Effect of water matrix on Vacuum UV process for the removal of organic micropollutants in surface water

Transcription

1 Effect of water matrix on Vacuum UV process for the removal of organic micropollutants in surface water by Clara Duca M.A.Sc., University of Torino, Italy 2009 B.Sc., University of Torino, Italy 2007 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF Doctor of Philosophy in THE FACULTY OF GRADUATE AND POSTDOCTORAL STUDIES (Chemical and Biological Engineering) The University of British Columbia (Vancouver) February 2015 c Clara Duca, 2015

2 Abstract UV-based advanced oxidation processes (UV-AOPs) have been demonstrated as effective technologies for the removal of micropollutants in water. One promising UV-AOP is Vacuum UV (VUV), which relies on the formation of hydroxyl radicals (HO ) by the photolysis of water induced by VUV photons. TiO2 /UV photocatalysis is another promising AOP. Both VUV and UV photocatalysis are greatly affected by the water matrix, inorganic ions and natural organic matter (NOM). Water constituents can absorb the UV and VUV radiations, they can act as HO radicals scavengers, and can produce radicals when photolyzed. The main objective of this research was to study the effects of water matrix on the efficiency of VUV for the degradation of micropollutants (with atrazine as a model contaminant). First, the absorbance of radiation at 254 nm and 185 nm was measured in the presence of different ions and NOM. All the inorganic ions showed a molar absorption coefficient equal to zero at 254 nm except nitrate with a ε= 3.51 M 1 cm 1. On the other hand, at 185 nm all the ions absorbed 185 nm radiation, with chloride showing the highest absorption coefficient ε=2791 M 1 cm 1. NOM showed a high absorption coefficient at both 254 and 185 nm ranging from 116 to 638 M 1 cm 1 at 254 and from 1137 to 1537 M 1 cm 1 at 185 nm). Second, the HO scavenging effects of different components were evaluated; nitrate showed a detrimental effect both with UV/H2 O2 and with VUV. The presence of 50 ppm of bicarbonate reduced the degradation rate of atrazine considerably. Sulfate seemed to photolyze at 185 nm to form HO. NOM was found to be a strong HO scavenger: in the presence of 9 ppm (DOC) NOM, less than 1% of the HO radicals were available to react with atrazine. The next component of this work involved developing a method for the measurement of quantum yield of atrazine at 185 nm. This allowed to measure the degradation of atrazine due to photolysis only. Finally, this research investigated the combination of VUV with TiO2 /UV. The results showed that incorporating photocatalysis cannot improve significantly the efficacy of VUV. ii

3 Preface Several manuscripts in journals and in conference proceedings have been published or are under consideration for publication pertaining to the results of this thesis. The following is a list of such manuscripts which have been published, presented in conferences or as posters. Journals: Synthesis, characterization, and comparison of Sol Gel TiO2 immobilized photocatalysts. Clara Duca, Gustavo Imoberdorf, and Madjid Mohnseni. Int. J. Chem. react. Eng., 2013, Volume 11, Issue 2, Pages Novel Collimated Beam Setup to Study the Kinetics of VUV-Induced Reactions. Clara Duca, Gustavo Imoberdorf, and Madjid Mohnseni. Photochem. Photobiol., 2014, Volume 90, Issue 1, Pages In addition chapter 6,7 and 8 are currently under preparation. Conference proceedings: Composite titania photocatalytist coating for Vacuum UV photoreactor. Clara Duca, Gustavo Imoberdorf, and Madjid Mohnseni. 2nd North American Conference on Ozone, UltraViolet and advanced Oxidation technologies Toronto, On September 19-20, Degradation of Micropollutants using Vacuum UV (VUV) Advanced Oxidation. Gustavo Imoberdorf, Clara Duca, and Madjid Mohseni. The17th International Conference on Advanced Oxidation Technologies for Treatment of Water, Air and Soil (AOTs-17) San Diego, Ca November Simultaneous inactivation of microorganisms and removal of micropollutants in VUV reactors. Laith Furatian, Clara Duca, Mehdi Bagheri, Gustavo Imoberdorf, and Madjid Mohseni BC water and waste water association annual conference and trade show Penticton, BC April iii

4 Effects of Inorganic Ions and NOM on the Degradation of Micropollutants during UV-H2 O2 AOP. Clara Duca, Gustavo Imoberdorf, and Madjid Mohseni. IUVA Moving forward: sustainable UV solutions to meet evolving regulatory challenges Washington DC, August 12-14, Study of the Kinetics of VUV-induced Degradation of Micropollutants. Clara Duca, Gustavo Imoberdorf, Madjid Mohseni 62nd Canadian Chemical Engineering Conference Vancouver BC October 14-17, Effect of NOM and Inorganic Ions on the Degradation of Micropollutants with VacuumUV Radiation. Clara Duca, Gustavo Imoberdorf, Madjid Mohseni 15th Canadian national conference and 6th policy forum on drinking water Kelowna BC October New methodology for the quantification of the quantum yield of micropollutants at 185 nm. Clara Duca, Gustavo Imoberdorf, Madjid Mohseni. IUVA World congress and exhibition. Las Vegas, NV September Treating fracturing flowback by biological and advanced oxidation processes. Yaal Lester, Tesfayohanes Yacob, Clara Duca, Kurby Sitterley, Julie Korak, James Rosenblum, and Karl Linden. 247th ACS meeting and exposition. Dallas TX March The last work was produced during my time I have spent working in the University of Boulder, Colorado. During that time I have worked on the flowback treatment and characterization. Poster presentations: Advanced oxidation using vacuum UV, catalyst development. Clara Duca, Madjid Mohseni (Part of UBC Advanced Oxidation and Small Systems Research Group Poster Presentation). The Consensus Conference on Small Water Systems Management for the Promotion of Indigenous Health. Victoria, BC, March 21-22, Composite photocatalytic coatings for vacuum-uv reactor. Clara Duca Res Eau Waternet: 3rd roundtable Vancouver BC March 12th, Effect of water matrix on UV based processes. Clara Duca, Gustavo Imoberdorf, Madjid Mohnseni. Big Value in Small Systems Vancouver BC, October 5-6, Behaviour of micropollutants and water matrix constituents exposed to 185 nm radiation. Clara Duca Res Eau Waternet roundtable Toronto ON November 3th, Degradation of micropollutants with Vacuum UV process (VUV). Clara Duca Res Eau Impact Vancouver BC October 2-3, iv

5 Influence of Water Matrix on UV based Processes for the Detoxification of Water Clara Duca Res Eau meeting Whistler BC May 28-29, v

6 Table of Contents Abstract ii Preface iii Table of Contents vi List of Tables x List of Figures xii Glossary xvii 1 2 Background Problem of micropollutants in drinking water Presence of natural organic matter (NOM) in raw surface water Conventional water treatments for drinking water Advanced oxidation processes (AOPs) Thesis layout Literature review Hydrogen peroxide/uv process Photocatalysis Vacuum UV Potential of combined VUV and UV/TiO2 for micropollutant degradation Effect of water matrix Effect of inorganic ions Effect of natural organic matter (NOM) Knowledge gaps Research objectives vi

7 4 Experimental setups and procedures Setups Photoreactors UV and VUV reactors VUV collimated beam set-up Experimental procedures Actinometry at 254 nm Actinometry at 185 nm Hydrogen peroxide measurement Absorption coefficient measurements at 185 and 254 nm ph and Dissolved oxygen (DO) measurements Reproducibility and data accuracy TiO2 photocatalyst development and evaluation of photocatalysis coupled with VUV Materials and methods Synthesis of the photocatalysts Physical characterization of the photocatalysts Photocatalytic activity assessment Results and discussion Physical characterization of the photocatalysts Photocatalytic efficacy Evaluation of the deactivation of the photocatalysts Evaluation of the adherence of the photocatalysts to the support Performance of the photocatalysts in a flow-through reactor Conclusions Effects of inorganics on the degradation of micropollutants with VUV advanced oxidation Material and methods Water samples and chemicals UV/H2 O2 and VUV irradiations Analytical methods Results and discussion Molar absorption coefficients at 185 nm and 254 nm Methodology for the determination of the observed kinetic constant Degradation of atrazine in millipore water Degradation of atrazine in the presence of different inorganic ions Effect of Na+ on VUV and H2 O2 /UV processes vii 75

8 Effect of nitrate on VUV and H2 O2 /UV processes Effect of bicarbonate on VUV and H2 O2 /UV processes Effect of sulfate on VUV and H2 O2 /UV processes Effect of different inorganic ions on the degradation of atrazine Conclusions Quantitative effect of natural organic matter on the efficacy of Vacuum UV oxidation of atrazine Materials and methods Water samples and chemicals UV/H2 O2 and VUV irradiations Analytical methods Results and discussion Molar absorption coefficients at 185 nm and 254 nm UV/H2 O2 irradiations VUV irradiations NOM sensitization effect during VUV treatment Conclusions The effect of chloride on Vacuum-UV and UV/H2 O2 photo-induced degradation of phenol Materials and methods Water samples and chemicals UV/H2 O2 and VUV irradiations Analytical methods Results and discussion Molar absorptions coefficients at 185 nm and 254 nm Degradation of phenol with UV/H2 O Degradation of phenol with VUV Effect of ph on the equilibrium of chloride Effect of real water matrix on the degradation of atrazine in real water samples119 Conclusions Conclusions Overall conclusions Significance of the research Recommendations for future research viii

9 Bibliography Appendix A Supplementary data for the photocatalysts deactivation tests and evaluation in a flow through reactor Appendix B Supplementary data for the UV/H2 O2 and VUV irradiation in the presence of different ions Appendix C Supplementary data for the UV/H2 O2 and VUV irradiation in the presence of different concentration an type of NOM Appendix D Supplementary data for the measurement of the absorbance of inorganic ions and NOM Appendix E Quantum yield of micropollutants Appendix F Degradation of micropollutants with ozone-generating Hg lamps (185 and 254 nm) ix

10 List of Tables Table 5.1 Percentage of the 4 polymorphous in the 5 photocatalysts Table 5.2 Particle size of the 5 different sols Table 6.1 Absorbtion coefficient of different compounds at 254 nm Table 6.2 Absorbtion coefficient of different compounds at 185 nm Table 6.3 Effect of different inorganic ions on the distribution of 254, 185 nm photons and on the HO radicals concentration Table 7.1 Absorption coefficient of various compounds at 254 nm and 185 nm Table 7.2 distribution of HO radicals and 254 nm photons during UV/H2 O2 with Nordic NOM Table 7.3 distribution of HO radicals and 254 nm photons during UV/H2 O2 with Suwannee NOM Table distribution of HO radicals and 185 nm photons during VUV process with Nordic NOM Table 7.5 distribution of HO radicals and 185 nm photons during VUV process with Suwannee NOM Table 8.1 Percentage of photons at 185 nm absorbed by water and by various concentrations of chloride ions Table 8.2 Effect of Bowen Island water matrix on the degradation of atrazine Table 8.3 Effect of peachland water matrix on the degradation of atrazine Table A.1 2,4-D conversion for photocatalysis/uv with millipore water and a flowrate of 1L/min Table A.2 2,4-D conversion for P25 slurry/uv with millipore water and a flowrate of 1L/min.141 Table A.3 2,4-D conversion for VUV process with millipore water and a flowrate of 1L/min. 142 Table A.4 2,4-D conversion for photocatalysis/uv with surface water and a flowrate of 0.25 L/min x

11 Table A.5 2,4-D conversion for VUV with surface water and a flowrate of 0.25 L/min Table D.1 Absorbance at 185 nm for different concentration of NaCl Table D.2 Absorbance at 185 nm for different concentration of NaHCO Table D.3 Absorbance at 185 nm for different concentration of NaNO Table D.4 Absorbance at 254 nm for different concentration of NaNO Table D.5 Absorbance at 185 nm for different concentration of Nordic NOM Table D.6 Absorbance at 254 nm for different concentration of Nordic NOM Table D.7 Absorbance at 185 nm for different concentration of HSO xi

12 List of Figures Figure 4.1 Batch experimental setup a) and differential reactor b). The batch reactor consisted of a sparging beaker, a pump, a flowmeter, the differential reactor and UV lamps Figure 4.2 Flow through reactor experimental setup; the reactor consisted of a annular reactor (a), a pump (b), a feed storage tank (c), and a treated water tank d).... Figure UV collimated beam set-up (a) and cross section of the UV collimated beam (b); the collimated beam consisted of a UV lamp, a reactor chamber and a stirrer. 34 Figure 4.4 Diagram of the collimated beam and of the top of the PVC enclosure. (1) motor, (2) reaction vessel, (3) stirrer (4) head of the enclosure, (5) enclosure, (6) VUV lamp, (7) Teflon cylinder (8) Orings, (9) quartz sleeve, (10) PVC head of the enclosure, (11) optical filter, and (12) Suprasil quartz Figure 4.5 Radial distribution of radiation on the surface of the reaction vessel Figure 4.6 Angular distribution of radiation on the surface of the reaction vessel Figure 5.1 SEM micrograph of photocatalyst D. The micrograph shows important fractures on the surface of the catalyst Figure 5.2 SEM micrograph of photocatalyst E. The micrograph shows an higher homogeneity compared to the one of catalyst D. Figure XRD of the five different coatings. The figure shows hat catalyst B is amorphous and catalyst A, B, D and E are a a mixture of rutile and anatase Figure 5.4 UV-VIS spectra of quartz, photocatalysts A, B and C Figure 5.5 Photocatalytic activity of the five different photocatalysts for the degradation of 0.1 ppm 2,4-D in millipore water. Error bars represent the standard deviations of three replicates samples Figure Attrition tests conducted with the photocatalyst D. The degradation of 0.1 ppm of 2,4 D in millipore water was followed. The curves show the photocatalytic activity of the fresh photocatalyst and after 24 h of H2 O recycling. Error bars represent the standard deviations of three replicates samples xii 53

13 Figure 5.7 Attrition tests conducted with the photocatalyst E. The degradation of 0.1 ppm of 2,4 D in millipore water was followed. The curves show the photocatalytic activity of the fresh photocatalyst and after 24 h of H2 O recycling.error bars represent the standard deviations of three replicates samples Figure Apparent first-order rate constants for the degradation of 2,4-D obtained with photocatalyst D and photocatalyst E after repeated photocatalytic experiments. Error bars represent the standard deviations of three replicates samples..... Figure ,4 D conversion for photocatalysis/uv, VUV and P25 slurry/uv with millipore water and a flowrate of 1 L/min Figure ,4 D conversion for photocatalysis/uv, and VUV with surface water and a flowrate of 0.25 L/min Figure ,4 D conversion for VUV with surface water and millipore water Figure 6.1 Water absorbance for different cell path length Figure 6.2 Apparent first order kinetic constant for the degradation of atrazine obtained with H2 O2 /UV and VUV.The fuence rates were 0.33mW/cm2 for 254 nm and 0.06 mw/cm2 for 185 nm Figure 6.3 Photolysis of atrazine with VUV and UV. The fuence rates were mW/cm2 and 0.29mW/cm2, respectively. Error bars represent the standard deviations of three replicates samples Figure First order rate constant for the degradation of atrazine using UV/H2 O2 in the with different concentrations of NaF. The fuence rate was 0.29mW/cm2. Error bars represent the standard deviations of three replicates samples Figure First order rate constant for the degradation of atrazine using VUV with different concentrations of NaF. The fuence rates was 0.03mW/cm2. Error bars represent the standard deviations of three replicates samples Figure First order constant for the degradation of atrazine using UV/H2 O2 in the presence of different concentration of nitrate. The fuence rate was 0.31mW/cm2. Error bars represent the standard deviations of three replicates samples..... Figure First order constant for the degradation of atrazine using UV in the presence of different concentration of nitrate. The fuence rate was 0.29mW/cm2. Error bars represent the standard deviations of three replicates samples Figure First order rate constant for the degradation of atrazine using VUV with different concentrations of nitrate. The fuence rate was 0.03mW/cm2. Error bars represent the standard deviations of three replicates samples xiii 80

14 Figure 6.9 First order constant for the degradation of atrazine using UV/H2 O2 in the presence of different concentration of bicarbonate. The fuence rate was 0.29mW/cm2. Error bars represent the standard deviations of three replicates samples Figure 6.10 First order constant for the degradation of atrazine using VUV in the presence of different concentration of bicarbonate. The fuence rate was 0.03mW/cm2. Error bars represent the standard deviations of three replicates samples Figure 6.11 First order constant for the degradation of atrazine using UV/H2 O2 in the presence of different concentration of sulfate. The fuence rate was 0.29mW/cm2. Error bars represent the standard deviations of three replicates samples Figure 6.12 First order constant for the degradation of atrazine using VUV in the presence of different concentration of sulfate. The fuence rate was 0.05mW/cm2 Error bars represent the standard deviations of three replicates samples Figure 7.1 Absorption spectra of 9 ppm of Suwannee river NOM and of Nordic NOM Figure 7.2 First order rate constant for the degradation of atrazine using UV/H2 O2 with various concentrations of NOM Suwannee river, and Nordic reservoir. The fluence rate was 0.29mW/cm2. Error bars represent the standard deviations of three replicates samples Figure First order rate constant for the degradation of atrazine using VUV with various concentrations of NOM (Suwannee river, and Nordic reservoir) and Methanol. The fluence rate was 0.03mW/cm2. Error bars represent the standard deviations of three replicates samples Figure First order rate constant for the degradation of phenol using UV/H2 O2 with various concentrations of chloride ion. The fuence rate was 0.25mW/cm2. Error bars represent the standard deviations of three replicates samples Figure 8.2 Schematic decomposition path of aqueous phenol [1] Figure 8.3 Phenol degradation (0.1 ppm) and byproducts formation during the UV/H2 O2 oxidation process. Error bars represent the standard deviations of three replicates samples Figure 8.4 Distribution of chlorinated by-products formed during UV/H2 O2 at various chloride concentrations, from 0 to 1.5 mm. Error bars represent the standard deviations of three replicates samples Figure 8.5 First order rate constant for the degradation of phenol using VUV with various concentrations of chloride ion. The fuence rate was 0.06mW/cm2. Error bars represent the standard deviations of three replicates samples Figure 8.6 Phenol degradation (0.1 ppm) and byproducts formation during the VUV process. Error bars represent the standard deviations of three replicates samples xiv

15 Figure 8.7 Distribution of chlorinated by-products formed during the VUV process at various chloride concentrations, from 0 to 1.5 mm Figure A.1 Degradation of 2,4-D in the presence of photocatalyst E after repeated photocatalytic experiments Figure A.2 Degradation of 2,4-D in the presence of photocatalyst D after repeated photocatalytic experiments Figure B.1 Degradation of atrazine in the presence of different concentration of NaF during UV/ceH2O2 treatment Figure B.2 Degradation of atrazine in the presence of different concentration of NaF during VUV treatment Figure B.3 Degradation of atrazine in the presence of different concentration of NaNO3 during UV treatment Figure B.4 Degradation of atrazine in the presence of different concentration of NaNO3 during UV/H2 O2 treatment Figure B.5 Degradation of atrazine in the presence of different concentration of NaNO3 during VUV treatment Figure B.6 Degradation of atrazine in the presence of different concentration of NaHCO3 during UV/H2 O2 treatment Figure B.7 Degradation of atrazine in the presence of different concentration of NaHCO3 during VUV treatment Figure B.8 Degradation of atrazine in the presence of different concentration of NaHSO3 during VUV treatment Figure B.9 Degradation of atrazine in the presence of different concentration of NaHSO3 during VUV treatment. In the experiments with NaHSO3 1 ppm of methanol was added in order to slow the degradation rate and have more stable measurements Figure B.10 Degradation of atrazine in the presence of different concentration of Cl during UV/H2 O2 treatment Figure B.11 Degradation of atrazine in the presence of different concentration of Cl during VUV treatment Figure C.1 Degradation of atrazine in the presence of different concentration of Nordic NOM during UV/H2 O2 treatment Figure C.2 Degradation of atrazine in the presence of different concentration of Nordic NOM during VUV treatment xv

16 Figure C.3 Degradation of atrazine in the presence of different concentration of Suwannee river NOM during UVH2 O2 treatment Figure C.4 Degradation of atrazine in the presence of different concentration of Suwannee river during VUV treatment Figure D.1 Absorbance of sodium chloride at 185 nm at different sodium chloride concentrations Figure D.2 Absorbance of sodium bicarbonate at 185 nm at different sodium bicarbonate concentrations Figure D.3 Absorbance of sodium nitrate at 185 nm at different sodium nitrate concentrations.156 Figure D.4 Absorbance of sodium nitrate at 185 nm at different sodium nitrate concentrations.156 Figure D.5 Absorbance of sodium sulfate at 185 nm at different sodium sulfate concentrations Figure D.6 Absorbance of Nordic NOM at 185 nm at different TOC concentrations Figure D.7 Absorbance of Nordic NOM at 185 nm at different TOC concentrations Figure D.8 Absorbance of Suwannee NOM at 185 nm at different TOC concentrations Figure D.9 Absorbance of Suwannee NOM at 185 nm at different TOC concentrations xvi

17 Glossary 2,4- D 2,4-Dichlorophenoxyacetic acid 4- CP 4-chlorophenol 4- NP 4-nitrophenol A Surface of the reaction vessel (cm2 ) C Concentration (M 1 ) CBZ Carbazepine DFC Diflonac DOC Dissolved organic carbon F Fluence mj cm 2 G Incident radiation mw cm 2 HPLC I High Performance Liquid Cromatography Specific spectral radiation intensity mwcm 2 s 1 Ionic Chromatography IC ICP / OES K Inductively Coupled Plasma Optical Emission Spectrometry Apparent kinetic constant M 1 s 1 LP Low pressure amalgam lamp MP Medium pressure mercury lamp MTBE MW Methyl-tert-butil ether Molecular weight xvii

18 N-nitrosodimethylamine NDMA NOM Natural Organic Matter SMX sulfamethoxazolone Specific ultraviolet absorbance SUVA THMFP Trihalomethane formation potential TOC Total organic carbon VUV Vacuum UV V Volume of the reaction vessel (L) Z Path lenght [Greek alphabets] α Absorbance coefficient of the propagating medium ε Molar absorption Coefficient ζ Fraction of photons absorbed φ Quantum yield [Superscripts and subscripts] 0 Initial condition nm nm AQ Aqueous ATZ Atrazine CON Consumed GEN Generated H2 O Water I i specie MEHO Methanol xviii

19 PHOT RV Photolysis Reaction vessel xix

20 Chapter 1 Background Most population, who lives in middle size and large urban centers, take drinking water for granted. However, a large number of small rural communities have difficulties to access drinkable water on daily basis, putting their health at risk for water born diseases. Given the remote settings, conventional water treatments are not suitable, since they require intense energy, high cost and experienced operators. For this reason, often, these communities rely on outdated treatment technologies. The result is that many communities live in a situation with frequent boil water advisories. Boiling water can solve the problem related to microbial contamination, but it cannot remove heavy metals, organics and taste and odor compounds. For these reasons, there is the need to develop effective and simple technologies that can be installed in remote locations and that can provide safe water, removing micropollutants and taste and odor compounds. 1.1 Problem of micropollutants in drinking water Micropollutants are contaminants which exist in trace amounts (ng/l, pg/l) but they pose potentially adverse health influences on humans. These emerging or new unregulated contaminants in water supplies have become an environmental problem. Micropollutants can have severals origins and they can cause various problems [2]. For example, they can be industrial products such as phthalates and polychlorinated biphenyls: in this case the problem associated with those compounds is the biomagnifications and the long range transportation [3]. These chemicals are slowly degraded 1

21 and therefore, they can be transported by water or air to locations far from their source. Micropollutants can also be products used in everyday life such as detergents, pharmaceuticals, hormones: these chemicals could be responsible for bacterial resistance or feminization of fish [4]. With micropollutants another class is biocides: some examples are atrazine and Dichlorvos which have been proven to bring toxic effects and create persistent metabolites [5]. In addition, micropollutants can have a geogenic source and they can be natural chemicals. Examples are heavy metals, cyanotoxins, human hormones: all these chemicals can cause drinking water quality problems due to their potentials negative effects on human health [6]. Another category of micropollutants are the disinfection byproducts such as trihalomethanes and haloacetic acid [7]. The source, behavior, and treatments of few macropollutants such as natural organic matter (NOM) and salts occurring at µg/l and mg/l is well understood [8]. However, it is far more difficult to assess the influence of the thousands of micropollutants that may be present at low concentrations. These chemicals have been found not only in industrialized areas, but also in more remote environments because of their low degradability. Because of their physico-chemical properties (high water solubility, often poor biodegradability, and low concentrations, usually in part per billion or trillion), most current wastewater treatment plants are not designed to treat micropollutants [9]. For this reason, so-called tertiary treatments such as carbon adsorption, and reverse osmosis are added to the system. These treatments, however, are expensive and not able to degrade the contaminants; they are able only to transfer the contaminants from one phase to another leading to waste generation and disposal issues. Hence, mitigation technologies to reduce the impacts of micropollutants require further development. The water supply industry is faced with a particularly difficult problem: low level (sub ppb to ppt level) of micropollutants must be inexpensively removed from large volumes of raw water. 1.2 Presence of natural organic matter (NOM) in raw surface water NOM consists of humic substances, humic and fulvic acid, and non-humic substances such as proteins, amino acids, sugars and polysaccharides. The chemical characteristics of NOM are influenced not only by the source materials, but also by the biogeochemical processes involved in carbon cy- 2

22 cling within the terrestrial and aquatic system. The heterogeneity of humic substances derives from their various molecular sizes and from their various chemical structures. The dissolved organic carbon (DOC) is an operational classification and it consists of compounds below 0.45 micrometers. In water it typically consists of 90% fulvic acid and 10% humic acid [10]. Aquatic humic substances have a moderate aromatic character, and contain primarily carboxyl groups, phenolic groups, alcohol and metoxyl groups, aldehydes and ketones. Characterization of the functional groups is important to understand their reactivity with oxidants and other species dissolved in the water. The presence of NOM in water could be troublesome for water treatment processes. NOM causes membrane fouling, absorbs ultraviolet (UV) irradiation, scavenges oxidants, among others. Further, NOM is a precursor of chlorination disinfection byproducts such as trihalomethanes and haloacetic acids [11] and may induce bacteria regrowth in the distribution system. Therefore, it is important to remove NOM from the source water before applying various treatment options and/or sending the water into the distribution system. Several processes have been used to remove NOM such as coagulation, which is the most well-established method of removing NOM. More recently some studies have been carried out on the coupling of various processes. For example, UV or Vacuum UV (VUV) radiation photooxidation followed by biological treatments has been an option under investigation, because pre-oxidation promotes the formation of biodegradable intermediates which can allow microbes to mineralize these oxidation products to CO2 [12]. 1.3 Conventional water treatments for drinking water The combination of coagulation, sedimentation, filtration and disinfection is used to provide clean and safe drinking water to the public [13] and is considered the so-called conventional water treatments that could reduce the concentration of NOM substantially. Coagulants, which may be aluminum or iron based, are chemicals that can be added to water to induce dissolved and colloidal species to agglomerate into larger particles known as flocs. These flocs are removed in a clarification step, which may be based on gravity or buoyancy. The effectiveness of coagulation to remove NOM depends primarily on the many variables of coagulation processes including the coagulant 3

23 type, ph, coagulant dose, flocculation time, and NOM characteristics. Other technologies used for the removal of NOM are membrane filtration, adsorption processes (granulated activated carbon), ion exchange resins, and advanced oxidations processes such as ozonation (O3 ), UV/H2 O2, UV/O3, and Vacuum UV radiation ( VUV) [14]. Conventional methods of water disinfection and treatment can address most problems related to water quality including NOM. However, these treatments are often chemically, energetically, and operationally intensive, focused on large scale applications. In addition, such treatments can produce residuals resulting from the process (sludge, toxic wastes) that can add to the problem of contamination and salting of freshwater sources [15]. Furthermore, conventional water treatments are not suitable for the removal of micropollutants [16]. To address this emerging problems, tertiary treatments are added to the process: examples of which being carbon adsorption and reverse osmosis [17]. These technologies are however expensive and bring additional problems, associated with the transfer of contaminants from one phase to another or generate some concentrated waste requiring further treatments. 1.4 Advanced oxidation processes (AOPs) Most organic compounds are resistant to conventional chemical and biological water treatments. For this reason, in the past years, new technologies have been developed to achieve the requirements of stringent standards for those substances exerting potentially toxic effects. Among these technologies, AOPs will probably constitute the most promising options and have been viewed favorably by industries and municipalities. AOPs, although using several reacting systems, are all characterized by the same chemical feature: production of hydroxyl radicals (HO ). HO radicals are extremely reactive species, they react with most organic molecules at rate constants in the order of M 1 s 1 and their steady state concentrations are between and M. Hydroxyl radicals are also characterized by their little selectivity, a useful attribute for an oxidant used in water treatment. The principal reaction pathways of HO with organic compounds include hydrogen abstraction from aliphatic carbon, addition to double bonds and aromatic rings, and electron transfer. These reactions generate organic radicals as transient intermediates, which then undergo 4

24 further reactions, resulting in final products corresponding to the net oxidative degradation of the starting molecule. AOPs can offer various possible ways for HO radicals production thus allowing a better versatility with the specific treatment requirements. Hydroxyl radicals can be generated by various approaches such as: Fenton reaction Fe2+ / H2 O2 ), Photo-Fenton reaction Fe3+ /H2 O2 / hν), UV-hydrogen peroxide (UV/ H2 O2 ), UV-ozone (O3 /UV), ultrasound, Vacuum UV (VUV) radiation, and photocatalysis (UV/TiO2 ). Another advantage is that AOPs usually operate at or near ambient temperature and pressure, making the process easier to control. One disadvantage of these AOPs is that, ideally the oxidation can cause total mineralization, but in reality partly oxidized byproducts which can be more toxics than the initial compounds are formed. Another problem lies in the high cost of reagents such as ozone, hydrogen peroxide or energy-light sources like ultraviolet light. However, for some AOPs, using solar radiation as an energy source can reduce costs. AOPs have found applications as diverse as groundwater treatment, soil remediation, municipal wastewater sludge conditioning, production of ultrapure water and volatile organic compounds treatment, and odor control. Depending on the properties of the waste stream to be treated and the goal of the treatment, AOPs can be employed either alone or coupled with other physicochemical and biological processes. While AOPs (e.g. UV/H2 O2 ) have found many applications in the water treatment industry for the removal of micropollutants, there are still challenges and issues with them. One such challenge is the effect of water characteristics on their efficiency. This research has tried to address these challenges by focusing on two emerging AOPs: VUV radiation and UV/TiO2 and studying the effects of water matrix (inorganic ions and NOM) on their efficiency in degrading micropollutants. 1.5 Thesis layout This thesis is organized in several chapters: Chapter One gives a general background on the challenge of water treatments and water contamination. Chapter Two presents a literature review of the advanced oxidation processes and the effects 5

25 of water characteristic on these AOPs. The literature on this topic is analyzed and the knowledge gaps are highlighted. Chapter Three presents the overall and the specific objectives of this study. Chapter Four describes the experimental setups and methodologies that are commonly used for achieving different objectives of the research. Setups and procedures specific to a particular objective or activity are presented in the respective chapters. Chapter Five contributes to meeting objectives of developing a TiO2 heterogeneous catalyst and of assessing the effects of the combination of photocatalysis with VUV. Chapters Six, Seven and Eight contribute primarily to meeting the objective of the effect of water matrix on the efficacy of the VUV process. In particular, in chapter six the effects of inorganic ions are studied, in chapter seven the ones of NOM and eight presents the effects of chloride. Chapter Nine presents the overall conclusions and the recommendations for future work. 6

26 Chapter 2 Literature review 2.1 Hydrogen peroxide/uv process UV/H2 O2 is one of the most widely studied AOP which has also been commercialized for the degradation of micropollutants. This process includes H2 O2 injection to the water and irradiating it with UV light (200 to 280 nm). During this process, ultraviolet radiation is used to cleave the O-O bond in the hydrogen peroxide and generate hydroxyl radicals. The overall quantum yield of this reaction is 1.0 for UV below 300 nm [18]. H2 O2 + hν 2 HO (2.1) The HO radicals produced are then able to react with the substrate leading to short chain acids byproducts or ideally to mineralization: HO + Substrate Products (2.2) In addition, the HO radicals produced can react again with H2 O2 : HO + H2 O2 HO2 + H2 O 7 (2.3)

27 High concentrations of H2 O2 are needed for the process because H2 O2 has a low molar absorption coefficient at 254 nm (19.9 M 1 cm 1 ) [19] and for this reason to produce enough HO radicals, relatively high doses are required. On the other hand, H2 O2 itself can scavenge HO radicals so high concentration of hydrogen peroxide can reduce the effectiveness of the process. In addition, high concentration of H2 O2 can reduce the penetration of photons reducing the direct photolysis of the target contaminant. For all these reasons, the ideal concentration of H2 O2 is in the range from 10 to 50 mg/l [20]. In addition, during the H2 O2 /UV process other reactive radicals are formed, such as superoxide radicals (HO2 ). Superoxide radicals are generated with dissolved oxygen according to the reaction [21]: 2HO + O2 2HO2 (2.4) The radicals formed can participate in the following range of reactions leading to the scavenging of HO radicals and to the formation of H2 O2 [21]: HO + HO2 H2 O + O2 (k = M 1 s 1 ) (2.5) HO + O2 + HO + O2 (k = M 1 s 1 ) (2.6) (k = M 1 s 1 ) (2.7) (k = M 1 s 1 ) (2.8) (k = M 1 s 1 ) (2.9) HO + HO H2 O2 HO2 + HO2 H2 O2 + O2 O2 + HO2 + H2 O+ H2 O2 + O2 + HO The application of UV/H2 O2 for commercial drinking water treatment involves a continuous flow system including one or more UV reactors and a H2 O2 dosing components. The two key parameters in the process are the UV fluence and the H2 O2 concentration. UV fluence is the total radiant energy of all wavelengths received by an infinitesimally small sphere. it is the product of the fluence rate and the exposure time [22]. The lamp technologies commonly applied in drinking water UV/H2 O2 applications are the next: low-pressure (LP) amalgam lamps and medium-pressure 8

28 mercury (Hg) lamps (MP). LP Hg and LP amalgam lamps emit two wavelengths 185 nm and 254 nm. LP lamps are constructed of a quartz envelope that permits transmission of only photons of 254 nm and above [23]. Medium Pressure Lamps have higher electrical power input compared with LPs: it is important to point out that UV radiation, in the range of nm, is absorbed by DNA disrupting its structure and leading to the deactivation of cell. For this reason, these types of lamps are used intensively for the disinfection of water, substituting chlorine methods [24]. The range of operating parameters for commercial UV/H2 O2 installation for the treatment of trace organic pollutants are the following: an initial H2 O2 concentration up to 15 mgl 1 and fluences up to 1500 mj cm 2. The commercial application of UV/H2 O2 in drinking water treatment processes dates to the early 1990s. Recently, research has focused on the efficacy of UV/H2 O2 on the removal of specific micropollutants. Kruithof et al. [25] studied the degradation of some organic contaminants in real surface water with a UV dose of 540 mj/cm2 (about 0.5 kwh/m3 ) and 6 mg/l H2 O2. Under those conditions, pesticides (atrazine), N-Nitrosodimethylamine (NDMA), methyl-t-butil etere (MTBE), dioxane, endocrine disruptors (bisphenol A), microcystine and pharmaceuticals (diclofenac, ibuprofen) could be removed by up to the required 80%. In addition, the effectiveness of H2 O2 /UV was studied for the degradation of other contaminants such as phenol [26], taste and odors compounds and methyl tert- butyl ether [27]. Some research focused also on the treatment of taste and odor causing compounds such as methyisoborneol (MIB) in water. UV/H2 O2 oxidized more than 70% of this compound at a UV fluence of 1000 mjcm2. The study was conducted with both MP UV and LP UV showing that MP UV consistently performed better than LP UV for MIB oxidation [28]. Zhang et al. [29] studied the photodegradation of 4-nitrophenol (4-NP) with UV/H2 O2 using a LP UV lamp placed in a double quartz sleeves annular photo reactor. The photon flux entering the reactor was 3.14 x10 6 Einstein 1. With these conditions 98% of 4-NP was removed within 12 minutes, and thus UV/H2 O2, in this case, was suggested suitable as tertiary treatments. There are numerous researches on the degradation of different organic compounds with UV/H2 O2 such as [30], [31], [32], [33]. It is difficult to compare the results among these studies since the conditions used such as source of irradiations, photo reactor geometry, and initial concentration of the 9

29 target compounds were different. However, some considerations can be made regarding the concentration of H2 O2 that has been used. Vogna et al. [30] and Lekkerkerker et al. [33] in particular, obtained good removal of the substrate with concentration of H2 O2 among 5-10 mg/l which is feasible for industrial applications. In particular, Vogna et al. [30] studied the degradation of carbamazepine with UV/H2 O2. The oxidation treatment caused an effective removal of the drug. The substrate (aqueous solution of 2.0 mm) was completely removed after 4 min treatment, and 35% removal of organic carbon (TOC) was obtained. Carbamazepine or benzoic acid solutions in water were irradiated with a nominal 17 W low-pressure mercury monochromatic lamp emitting at 254 nm (Helios Italquartz) in a L photoreactor. The power output of the lamp was 2.7x10 6 E s 1. Lekkerkerker et al. [33] studied the transformation of atrazine (ATZ), carbazepine (CBZ), diflonac (DFC) and sulfamethoxazole (SMX) by UV/H2 O2 treatment. Irradiations were carried out in a collimated beam. The treatments employed a range of doses between 300 to 700 mj/cm2 and a concentration of H2 O2 from 0 to 10 mg L 1. The results showed that DFC and SMX were removed by 100% after an irradiation of 230 mj/cm2 with a low pressure lamp. The addition of H2 O2, in this case, had little effect to the direct photodegradation. For ATZ, the addition of H2 O2 had an effect on its removal. The addition of 5-10 mg/l of H2 O2 increased the removal efficiency by 10-15%. CBZ was poorly removed with each doses and also with the addition of different concentration (from 0 to 10 mg/l) of H2 O2. In general, it is important to point out that the H2 O2 dose will depend on the DOC load of water. On the other hand Johnson et al. [31] and Li et al. [32], although they showed a very appealing removal of the substrate, used concentration of H2 O2 too high for any industrial applications ( 25 and 100 mg/l respectively). Johnson et al. [31] studied the metronidazole degradation by UV/H2 O2 process. The irradiation experiments were performed in a UV collimated beam equipped with a LP UV lamp emitting at nm. The initial concentration of metronidazole was 1 mg L 1 and the one of H2 O2 was 25 mg L 1. The rate constant was found to be 1.98 x 109 M 1 s 1, and thus the efficiency was found to be high, however, the concentration of H2 O2 was too high for any industrial applications (the concentration used is usually 10 ppm). 10

30 Li et al. [32] explored the photochemical degradation of typical herbicide simazine by UV/H2 O2 in an aqueous solution. The experiments were performed in a 20 L double cylindrical stainless reactor. The annular reactor was equipped with an immersible low pressure lamp with an output at 254 nm. The concentration of simazine was 100 µg L 1 and the light intensity was 45x10 6 einstein L 1 s 1. With a concentration of H2 O2 of 100 mg L 1 the removal was 100%. The study showed complete removal, however the concentration of H2 O2 was too high for any industrial application. These studies showed that UV/H2 O2 can successfully remove micropollutants from water however often the H2 O2 concentration used in these studies is not feasible for any industrial applications. In addition all these researches studied the removal of micropollutants in Millipore water, and the effect of real raw surface water on the efficacy of this AOP was not analyzed. UV/H2 O2 can be also used for the removal of NOM: some studies were conducted in order to assess the capability of UV/H2 O2 to degrade NOM. Kleiser and Frimmel [34] exposed river water, with DOC of 2.3 mg L 1 and to a LP lamp in the presence of 4, 8, and 16 mg L 1 H2 O2. Under these irradiation times DOC removal was minimal. The trihalomethane formation potential (THMFP) was observed after irradiation times, of 100 min. However, THM-FP dropped at an irradiation time longer than 100 min and that was attributed to the observed mineralization of DOC. Fluences were not reported, but it is likely that the oxidation conditions at 100 min were similar to those typically found in drinking water applications. Wang et al. [35] evaluated LP UV/H2 O2 oxidation for the degradation of NOM in water. About 90% of humic acid was removed after 30 minutes of irradiation in the 10 L batch reactor using a 450 W high-pressure Hg vapor lamp in the presence of 0.01% H2 O2. Specific UV fluences were not reported but due to the observation of TOC destruction it was likely fluences were very high (i.e., far greater than those in commercial applications). UV/H2 O2 AOP has proved to be an effective treatment for the removal of micropollutants, but it has several disadvantages. First H2 O2 has to be added and removed downstream or it needs to be coupled with BAC (Biological activated carbon) treatment: this increases the cost of treatment and it can make the technology complicated. Second, complete mineralization is rarely achieved, and sometimes the by-products created are more toxic than the initial contaminants. In addition, many 11

31 organics have a high absorption coefficient in the UV (254 nm) and for this reason they can act as inner filter therefore fewer H2 O2 molecules are photolyzed. On the other hand, other technologies should be applied for the removal of NOM, since THMFP was observed and the TOC removal took place at very high fluences. 2.2 Photocatalysis Photocatalysis with a semiconductor is another AOP which has shown great potentials for the removal of pollutants [36]. Photocatalysis with semiconductor is a process where a semiconductor catalyst is activated by UV radiation. When a semiconductor absorbs photon with energy higher than its band gap, an electron/hole couple (e /h+ ) is generated [37]. The electron/hole couple can either be used to generate electricity in photovoltaic cells or to drive a chemical reaction (photocatalytic process). One of the most widely studied photocatalysts in environmental decontamination is titanium dioxide, TiO2. It is used to degrade a large variety of organics, which can be totally degraded and mineralized to CO2, H2 O, and harmless inorganic anions. This performance is attributed to titania s highly oxidizing holes that can react directly with the adsorbed substrates or oxidize water to give HO. TiO2 can exist in different polymorphous: brookite, rutile and anatase. Anatase is the crystalline structure that shows greater photocatalytic activity for most reactions [38]. It has been suggested that this increased photoreactivity is because of anatase s slightly higher fermi level, lower capacity to adsorb oxygen and higher hydroxylation [39]. When TiO2 absorbs a photon with an energy equal to or higher than its band gap (Eg ), an electron and hole pair are generated: hν Eg + TiO2 e + h+ (2.10) The hole formed (h+ ) can then generate HO : h+ + H2 O HO + H+ 12 (2.11)

32 Studies on the degradation rate of various pesticides with various TiO2 catalysts have been performed. The degradation rates of dichlorovoros and phosphamidon in a Degussa p25 slurry (1 gl 1 ) at 0.5 mm and 0.25 mm, respectively, found to be and 0.02 x10 3 L 1 min 1 [40]. Gelover et al. [41] studied the degradation of chlorophenol using degussa P25 as catalyst. After three hours of irradiation the final dissolved concentration of 4-chlorophenol was 24% of its initial value. In addition, laboratory scale studies with artificial irradiation source and small volumes of water have shown complete contaminant disappearance for a range of pesticides such as triazines [42], lindane [43], methyl paration [44] at various concentrations. Some authors have studied the degradation of taste and odors compounds, in particular geosmin, with suspensions of Degussa P25 [45]. The suspension was 1% P25 and the samples were irradiated with a xenon lamp with a spectral output from nm. The photonic output was found 2.15x10 5 einstein min 1. Geosmin was degraded with more than 99% decomposition achieved in one hour. Another group [42] studied the degradation of atrazine by using Degussa P25 TiO2 as photo catalyst under simulated solar light. The irradiation was carried out by 1500 W xenon lamp and the flux in the range of nm, was 5.8x10 5 einstein min 1 cm 2.The results showed that in the part-per-million domain and beside TiO2 slurries and illumination, s triazine herbicides were rapidly degraded: only a few minutes were needed to reduce an initial concentration of 2 ppb down to less than 0.1 ppb. Other studies showed that the photocatalytic activity was dependent to the light intensity [46], [47]. Dalrymple et al. [48] published a review on the removal of pharmaceutical and endocrine disrupting compounds from wastewater by Degussa P25 TiO2 photocatalysis. The contaminants included were approximately 30. The rate constants were between 1 to 11.2 x10 6 M 1 s 1. Furthermore it was found that the addition of catalyst has shown to increase the degradation rate within a range of concentration, from 0.1 to 4 gl 1. All these studies reported the degradation of micropollutants obtained in the presence of TiO2 P25 Degussa. The results were promising, meaning that TiO2 can efficiently remove organics, 13

33 however, the TiO2 was in the form of a concentrate suspension which needs to be removed downstream and this an important disadvantage. For this reason the process is not easily manageable for industrial application, since the filtration of slurry can be challenging and expensive. Other researchers worked with immobilized photocatalysts which can be used in real application since it does not require to be removed downstream. For example, Carbonaro et al [49] and Zabar et al [50] studied the degradation of different organics. The conditions used in the studies, such as the type of photocatalysts, the source of irradiation, the type of photoreactor and the initial concentration of the micropollutants were different. However in both cases, the removal rate of the organics was lower compared to the one obtained with the Degussa P25 suspension (46% and 99% removal respectively). Carbonaro s group in particular studied the continuous-flow photocatalytic treatment of pharmaceutical micropollutants. The activity of immobilized photocatalyst thin films was studied in a serpentine-pattern plug flow reactor. The reactor was designed with five channels, each with the width of 1 standard microscope slide (25 mm) and length of 5 slides (375 mm), for a total of 25 film-coated slides placed along the entire reactor flow path. Five 15 W 18 inches long UV-A lamps with spectrum centered at 365 nm were suspended and centered over each of the five channels. The degradation of four pharmaceutical micropollutants (iopromide, acetaminophen, sulfamethoxazole, and carbamazepine) was monitored. The kobs values were found to be 0.97 h 1 for acetaminophen, 0.50 h 1 for carbamazepine, 0.49 h 1 for iopromide and 0.79 h 1 for sulfamethoxazole. Zabar et al [50], on the other hand, describes the photocatalytic degradation of 6-chloronicotinic acid (6CNA). Photocatalytic experiments were performed using immobilized titanium dioxide on six glass slides in a spinning basket inside a photocatalytic quartz cell. Three low-pressure mercury fluorescent lamps were used as a UVA (315 to 400 nm) radiation source (CLEO 20 W, 438 mm x 26 mm, Phillips; broad maximum at 355 nm). The photon flux in the cell was determined to be 2.3 x 10 5 Einstein L 1 s 1. The degradation within 120 min, obeyed first-order kinetics. The observed disappearance rate constant was k = min 1. Furthermore 46% mineralization was achieved within 120 min of irradiation. Another promising application of UV/TiO2 can be the removal of cyanotoxins [51]. The hy14

34 droxyl radicals, produced during TiO2 /UV process reacts in the order microcystins (1.1 x 1010 M 1 s 1 ) > cylindrospermopsin (5.5 x 109 M 1 s 1 ) > anatoxin (3.0 x 109 M 1 s 1) [52]. Because the rate constant with HO is several orders of magnitude larger than that for other oxidants (rate constant for ozone is 104 M 1 s 1 and for chlorine 1 M 1 s 1 or less depending on the conditions), TiO2 could be used succesfully for the inactivation of cyanotoxins, as alternative of UV/O3 and chlorinated processes. 2.3 Vacuum UV Vacuum UV is another AOP which was found to be a highly effective process for the degradation of micropollutants. It is an oxidant free, UV-based AOP, which relies on the formation of reactive through the photolysis of water by irradiating it with species such as HO, H, e aq, HO2, O2 VUV (photons of less than 200 nm). VUV photons can be generated by several sources, the most common being excimer lamps and ozone-generating low-pressure Hg lamps (VUV-Hg lamps), each presenting some advantages and disadvantages. Excimer lamps can emit a high powered quasimonochromatic radiation at various wavelengths (depending on the gas in the lamp) [53]. The electrical efficiency is in the order of 5-40%. For water treatments, Xe2 -excimer lamps, which emit at 172 nm have been proposed [54]. Because water absorptivity at 172 nm is high (550 cm 1 ) [55], photons are mostly absorbed in a 10 micron layer close to the lamp, with the generation of important mass transfer resistance that lowers the efficacy of the process. VUV-Hg lamps, on the other hand, emit about 10% of the radiation at 185 nm and 80-90% radiation at 254 nm, and a low percentage in the visible range. The absorptivity of water at 185 nm is reported to be 1.80 cm 1 [19]: in this case, photons are absorbed in about 0.3 cm layer close to the lamp. As a result, diffusive resistances are less significant and could potentially be minimized or eliminated by increasing the turbulence in the reactor. The problem of these lamps, however, is that only 10% of all radiation is in the VUV range. Two of the most important reactions occuring during the VUV radiation/water systems are [56]: 15

35 The photochemical homolysis of water: H2 O + hν<200nm HO + H (2.12) The photochemical ionization of water: H2 O + hν<200nm HO + e + H+ (2.13) The quantum yields at 185 nm of the reactions are 0.33 and 0.045, respectively. HO can be involved in a number of reactions among them being [56] : O2 + H HO2 (k = ) (2.14) HO + HO H2 O2 (k = M 1 s 1 ) (2.15) (k = M 1 s 1 ) (2.16) (k = M 1 s 1 ) (2.17) (pk = 4.8) (2.18) HO + H2 O2 HO2 + H2 O H2 O2 + HO2 HO + O2 + H2 O HO2 + H2 O H3 O+ O2 During the VUV process the degradation of pollutants can occur through several mechanisms, with the most significant of those being the reactions between HO and pollutants: Pollutant + HO intermediates CO2 + H2 O + mineral acids (2.19) Another possible reaction occurring in the system is the direct photolysis of the pollutant: Pollutant + hν254nm products (2.20) Pollutant + hν185nm products (2.21) Different researches have focused on the degradation of organics with VUV process. Oppenlander et al. [57]. studied the VUV-induced oxidation of organic micropollutant (C1 -C8 ) in ho16

36 mogeneous aqueous solution using a xenon-excimer (with an electrical input power of 155W) flow through reactor. The degradation efficiency depended on the structure of the molecules: after 80 minutes of irradiation the concentration decreased from 48 to 13 mg/l in DOC. However, due to the complexity of the reactions, further investigation concerning the TOC diminution, kinetics, the mechanism of intermediary product formation and of the subsequent mineralization would be valuable. In addition, the electrical energy dosage used in this study is not appealing (since it is too high) for industrial applications. Baum et al. [58] investigated the application of VUV to the degradation of chloroorganic compounds in a flow through reactor. The initial concentrations of trichloroethane, dichloroethene and tetrachloroethene were 103 mg/l, 25 mg/l and 0.4 mg/l respectively. After 60 minutes of VUV radiation, the concentration of chloroorganic water pollutants decreased by more than 97%. In another study involving citric acid and gallic acid as model pollutants, the apparent first-order rate constants of the degradation in aqueous solution were s 1 and s 1, respectively [59]. Many of these studies were performed in a batch and with a small reactor. Hence, there was no consideration given to the amount of energy used to degrade a specific concentration of contaminant. In addition the degradation of organics was performed preparing the samples with Millipore water hence without taking into account the effects that any compounds present in the water matrix may have. The removal of NOM, as already mentioned, is a major objective of drinking water treatments. Several studies on the application of VUV irradiation for NOM removal have been conducted and generally it could be shown that VUV light increases the biodegradability and eventually mineralize NOM: VUV was found the most effective AOPs among UV/H2 O2 and UV in the degradation of NOM. It was reported that VUV irradiation emitted by a low pressure mercury vapor lamp achieved a five times greater NOM removal, measured by the reduction of DOC concentration than UV irradiation at 254 nm [60], [12]. Imoberdorf et al. [60] used VUV radiation to degrade natural organic matter (NOM) in an annular reactor. It was found that after 180 min of irradiation the TOC of raw water decreased from 4.95 mg L 1 to 0.3 mg/l 1. In addition, it was found that the efficiency of the VUV process depends on initial water quality because the inorganic ions present 17

37 in water can absorb photons or scavenge HO radicals. This study showed a good removal of TOC, however the energy efficiency of the treatment was not discussed. Ratpukdi et al. [61] studied the removal of DOC with a stainless steel reactor. The reactor had a diameter of 30 cm and a height of 25 cm, and was filled with 16 L of the filtered water sample and it was equipped with 4 VUV lamps which had have a power input of 30 W per lamp, Mixing (60 rpm) was provided by a magnetic stirring system. The initial concentration of DOC was 4.18 mg/l and after 60 minutes of irrafiation the DOC was removed up to 30%. The consequence of the sequential NOM transformation under VUV irradiation for the disinfection byproducts formation potential were studied by Buchanan et al. [12]. Buchanan [12] studied the formation of hazardous by-products resulting from the irradiation of natural organic matter with VUV. It was found that after an initial increase the trihalomethane formation potential (THMFP) was reduced and the reduction correlated well with DOC mineralization. The study showed that during the VUV treatment the THMFP increased with the increment of the mineralization. The work did not provide any comparison with other AOPs regarding THMFP. On the other hand, the energy used to observe mineralization in this study (20000 kj/m3 ) was the same order as the energy required for reverse osmosis of salt water [62] and about 3 order of magnitude more than that for some commercial AOPs. For this reason the use of VUV for the removal of NOM or micropollutants, with the current technology, cannot be use for any industrial application due to its extremely high energetic cost. Other recent studies were focused on disinfection obtained with VUV radiation. The efficacy of UV and VUV disinfection of Bacillus (B.) subtilis spores in aqueous suspension at 172 nm and 254 nm was evaluated [63]. A Xe2 excimer lamp and a low pressure lamp were used as irradiation sources for these two wavelengths. The first order inactivation rate constant at 172 and 254 nm were and cm2 mj 1 respectively. A 2 log reduction of B. subtilis spores was reached with fluences (UV doses) of 870 and 40.4 mj cm2 at these individual wavelengths, respectively. Therefore, for the inactivation of B. subtilis spores, VUV exposure at 172 nm is much less efficient than exposure at 254 nm, since the deactivation of microorganism is due to the interaction of DNA with 254 nm photons and not due to the attack of HO radicals. This research indicated quantitatively 18

38 that VUV at 172 nm is not practical for microorganism inactivation in water and wastewater treatment. On the other hand, as already mentioned, if the source of VUV radiation is a low pressure Hg lamps, two wavelengths are emitted: 185 and 254 nm. In this case, the removal of micropollutants and disinfection (obtained with 254 nm ) are achieved simultaneously. 2.4 Potential of combined VUV and UV/TiO2 for micropollutant degradation With the mercury VUV lamps emitting 10% of radiation at 185 nm (VUV radiation) and 90% at 254 nm, there is a potential to combine the VUV process with photocatalysis. The radiation at 254 nm can be used either for water disinfection purpose or for the activation of the photocatalyst while the one at 185 nm can photolyse the water molecules for HO formation. This approach could enhance the overall efficiency and could use all the photons emitted by the lamps. Han et al. [64] chose pcba as a model compound and investigated the difference between water photolysis at less than 185 nm radiation and photocatalysis with irradiation at 254 nm. The pseudo first order reaction rate constant was found to be min 1 for TiO2 /UV, 0.21 min 1 for VUV and 0.23 min 1 for TiO2 /VUV. As shown, the rate constant was higher for the combination of TiO2 with VUV. In another work, Han et al. [65] measured the photocatalytic decomposition and mineralization of 4chlorophenol (4-CP), hydroquinone and 4-nitrophenol (4-NP) in aqueous solution using two kinds of low-pressure mercury lamps: one was UV lamp emitting at 254 nm and the other was VUV lamp emitting at both 254 and 185 nm. It was demonstrated that VUV irradiation led to the most efficient degradation of the organics. No significant difference of degradation rate was observed due to TiO2 catalyst under UV or VUV irradiation except the UV irradiated 4-chlorophenol solutions, which indicated that the main degradation reaction occurred on the catalyst surface. 2.5 Effect of water matrix Water matrix makes a significant contribution to the efficiency of every AOP. Two of the primary influencing constituents within the natural water matrix are inorganic ions and NOM which can react with HO radicals leaving less HO radicals available for the degradation of the target micropol- 19

39 lutants. In addition these constituents can absorb photons, acting as inner filter or being photolyzed. In the following section, the detailed effects of inorganic and NOM on AOPs are presented Effect of inorganic ions In natural water, there are several inorganic ions which could act in various ways towards VUV and TiO2 /UV. In general terms, inorganics can: Absorb the incident radiation making the process less efficient because less photons will be available for the photolysis of water, and/or H2 O2, or for the activation of TiO2. Scavenge HO reducing the concentration of HO available for the reaction with target contaminants. Produce radicals when photolyzed. There are some literature reports on the effect of several inorganic ions on the efficacy of various UV based AOPs. Usually a decrease in the efficiency has been reported with some ions [66]. It is unclear whether the efficiency is decreased because of the absorbtion of photons by those ions or because of the reactivity of ions toward HO or the combination of both. HNO2, NO2 and NO3 are known to absorb radiation in the UV region of the electromagnetic spectrum. The absorbtion spectra of HNO2, NO2 and NO3 show intense π π* bands in the UVC region (ε HNO2,371nm 2900 M 1 cm 1, ε NO2,205nm 5500 M 1 cm 1, ε NO3,201nm 9900 M 1 cm 1 ) and weak n π* at longer wavelengths [67]. Weeks et al. [19] reported the extinction coefficient of different compounds at 185 nm: for ethyl acetate at 0.03 mol L 1 was 158 M 1 cm 1, for H2 O2 at mol L 1 was 289 M 1 cm 1 and for sulfuric acid at mol L 1 was 186 M 1 cm 1. These results show that the absorptivity of inorganic ions at 185 and 254 nm could be important and should be considered. At the same time, it is well established that many inorganic ions, naturally present in water, can react with HO. The rate constant of most of the inorganic ions with HO is quite high ( L mol 1 s 1 ) [68]. On the other hand, there are some inorganic ions able to increase the efficiency of the AOPs because they can generate HO. Examples are: nitrous acid/nitrite, nitrate, and bicarbonate/carbon20

40 ate systems. The mechanisms of photolysis of nitrous acid, nitrite, and nitrate involve photolytic pathways that result in the formation of HO and different nitrogen species such as NO, NO2 and ONOO t, as primary photoproducts. The primary photoprocesses and the main subsequent reactions leading to the production of HO may be represented as follows [69]: Nitrous acid/nitrite systems: HNO2 + hν HO + NO φ = 0.01 (2.22) NO2 + hν O + NO φ = 0.01 (2.23) O + NO + H+ HO + NO (2.24) In the nitrate system, during irradiation at wavelength < 280 nm, peroxynitrite anion (ONOO ) is formed via isomerization of NO3 (the excited specie of NO3 ) NO 3 + hν NO3 NO3 ONOO + H+ ONOOH (NO3 + H+ )72% + (HO + NO2 )28% NO3 + hν O + NO2 + H2 O 2HO + NO2 (2.25) pk = 6.6 (2.26) (k = 0.9s 1 ) (2.27) φ = (2.28) Quantum yields reported for nitrous acid and nitrite photoinduced production of HO are higher than the ones reported for nitrate [70]. In general, the quantum yields increase with decreasing wavelength, this dependence being stronger for nitrate than for nitrite and still less significant for HNO2 [71]. 21

41 While NO3 exhibits little reactivity towards hydroxyl radicals, the kinetic role of HNO2 and NO2 as HO scavengers has to be taken into account [72]. NO2 + HO NO3 H (k = M 1 s 1 ) (2.29) NO2 + HO HO + NO2 (k1 = M 1 s 1 ) (2.30) Another ion that can have a large impact on water photochemical process is bicarbonate. The radical CO3 can be produced from HO and CO3 2, HCO3 or 3 DOM (excited triplet state). 3 HO + CO3 2 HO + CO3 (k = M 1 s 1 ) (2.31) HO + HCO3 H2 O + CO3 (k = M 1 s 1 ) (2.32) (k = M 1 s 1 ) (2.33) DOM + CO3 2 DOM + CO3 The radical CO3 can react with organic contaminants in water and provide some degree of decontamination. For instance, the reaction rate constants between CO3 and some pesticides such as fenthion, atrazine, malathion and phorate are, respectively, M 1 s 1, M 1 s 1, M 1 s 1, and M 1 s 1 [73]. Furthermore, the rate constant for benzene is M 1 s 1 [74]. The radical CO3 is less reactive than HO toward the degradation of organic compounds. At the same time, it is scavanged to a lesser extent by DOM in surface water. For this reason, CO3 can reach a higher steady-state concentration than HO, which may compensate for its lower reactivity [75]. Another ion that plays a significant role is chloride, which can absorb radiation at 185 nm and 254 nm and can be photolyzed forming HO radicals [76]. Also, it can scavenge photons and HO radicals making the AOP less efficient [77]. In addition, since Cl is formed it can react with the organic substrate forming toxic chlorinated byproducts. The reactions in which chloride is involved are: Cl + HO Cl + HO 22 (2.34)

42 Chloride can react with HO radicals and form chloride radicals which can react with the substrate and create chlorinated byproducts. On the other hand, the chloride radicals in these way formed can be scavenged according to: Cl + H2 O HOCl + H+ (k = M 1 s 1 ) (2.35) The hypochlorous radical ions can be involved in the following reactions [77]: HO + Cl HOCl (k = M 1 s 1 ) (2.36) HOCl HO + Cl (k = M 1 s 1 ) (2.37) HOCl + H+ Cl + H2 O (k = M 1 s 1 ) (2.38) (k = M 1 s 1 ) (2.39) Cl + H2 O HOCl + H+ As this manifold of reactions show, ph makes an important contribution to the production of chlorine radicals and it is clear that hydroxyl radicals and chloride ions are in equilibrium with the hypochlorous radicals. It is clear that the role of chloride is complicated; on one hand it can create HO radicals, on the other it can scavenge HO radicals. One research in the literature [78] studied the photooxidation with UV/H2 O2 of reactive orange dye with and without NaCl. The photoreactor used in the study consisted of 8 medium pressure mercury vapor lamps set in parallel emitting 365 nm wavelength. With 20 mm of H2 O2, 98.3% of the dye was degraded after 150 min of irradiation. With the addition of 1g/L of NaCl a small decrease was observed (5%). This behavior was explained based on the scavenging effect of chloride towards HO radicals. Another study [79] investigated the decolorization and degradation of commercial reactive azo dyes in the presence of different inorganic ions. It was deduced that all the anions examined affect the decolorization rate adversely but to a varying degree. The trend was somewhat different at lower anion concentration (0.01 M) when compared to a higher one (0.1 M). Specifically, at 0.01 M, the order of inhibition measured was as follows: H2 PO4 > CO3 2 > HCO3 > NO3 > Cl > SO4 2 while at the higher concentration of 0.1 M, the order of inhibition measured was: H2 PO4 > Cl > 23

43 HCO3 > CO3 2 >SO4 2 > NO3. Unfortunately, the researches did not do a systematic study on the different effect that each ion could have on the process, such as the absorption coefficient and their capability to scavenge HO radicals. Yuan et al. [80] investigated the effects of chloride ion on the degradation of Acid orange with sulfate radical-based advanced oxidation. It was found that chloride at concentration higher than 5mM can increase the efficacy of the process since chloride radicals are formed and they can assist HO radicals during the degradation of the organic Effect of natural organic matter (NOM) NOM plays an important role in the degradation of micropollutants. At high concentration, NOM can decrease the overall process efficiency. It is known that NOM absorbs radiation and attenuates UV. Hence, the photons absorbed by NOM are no longer available for photochemical reactions. Furthermore, NOM can scavenge HO with a very high reaction rate constant of about M 1 s 1 [81]. The reactions of HO, produced by different AOPs, with NOM proceeds in three different ways: i) addition of HO to double bonds, ii) abstraction of an H-atom, which yields carbon centered radicals, and iii) reactions whereby HO gets an electron for an organic substituent. The carbon centred radicals then react very rapidly with oxygen to form organic peroxyl radicals. The reactions of peroxyl radicals among themselves can lead to the production of ketones and aldehydes and/or carbon dioxide [82]. Another problem specific to heterogenous TiO2 /UV is that NOM could adsorb on the catalyst, leading to its deactivation [82]. Low concentration of NOM, on the other hand, could actually accelerate the degradation of target contaminants due to the formation of reactive species [83]. NOM can act as a sensitizer and the energy transfer, in the presence of oxygen, can lead to singlet oxygen formation, which is an efficient oxidant for a variety of unsaturated organic compounds. Photosensitized transformations of organic chemicals in surface water are mostly initiated through light absorption by chromophores present in dissolved organic material (DOM). The occurrence of several reactive photooxidants, including the hydroxyl radical, the carbonate radical, singlet molecular oxygen, and solvated electrons, has been documented, and photo stationary steady state concentrations of these species have been calculated [84]. Recently, details of the fast photooxidation of organics by short lived triplet 24

44 states of DOM have been reported. It was postulated that 3 DOM reacts with organics by electron abstraction and/or hydrogen transfer [85]. The reactions presented here are the ones that have been proposed: DOM + hν DOM DOM + O2 1 O2 (2.40) (2.41) DOM + hν 1DOM (2.42) DOM 3 DOM (2.43) 1 3 DOM + P 3P (2.44) DOM + P POx + DOMRed (2.45) 3 where DOM is the dissolved organic matter in an excited state, 1 DOM and 3 DOM are DOM in a singulet and triplet state, respectively, Pox is the oxidized substrate, 3 P is the substrate in a triplet excited state, and DOMred is the DOM reduced. The impact of NOM was studied by different researchers [86], [85]. Depending on the NOM initial concentration and on the type of irradiations, different results were obtained: NOM could act as HO scavenger or as a sensitizer. In particular, Wu et al. [86] studied the degradation of metolachlor with UV and UV/H2 O2. UV irradiations were carried out using a 15 W germicidal low pressure UV lamp. The concentration of metolachlor was 2 µm and the concentration of H2 O2 varied from 0 to 50 mgl 1. An increase of the NOM concentration from 1.6 to 4.7 mg L 1 resulted in a decrease of the destruction rate of metolachlor. With 25 mgl 1 of H2 O2 and 1.57 mgl 1 DOC, the first order destruction rate of metalochlor was 6.3 x 103 s 1, and with 4.71 mgl 1 decreased to 2.7 x 103 s 1. This effect was explained with the scavenging effect of NOM towards HO radicals. On the other hand, Canonica et al. [85] studied the transformation of phenols in water and their photosensitization by dissolved natural organic material in water. Different dissolved natural organic materials photo- sensitized the transformation of a series of methyl and methoxy phenols at ph 8 with a very similar high selectivity (reactivity range 50). It was found that the photooxidation at ph 8 was not controlled by singlet oxygen. Deuterium isotope effects suggested an electron 25

45 transfer mechanism. The reactive triplet state concentration was estimated to be M in the top meter of Lake Greifensee under summer noon sunlight, and this leads to a half-life of 7 h for 2,4,6- trimethylphenol. Further, it was found that still uncharacterized photooxidants derived from the dissolved organic material are also involved in the phototransformation of the phenols The relative contribution of inorganic ions and NOM on the scavenging effect and inner filter effect they may have is not very well understood for VUV process. Most literature studies have focused on the reaction mechanisms of different micropollutants [87], [59], and the degradation of NOM and the formation of byproducts [12]. Only few researchers conducted a study on the efficacy of VUV in the presence of NOM or inorganic ions [88], [89], [90]. A research group [88] conducted a study on the degradation of 2,4 dichlorophenoxyacetic acid (2,4-D) in Milli-Q water and in the presence of raw water with 5.66 mg/l TOC. The 2,4-D degradation rate decreased significantly (72%) when raw surface was used. The results showed that water matrix can play an important role on the efficacy of VUV. However, in this study there was not any discussion on the mechanism responsible to the decrement of VUV efficiency in the presence of NOM. Other studies [89], [90] were focused on the effect of nitrate on VUV. All these studies have reported a decelerating effect of the degradation in the presence of nitrate due to its capability to absorb 185 nm photons. For this reason it can act as inner filter, and thus less photons are available for the photolysis of water. However, the mechanism responsible to this decrement of efficiency was not studied in details and the research analyzed only the effect of nitrate and not any other ion. 2.6 Knowledge gaps VUV oxidation and photocatalysis with TiO2 can be used in drinking water treatment. VUV has not been studied thoroughly and in details; the research around it is limited. The chemical mechanism of the process has always been a challenge because nearly every species absorbs 185 nm photons, and many are photolyzed. For this reason it is difficult to separate the effect of photons from the one of HO radicals on the overall degradation. Understanding the effect of water matrix, in particular, the effect of NOM and inorganic ions, will be valuable for real and commercial applications of VUV process because they make an important contribution in the efficiency of process. Therefore, 26

46 research is necessary to enrich existing knowledge of the scavenging effect of the water matrix toward 185 nm radiation and HO radicals. In addition, to the best of my knowledge, no prior research has investigated the process with just 185 nm radiation, because the lamps mostly used emit at 185 nm and 254 nm. In this case the total output of the lamp is a mixture of two wavelengths (254 and 185 nm). Proper understanding of the process with just one type of radiation and of the water matrix would be valuable for future applications since it will possible to quantify the impact of different water matrices on the VUV process. Therefore, there is an opportunity to develop some tools (absorption coefficients and scavenging effects of different components of water matrix) that can be used to predict the efficiency of VUV in different conditions. Furthermore, as pointed out before, the VUV lamps emit at 254 nm and 185 nm. Hence, it is valuable to study the combination of UV/TiO2 and VUV in a reactor, allowing for all the photons to be used: those at 185 nm for the photolysis of water, leading to the formation of HO radicals, and those at 254 nm for the activation of the semiconductor. To the best of my knowledge no prior research has explored this possibility. 27

47 Chapter 3 Research objectives The main objective of this research was to study the effect of VUV process at degrading target micropollutants and to assess the effects of water matrix on the VUV treatment. These overall objectives were achieved through the following specific objectives: Study the absorbance of different ions at 254 and 185 nm and their scavenging effect towards HO radicals. Study the absorbance of different sources of NOM at 254 and 185 nm and their scavenging effect towards HO radicals. The secondary goal was to study the combination of VUV with TiO2 /UV process. This objective was achieved through the following specific objectives: Develop a heterogeneous TiO2 photocatalyst through sol-gel technique with the following characteristics: robust, high efficiency, easy to produce and stable on the support. Assess the synergistic and beneficial effect or both of incorporating heterogeneous TiO2 photocatalyst in VUV reactors. In order to study in detail the effect of water matrix, first the absorption coefficient of each compounds was determinated. Secondly, the scavenging effect was explored following the degradation of a target micropollutant in the presence of different compounds with UV/H2 O2. To study the synergistic effects of incorporating heterogenous TiO2 photocatalyst in VUV reactor, the internal walls 28

48 of the annular reactor were coated with the previously developed catalyst that showed the highest efficiency and the highest stability. 29

49 Chapter 4 Experimental setups and procedures In this chapter, the general and common experimental setup and procedures that have been used in this work will be described and explained. Specific experimental procedure associated with each task or objective will be explained in the chapter dedicated to that task. 4.1 Setups Photoreactors For the photocatalytic experiments two reactors were used, a batch reactor and a flow through reactor Batch reactor The batch reactor was used in order to evaluate the photo efficiency of the synthesized TiO2 coatings. The reactor was a differential reactor (Figure 4.1 (b)) which consisted of a 63-mm-wide aluminum reactor designed to allow water to flow through a 225 mm long passage of 25 mm (width) and 3 mm (height), and over the coated glass plates. The reactor was coupled with a pump, a storage tank (a 4 L flask), a flow meter, and connecting tubing (Figure 4.1 (a)). A UV lamp emitting at 254 nm (Light Source Inc. GPHVA357T5L with an output power of 11 W) was used as the source of radiation, and the volume of water utilized in the experiments was 1 L. Microscope slides (25x75 mm) coated with the TiO2 catalysts were placed in the reactor channel, which was closed 30

50 Figure 4.1: Batch experimental setup a) and differential reactor b). The batch reactor consisted of a sparging beaker, a pump, a flowmeter, the differential reactor and UV lamps. with a quartz plate. Two UV lamps were placed overhead mounted under an aluminum reflector. The lamp centerlines were located 45 mm above the quartz plate and 20 mm away from each other. The photocatalytic reactor was operated under differential mode with complete recirculation of the reagents Flow through reactor The flow through reactor was used to test the coupling of UV/TiO2 with VUV. The interior of the reactor was coated with the photocatalyst that showed the best efficiency and adherence to the support in the batch reactor. The photoreactor (4.2) was made of a suprasil quartz envelope with an annular configuration. A VUV-Hg lamp (Light Sources, Inc. G10T51-2-VH), which emits radiation at 254 nm (90%) and 185 nm (10%) or a UV lamp (Light Source Inc. GPHVA357T5L), was placed 31

51 Figure 4.2: Flow through reactor experimental setup; the reactor consisted of a annular reactor (a), a pump (b), a feed storage tank (c), and a treated water tank d). longitudinally at the axial center of the reactor in such a way that the envelope of the lamp formed the internal wall of the reactor. The external and internal diameters of the annular reactor were 2.5 and 1.5 cm, respectively, giving an annulus thickness of 0.5 cm and a total reactor volume of 85 cm3. The reactor was coupled with a peristaltic pump, a feed storage tank, a treated water tank and connecting tubing UV and VUV reactors UV and VUV irradiations were performed in two collimated beams, one involving only 254 nm (UV collimated beam) and the other one involving only 185 nm (VUV collimated beam). The kinetics of the photochemical reactions associated with UV/H2 O2 were studied in details with collimated beams. When properly designed, collimated beam can provide quasi-collimated and uniform radiation, which allows obtaining valuable kinetic data that can be easily analyzed. In addition, a small volume of the reacting solution is required for each test and, if the solution is thoroughly mixed, the concentration of reagents in the reaction vessel is uniform. Conventional UV-collimated beam setups, however, cannot be used to study VUV process since oxygen in air partially absorbs photons 32

52 at wavelengths shorter than 190 nm. In spite of the advantages of collimated beams, there had not been any setup specifically built to study the kinetics of VUV-induced reactions. To the best of my knowledge, VUV process had only been studied in batch and flow-through lab scale reactors. In such cases, the concentration of micropollutants and the local incident radiation may not be uniform in the reactor volume, which represent a clear obstacle to interpret the kinetic data obtained. For all these reasons a VUV collimated beam was designed and built UV collimated beam set-up The experiments were conducted with a collimated beam bench scale reactor equipped with a low pressure amalgam lamp emitting light at 254 (Light sources Inc. GPHVA357T5L) positioned 28 cm above a circular stirred reactor chamber. The reactor chamber was a petri dish of 5 cm in diameter, and the water path length was 3.33 cm (Figure 4.3). The samples were positioned under the collimated beam on top of a magnetic stirrer. The flux of the lamp will be measured with an actinometry. The method of choice was the potassium-ferrioxalate actinometry. 33

53 Figure 4.3: UV collimated beam set-up (a) and cross section of the UV collimated beam (b); the collimated beam consisted of a UV lamp, a reactor chamber and a stirrer. 34

54 4.1.3 VUV collimated beam set-up The VUV collimated beam (Figure 4.4) is comprised of an ozone generating amalgam Hg lamp (Light Sources GPHVA357T5VH/4W) placed in a T-shape PVC enclosure, which is continuously purged with nitrogen to remove oxygen present in air. Two Teflon cylinders were placed around the quartz envelope to cover the ends of the lamp in order to improve the collimation of the radiation. The top of the T-shaped enclosure is sealed by a PVC head with a Suprasil quartz and/or an optical filter, which allows studying of a specific wavelength during the batch kinetics studies. The quartz and filter are held by three Teflon Orings, which seal the enclosure. Two different combinations of quartz windows and filters are used. To conduct experiments with only 185 nm radiation, a special optical filter is used to block 254 nm. This filter is formed by a MgF2 flat quartz coated with a metal-dielectric-metal (MDM) aluminum thin film (esource Optics). Its peak transmission wavelength is at nm and its diameter and thickness are 50.8 mm and 4.2 mm, respectively. An analytical quality flat UV grade Fused silica window was placed under the 185 nm filter to protect the metallic coating from ozone that could be formed inside the enclosure. To expose samples to both 254 nm and 185 nm photons, a head with an analytical quality flat UV grade fused silica can be used. When only 254 nm is required, a regular germicidal Hg lamp (Light Sources 35 GPHVA357T5L/4W) is used instead of the ozone generating Hg lamp. To irradiate water samples, two special cylindrical reaction vessels were built. The first vessel is an open top container made with regular quartz, except the bottom part made of Suprasil quartz to allow 185 and 254 nm radiation to be transmitted. The second vessel is similar as the first one, but with the top part closed, to make it possible to work with volatile compounds. The diameter and height of both vessels are 4.8 cm and 1.5 cm, respectively. A stirrer is mounted on top of the setup to mix the solution during the irradiation. Ideally, a collimated beam setup should be able to provide uniform and collimated radiation to the solution. That is, the incident radiation at the window of the reaction vessel should be uniform and the direction all the photons should be normal to the window of the vessel. However, in real setups, these two requirements cannot be completely satisfied, and quasi-collimated radiation with small changes in the incident radiation is achieved. To analyze the uniformity and extent of collimation of the radiation reaching the reaction vessel, 35

55 a radiation model based on the Monte Carlo technique was developed in collaboration with Dr. Gustavo Imoberdorf (research associate in the department of chemical and biological engineering, UBC). The approach and hypotheses considered in the model were the same as those presented elsewhere [91]. Figure (4.5) shows the normalized incident radiation at the bottom of the vessel. A fairly uniform distribution of the radiation was obtained and the difference between the maximum incident radiation at the center of the vessel and the minimum at the borders was less than 5%. Similarly, modeling results shows a small divergence of the radiation figure 4.6 and 90% of the beams are deviated less than 20 degrees from the normal. The low dispersion of radiation was reduced significantly by the two Teflon tubes surrounding the quartz sleeve (Figure 4.4), which cut the beams with a high divergence of radiation. Modeling results show that the developed collimated beam can provide uniform and quasi-collimated radiation. 36

56 Figure 4.4: Diagram of the collimated beam and of the top of the PVC enclosure. (1) motor, (2) reaction vessel, (3) stirrer (4) head of the enclosure, (5) enclosure, (6) VUV lamp, (7) Teflon cylinder (8) Orings, (9) quartz sleeve, (10) PVC head of the enclosure, (11) optical filter, and (12) Suprasil quartz. Figure 4.5: Radial distribution of radiation on the surface of the reaction vessel. 37

57 Figure 4.6: Angular distribution of radiation on the surface of the reaction vessel. 38

58 4.2 Experimental procedures The followings are common and general procedures used in this project: Actinometry at 254 nm The protocol that was used for measuring the flux at 254 nm was an iodide/iodate actinometry [92]. An iodide-iodate-borate solution was prepared and irradiated with UV desired geometry. The iodideiodate chemical actinometer on exposure to UV forms triiodide, the proposed reaction having the following stoichiometry according to [92]. 8KI + KIO3 + 3H2 O + hν 3I + 6HO + 9K+ (4.1) The absorbance of triiodide can be determined spectrophotometrically and used to calculate the UV fluence. After that the absorbance was measured at 450 nm. In order to calculate the the quantum yield, the following equation was used: φ = 0.75 [ (T 20.7)] [ (C 0.577)] (4.2) where φ is the quantum yield, T is the solution temperature in Celsius degrees and C is the molar concentration of iodide. Once the quantum yield was calculated, the fluence was quantified: F= A450nm V 1000 ε450 Φ A (4.3) where F is the fluence (mj/cm2 ), A450nm is the change in absorbance at 450 nm (cm 1 ), V is the solution volume (L), ε450nm is the molar absorption coefficient of triodide at 450 nm (1600 M 1 cm 1, φ is the quantum yield (mol/einstein), A is the area irradiated (cm2 ), 4.72x 105 is the converting factor from einstein to Joules and 1000 is the converting factor from joule to millijoule. The fluence that was obtained with the geometry used during the UV experiments was found to be 0.29 mw/cm2. 39

59 4.2.2 Actinometry at 185 nm The protocol that was used for measuring the flux at 185 nm was a methanol actinometry [93]. The degradation of methanol during VUV was used to calculate the incident radiation, since the parameters affecting primarily the rate of methanol degradation, i.e., the incident photon rate, the concentration of dissolved molecular oxygen, the initial methanol concentration were determined and optimized for a general actinometric procedure. The incident radiation was calculated following: P0 = d[meoh] V dt φh2 O ζh2 O + φmeoh ζmeoh (4.4) where P0 is the incident radiation, φh2 o is the quantum yield of water (0.375), φmeoh is the quantum yield of methanol (1), ζh2 O is the fraction of photon absorbed by water (0.99), ζmeoh is the fraction of photons absorbed by methanol (0.0038), and d[meoh] dt is the apparent measured rate of methanol photolysis (mol L 1 s 1 ). The concentration of methanol was measured following an enzymatic procedures. The methanol was determinated through the oxidation of methanol to formaldehyde with an alcohol oxidase enzyme (Pichia Pastoris), followed by condensation with 2,4-pentanedione to yield the colored product 3,5-diacetyl-1,4-dihydro-2,6-dimethylpyridin [94] Hydrogen peroxide measurement The residual concentration of H2 O2 for the UV/H2 O2 and the formation of H2 O2 during the VUV process were measured with the protocol described by klassen et al.[95]. The procedure was as follows: first reagent A, which consisted of 10 gr of Potassium hydrogen phtalated dissolved in 500 ml of Distilled water (DW) was prepared. After that Reagent B which consisted of 33 gr of potassium iodide, 1 gr of NaOH and 0.1 g of Ammonium molybdate tetrahydrate in 500 ml of distilled water was prepared. The analysis consisted in the measurement of the absorbance (A) at 351 nm of a solution prepared with 2.5 ml of reagent A with 2.5 ml of reagent B with 0.5 ml of sample diluted with distilled water in a 10 ml volumetric flask. The blank (A0 ) was prepared in the same way without the addition of the sample. In order to calculate the concentration of H2 O2 the 40

60 following equation was applied: H2 O2 (ppm) = (A A0 ) D S (4.5) where D is the additional dilution (1 if none) and S is the sample volume (0.5 ml) Absorption coefficient measurements at 185 and 254 nm The absorbtion coefficient of different compounds was measured at different concentrations with an Agilent Cary 4000 spectrophotometer. The spectrophotometer worked in a double beam mode under nitrogen purging (20 L/min). The spectra were collected from 300 nm to 180nm. The scan rate was 80 nm, the temperature was set at 25 C and the spectral bandwidth was set to 2. The baseline was a cuvette with Millipore water. Each sample was placed in the measurement compartment and purged with N2 for 15 minutes in order to remove all the oxygen before the measurement started. Measuring the absorbance at 185 nm is difficult for the typical spectrophotometers because it is at the edge of the vacuum UV band where air and, in particular, oxygen absorbs the bulk of these high energy photons. For this reason, it is necessary to purge the optical compartments of the spectrophotometer with dry, pre-purified N2 gas. Another problem is that the transmission of the quartz suprasil cuvette falls significantly at shorter wavelength. In addition, the filters, lenses, and gratings in the spectrophotometer may perform poorly at these high energies as they are optimized for the UV and IR regions. For all these reasons the absorbance of different compounds was measured at different concentrations in order to assess if the absorbance was linear and followed Lambert Beer law. If a straight line was obtained the measurements were precise and the absorption coefficient was the slope of this straight line ph and Dissolved oxygen (DO) measurements The ph was monitored during the irradiations using a ph meter (Thermo Orion PerpHecT LogR 1330 meter, 9206BN electrode).the concentration of dissolved oxygen was monitored with a dissolved oxygen meter (YSI 52 meter, YSI 5909 probe). The ph was found to be 5.2 a the beginning of the irradiations and 5.6 at the end. The effect of ph on the absorbance of ions and its effect on 41

61 the degradation of the substrate was not explored. The effect of ph on the percentage of HO radicals formed in the system was studied only in the presence of different concentration of chloride. The dissolved oxygen was monitored at the beginning of the irradiations and it was found to be 5.5 mg/l Reproducibility and data accuracy For the absorption coefficient measurements, for each ion the absorbance was measured at three different concentrations in order and since the absorbance was found to be linear the validity of the methodology was proven. H2 O2, TOC, HPLC analysis were all done in duplicate and the value considered was the average of the two runs. In other words, the respective analytical methods were carried out twice for each sample and the average reported. In order to consider the reproducibility of the UV/H2 O2 and VUV treatment experiments, experimental runs were done in triplicates so each point that is plotted is the average of three runs with its own standard deviation. Thus the error bars on figures represent the standard deviation for the three replicate of each sample. Thus, error bars are presented to indicate the statistical reproducibility of the tests. 42

62 Chapter 5 TiO2 photocatalyst development and evaluation of photocatalysis coupled with VUV This chapter presents the results regarding the secondary objectives of this thesis. As discussed in the literature review chapter, one emerging and promising UV-AOP is Vacuum UV (VUV) irradiation which relies on the formation of hydroxyl radicals (HO ) by the photolysis of water induced by VUV radiation (Gonzalez et al., 2004). Low pressure Hg or amalgam Hg VUV lamps, which emit about 10% at 185 nm (VUV) and 90% at 254 nm (UV), can be used as the source of radiation. Among the advantages of Hg-VUV lamps are that they are easily available in different sizes and powers and are very affordable. Since only 185 nm photons are capable of promoting water photolysis, 90% of the energy (254 nm photons) is not utilized. A potential option to take advantage of the 254 nm radiation is the integration of VUV process with photocatalysis, whereby 254 nm photons can activate the photocatalyst to promote redox reactions leading to the degradation of organic contaminants. In this work, five different TiO2 coatings were synthesized and their photoactivities, physical characteristics as well as their potentials for integration with VUV photolysis were evaluated. 43

63 Materials and methods Synthesis of the photocatalysts TiO2 coatings were synthesized with different sol-gel techniques, utilizing titanium tetraisopropoxide (TIP) as the precursor. TiO2 colloids were prepared and deposited on microscope slides (75mm x 25mm). In total, five different photocatalysts were synthesized. Photocatalyst A was prepared based on the method proposed by Yip et al. [96]. TiO2 colloids were synthesized by adding 25mL of TIP to 5mL of glacial CH3 COOH (Fisher Scientific Canada, 99.7%) at 4 C. To this solution, 250 ml of 1M concentrated HNO3 (Fisher Scientific Canada, 70%) was added, and the sol was stirred for 12h. Photocatalyst B was prepared with the colloidal suspension obtained by mixing 5 ml of TIP, 1mL of polyethylene glycol (PEG) (SigmaAldrich Mn400), and 0.5mL of CF3 COOH (SigmaAldrich 99%). Ethanol was added at the end until the volume reached 50mL. Photocatalyst C was prepared following the recipe developed by Yamazaki-Nishida et al. [97]. A suspension of TiO2 was synthesized by mixing 415mL of distilled water with 4.5mL of HNO3 (Fisher Scientific Canada 65%) and 35 ml of TIP. The sol obtained was stirred at 80C for 10 h. Photocatalyst D was prepared according to the method proposed by Keshmiri et al. [98]. Ninty-six millilitres of denatured ethanol, 6.4mL of H2 O, 16 ml of concentrated hydrochloric acid (HCl) (Fisher Scientific Canada, 99%), and 120mL of TIP were mixed, and stirred for 2 h. After that, 38 g of Degussa P25 (sigma Aldrich) was added to the sol and the suspension was stirred for 12 h. Photocatalyst E was prepared in the same manner as photocatalyst A, except here 50 g of Degussa P25 TiO2 was added to the sol and the suspension was stirred for 12 h. The sols obtained were deposited on glass microscope slides with the dip coating technique with a deposition speed of 1 cm min 1. The obtained coatings were calcinated at 400 C for 3 h. The 44

64 ramping up temperature rate was 5 C every 5 minutes. At the end of the calcination process, the furnace was turned off, and the coatings were left inside the furnace to cool down for 2 h Physical characterization of the photocatalysts The surface and morphology of the coatings were examined using a scanning electron microscope/energy dispersive X-ray spectrometer (SEM/EDX PHILIPS XL30 with Bruker Quantax 200 Microanalysis system and light element detector Silicon DriftDetector XFLASH 4010). Solid samples were ground into fine powder with a corundum mortar and then smeared them onto a glass slide with ethanol. Step-scan X-ray powder-diffraction data were collected over a range of 3802q with CoKa radiation on a Bruker D8 Focus Bragg-Brentano diffractometer equipped with an Fe monochromator foil, 0.6-mm (0.3) divergence slit, incident- and diffracted-beam Soller slits, and a LynxEye detector. The long fine-focus Co X-ray tube was operated at 35 kv and 40 ma, using a take-off angle of 6 C. Phase identification was done using the International Centre for Diffraction Database PDF-4 and Search-Match software by Siemens (Bruker). To determine the UV-Vis spectral transmittance of the coatings, the sols were immobilized on quartz slides instead of on the microscope slides since glass absorbs UV radiation. Spectra were registered with a UV-Vis-Nir Cary spectrophotometer. The transmittance of the coating was registered from 400 to 200 nm. The sizes of the TiO2 particles of the colloidal suspensions were measured with light scattering technique (Mastersizer 2000). The reflection index was set at The size range of the instrument was from to 2,000 µm. The light source was a helium neon laser Photocatalytic activity assessment Tests with a lab-scale glass plate photocatalytic described in Section were performed in order to measure the photocatalytic activity for the degradation of a target micropollutant, 2,4-D. Contaminated water samples were prepared by adding 2,4-D (10 ppm) to Millipore water. Contaminated water was placed in the tank, and it was recycled through the reactor at 0.25 L min 1. Each run lasted 90 min. In the first 30 min, the UV lamp was off to assess the adsorption of the micropollutant on the photocatalyst under dark condition. Then, the lamp was turned on, and samples were taken every 10 min for 60 min. The concentration of the micropollutant was quantified 45

65 using a high-performance liquid chromatograph (HPLC, Dionex 2695) equipped with C-18 column (4-micronmeter particle diameter) and a UV detector. Methanol/water/acetic acid (58%:40%:2% v/v) were used as mobile phase. The injected sample volume was 100 µ L. The flow rate of the mobile phase was 1 ml min 1 and analyses were conducted at λ = 280 nm via UV detection. Three replicate tests were conducted for each of the five photocatalysts. In order to test the stability of the TiO2 films, attrition tests were performed. Initially, the photocatalytic efficiency of the fresh films was assessed. Then, water was recycled over the photocatalyst surface at a high flow rate (1 L min 1 ) in order to determine the amount of photocatalyst lost under a high turbulence flow for 24 h. Finally, the photocatalytic efficacy was evaluated again. Differences in the photocatalytic activity of the fresh photocatalyst and that after 24 h could represent the potential loss of photocatalyst from the surface due to continuous exposure to a turbulent flow during an extended period of time. As for the deactivation test, consecutive experiments (seven runs) were performed using the same photocatalyst in order to determine the potential loss of activity due to deactivation that may be caused by the deposition of photooxidation by-products on the photocatalyst surface Results and discussion Physical characterization of the photocatalysts TiO2 films in photocatalysts A and C were not observed in SEM micrographs, possibly because those films were too thin. SEM micrographs of composite photocatalysts D and E are shown in Figure 5.1 and Figure 5.2, respectively. Photocatalyst D was not homogeneous and showed significant fragmentation and cracking, probably due to the evaporation of the solvent during calcination. Photocatalyst E, on the other hand, was more homogeneous and better attached to the support. These early results gave an indication that photocatalyst E would provide a better attachment to the support, providing higher attrition resistance. The percentages of crystallinity and different polymorphs in TiO2 play a major role in the efficiency of the photocatalyst. It is known from the literature that rutile has a lower photocatalytic 46

66 activity than anatase [99]. However, the activity of Degussa P25, which consists of anatase and rutile (4/1 w/w), exceeds that of pure anatase. For this reason, it is widely believed that photocatalysts with mixtures of rutile and anatase show a higher degree of activity, and hence, this ratio was considered important in this investigation. Figure 5.3 shows the XRD spectra obtained for the five photocatalyst coatings, and Table 5.1 shows the percentage of crystallinity and the percentages of anatase, rutile, and brookite in each coating. Photocatalyst A was 100% crystalline, but the ratio of anatase to rutile was not close to the one of Degussa P25. Photocatalyst B was completely amorphous, while catalyst C was 100% crystalline, even though it was a mixture of anatase and brookite, which does not show any photocatalytic activity. Photocatalyst D was not completely crystalline, but it had a high ratio of anatase to rutile: the percentages of anatase and rutile were 57.8 and 12%, respectively. Photocatalyst E was 100% crystalline with a composition close to that of Degussa P25. Light scattering technique was used to determine the particle size of the colloidal suspension used to synthesize the immobilized photocatalysts. As shown in Table 5.2, the particle size varies from 0.07 to 5.50 µm. The optimum particle size is the result of competing impacts of the effective particle size on light absorption and scattering efficiency, charge-carrier dynamics, and surface area. Small-size particles have a larger surface area but a higher rate of electron and hole recombination at the particle surface because of the proximity of the charges. Large-size particles have slower electron/hole recombination rates, but they possess smaller surface area [100]. Figure 5.4 shows the UV-Vis spectra of photocatalysts A, B and C. As expected, all the TiO2 films have very low transmittance at wavelengths below 300 nm, indicating that they are able to absorb all the photons received at 254 nm, which is the main wavelength emitted by low-pressure UV lamps. Photocatalysts D and E showed a zero transmittance for the entire range of wavelengths. This is probably due to the scattering produced by the P25 powder incorporated in the composite coating and due to their thickness. 47

67 Figure 5.1: SEM micrograph of photocatalyst D. The micrograph shows important fractures on the surface of the catalyst. Table 5.1: Percentage of the 4 polymorphous in the 5 photocatalysts. Photocatalyst % Crystallinity Anatase % Rutile % Brookite A B C D E Table 5.2: Particle size of the 5 different sols. Colloid Particle size (micron) A B C D E

68 Figure 5.2: SEM micrograph of photocatalyst E. The micrograph shows an higher homogeneity compared to the one of catalyst D. Figure 5.3: XRD of the five different coatings. The figure shows hat catalyst B is amorphous and catalyst A, B, D and E are a a mixture of rutile and anatase. 49