption, total dissolved iron and pH. Process parameters: pH0 = 5.0 and total added H2O2 = 90 mg L-1. Effect of radiation intensity (♦ - I = 24.7 WUV m-2; ▲ – I = 37 WUV m-2; ■ - I = 44 WUV m-2), T = 25 °C. Taking these results into consideration, the environmental/natural variation of these two parameters do not substantially affects the process efficiency, since the illumination time required for antibiotic disappearance and the achieved mineralization levels were otherwise similar under the conditions assayed. 6.3.2.4 Influence of inorganic anions and humic acids Keeping in mind the perspective of applying the proposed process to treat concentrated micropollutants in retentates originating from membrane processes operation, the retention of other constituents which may affect the photo-Fenton process also needs to be considered (Martínez et al., 2013). Inorganic anions, for instance, can concentrate up to the order of the hundreds of milligrams (Miralles-Cuevas et al., 2013), and may affect the process by i) complex formation with Fe (II) and Fe (III), ii) hydroxyl radicals scavenging and formation of less reactive inorganic radicals and iii) oxidation reactions involving these inorganic radicals (De Laat et al., 2004). The presence of dissolved natural organic matter, such as humic acids (HA), may also hamper the removal process since it may bind the micropollutants themselves (Gu et al., 2007), attenuate incoming radiation (Bautitz and Nogueira, 2007), act as a hydroxyl radical sink and finally, although their binding with iron may increase its solubility, it alters its redox cycle and changes the formation rate of hydroxyl radicals (Lindsey and Tarr, 2000). 0.0 0.2 0.4 0.6 0.8 1.0 OTC (C/C0) RAD-ON 0.0 0.5 1.0 1.5 2.0 Consumed H2O2 (mM) RAD-ON 010 20 30 40 50 0 4 8 12 16 +OA DOC (mg C L-1) Time (min) RAD-ON 010 20 30 40 50 0.0 0.5 1.0 1.5 2.0 +OA Total Dissolved Iron (mg Fe L-1) Time (min) RAD-ON 010 20 30 40 50 3.0 4.0 5.0 6.0 7.0 pH Time (min) Chapter 6 126 For these reasons, a set of experiments was performed with 1 g L-1 of Cl-, SO42-, NO3-, 0.1 g L-1 of HCO3and 5 mg C L-1 of HA to assess the individual effect of each interfering species in the degradation of OTC under the chosen conditions (Figure 6.8). The required illumination time to bring the concentration of OTC under LOD levels only marginally increased in all experiments, except in the case of HA, which was longest. Likewise, HA was the only interfering species whose effect on the mineralization process was noteworthy, since initial DOC was only reduced by 35% against an average of 41% reduction compared to the other interfering species, and 46% in their absence, after 40 min of illumination time (QUV = 2.3 kJ L-1). The overall consumption profiles of H2O2 reflected well these differences in mineralization degrees. Figure 6.8. Follow-up of OTC degradation (C0 = 20 mg L-1), DOC removal, H2O2 consumption, total dissolved iron and pH in the absence (■) and in the presence of 1 g L-1 of Cl- (●), SO42- (), NO3- (♦), 0.1 g L-1 of HCO3- (▲), 5 mg C L-1 of HA (×) using the solar photo-Fenton process mediated by 2 mg L-1 iron (III) and a 1:3 iron/oxalate molar ratio. Process parameters: pH0 = 5.0, T = 25 °C, I = 44 WUV m-2 and total added H2O2 = 90 mg L-1. This apparent weak impact in the reaction efficiency in the presence of iron-complexing species such as Clor SO42could be also explained recurring to the study of the speciation of iron in solution via the MINEQL+ software. The equilibrium constants of ferrioxalate complexes are much higher than those of both Fe-Cl and Fe-SO4 complexes (Truong et al., 2004), and thus it can be seen in Figure 6.4b that, for instance, the introduction of 1 g L-1 of SO42when plotting the iron complexes in solution (with the respective ionic strength correction, 30 mM) only alters the distribution of the Fe(C2O4)2complex in favor of Fe(C2O4)33-, whereas none of the Fe-SO4 complexes is shown to be formed at the pH range of interest. 0.0 0.2 0.4 0.6 0.8 1.0 OTC (C/C0) RAD-ON 0.0 0.5 1.0 1.5 2.0 Consumed H2O2 (mM) RAD-ON 010 20 30 40 0.0 0.3 0.6 0.9 1.2 1.5 DOC (DOC/DOC0) Time (min) RAD-ON 010 20 30 40 0.0 0.5 1.0 1.5 2.0 Total Dissolved Iron (mg Fe L-1) Time (min) RAD-ON 010 20 30 40 3.0 4.0 5.0 6.0 7.0 8.0 pH Time (min) Chapter 6 127 6.3.2.5 Influence of the matrix After assessing the potential influence of individual common wastewater components on the Fe3+/Oxalate/H2O2/UV-Vis process, two different effluents were spiked with OTC (C0 = 20 mg L-1) to complementarily study the influence of real matrixes. Even though pH0 = 5.0 was set beforehand as the chosen pH to study the process efficiency, experiments were also performed with pH0= 4.0 in each media. This was due to the fact that after the oxalate and iron (III) addition steps, the pH of the TF effluent lowered to the same unadjusted pH of the DW solution, around 4.0. On the other hand, given composition of WW medium (Table 6.2), additional pH adjustment with acid was required, but it also helped to lower some of the initial inorganic carbon. Table 6.2. Main characteristics of the tested effluents, before the OTC-spike step. Effluent pH DOC / IC Average anion concentrations (mg L-1) (mg L-1) [PO43-] [Cl-] [SO42-] WW 6.46 5.5 / 5.1 13.5 61.5 42.9 TF 6.70 3.9 / 1.8 0.95 The results comparing the efficiency of the process carried out in deionized water (DW), domestic wastewater (WW) and trout farm (TF) effluents are thus presented in Figure 6.9a. With pH0 = 4.0, OTC was no longer detected after 10 min of illumination in WW effluent (QUV = 0.7 kJ L-1), and after 6 min in TF effluent (QUV = 0.4 kJ L-1). Required illumination times were higher in pH0 = 5.0: 20 min in WW effluent (QUV = 1.4 kJ L1), and 10 min in TF effluent (QUV = 0.7 kJ L-1). The distribution of the iron (III) species in both media (Figure 6.9b) seems to explain the similar kinetic degradation rates achieved with TF effluent, and the worse with WW (Table 6.1). At pH0 = 4.0, some fraction of iron (III) in the form of the mineral Strengite (FePO4.2H2O (s)) is already expected to be found in WW, while in TF medium results, there is a relatively higher proportion of the more photo-active Fe(C2O4)2complex compared to that in DW (Figure 6.4b). The proportion of FePO4.2H2O (s) at pH0 = 5.0 in both effluents clearly reflects the lower OTC kinetic degradation rates, especially in the case of WW effluent. Also to be considered is the abovementioned effect of the additional DOC/NOM content of both effluents in the overall process efficiency. Compared to the DW experiments, for instance, less H2O2 was consumed by the end of the photo-treatment period (45 min of illumination, QUV = 3.3 kJ L-1), which may also reflect the lesser reduction in DOC content comparing with DW experiments at both tested pH values. Accounting for each effluent’s initial DOC values, DOC reduction in WW effluent was of 39 and 31% for pH0 = 4.0 and 5.0, respectively, whereas in TF effluent it was 56 and 49%, in the same order.