Thesis Template

UV-ADVANCED OXIDATION TREATMENT OF MICROPOLLUTANTS IN SECONDARY WASTEWATERS

Jacque-Ann Natacia Grant

A thesis submitted in conformity with the requirements for the degree of Doctor of Philosophy Graduate Department of Civil Engineering University of Toronto

UV-ADVANCED OXIDATION TREATMENT OF MICROPOLLUTANTS IN SECONDARY WASTEWATERS Jacque-Ann Natacia Grant Doctor of Philosophy Graduate Department of Civil Engineering University of Toronto 2015

Abstract

Ultraviolet light+hydrogen peroxide (UV/H2O2) advanced oxidation is known to effectively oxidise micropollutants in wastewater. Its relatively high cost and energy requirements in comparison to other treatment options, however, have limited its implementation. This research evaluated the use of coagulation as a modification to the pre-tertiary component of the wastewater treatment process. The objective was to reduce the background scavenging capacity of the wastewater as a means of improving the oxidation efficiency of UV/H2O2 while reducing the cost and energy requirements.

Effluent organic matter (EfOM) was identified as the primary wastewater component responsible for scavenging of the hydroxyl radicals required for oxidation in UV/H2O2 treatment. The organic constituents of EfOM that are the driving force for its reactivity with the hydroxyl radical were identified as the high molecular weight components (biopolymers) and tryptophan-protein like components. Since EfOM concentration and composition can vary from one water matrix to another, a comparison of secondary wastewaters from membrane bioreactor (MBR) and activated sludge (AS) treatment systems showed that the average scavenging capacity of the AS systems exceeded the average of the MBRs such that MBRs may be more amenable to UV/H2O2 treatment.

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Coagulation of the wastewater using ferric chloride, aluminium sulphate, and polyaluminium chloride primarily removed the high molecular weight components and significantly reduced the

EfOM scavenging capacity. This improvement in wastewater quality also resulted in an improvement in the degradation rates of micropollutant compounds, reduced the energy requirements required to achieve 1-log removal of the compounds, and significantly reduced the costs associated with the UV/H2O2 system. No one coagulant outperformed any of the others; however the study demonstrated that coagulation is a feasible modification for a wastewater treatment plant upstream of an UV-AOP system. It was also found that the cost benefit to

UV/H2O2 exceeded the chemical costs of coagulation. Nevertheless, facilities considering UV-

AOP systems should weigh the benefits of increasing the H2O2 concentration or using coagulation as the cost savings are relatively similar at high H2O2 concentrations.

Acknowledgments

O taste and see that the Lord is good; blessed is the man that trusteth in Him. I thank Almighty God for His continual grace and strength throughout this project, for it is only in Him and through Him all things are possible.

This project has received funding support from the Government of Ontario. Such support does not indicate endorsement by the Government of Ontario of the contents of this material. Financial support was also provided by the Ontario Research Fund through the Centre for the Control of Emerging Contaminants.

I am deeply grateful to my thesis supervisor, Professor Ron Hofmann, for his support and guidance throughout the course of this work academically and professionally. I also thank Professor Robert Andrews and Hugh Monteith for being on my supervisory committee and the suggestions and feedback provided. I thank Sonya Kleywegt and David Poirier of the Ontario Ministry of Environment and Climate Change, Laura Meteer of the Regional Municipality of York, and Keith Bircher of Calgon Carbon Corporation for the support provided throughout this project. I am also grateful to Hydromantis Environmental Software Solutions Inc. (Hamilton, Ontario, Canada) for allowing the use of the proprietary CapDet Works software for the cost modelling analysis, to Dr. Monica Tudorancea and Dr. Sigrid Peldzsus (University of Waterloo) for performing LC-OCD analyses, to Dr. Viviane Yargeau and Rachel Benoit of Yargeau Laboratories (McGill University) for estrogenicity analyses, and Dr. Jim Bolton of Bolton Photosciences Inc. for the use of UVCalc 2B.

I would also like to express my gratitude to Jim Wang, Hong Zhang, Russell de Souza, Ding Wang, Anwar Sadmani, Iolanda Montagnese, Montaseer Rahman, Min (Talia) Xu, Jiafan Yang, Clare Lin, and Elena Li for their assistance in the lab, and to my DWRG colleagues for their general encouragement along the way.

Finally, I am grateful and thankful for my mom, Mary Elizabeth Grant, for her unwavering support, encouragement, and belief in me throughout this journey. v

Table of Contents

Acknowledgments ...... iv Table of Contents ...... v List of Tables ...... x List of Figures ...... xii Nomenclature ...... xiv

Chapter 1 - Introduction ...... 1 1.1 Background ...... 1 1.2 Thesis Format ...... 3 1.3 Objectives ...... 3

Chapter 2 - Literature Review ...... 5 2.1 Micropollutants in the Environment ...... 5 2.1.1 Classification ...... 5 2.1.2 Occurrence ...... 6 2.1.3 Potential Impacts ...... 7 2.1.4 Current Status of Regulations ...... 8 2.1.5 Treatment and Removal of Micropollutants ...... 9 2.2 Advanced Oxidation Processes ...... 10 2.2.1 Mechanism and Reaction Kinetics ...... 10 2.2.2 Limitations to AOPs ...... 13 2.2.3 Ecotoxicological Effects of UV-AOP ...... 13 2.3 Pretreatment Methods for EfOM ...... 15 2.4 Energy Requirements of AOP Processes ...... 17 2.5 References ...... 19

Chapter 3 - Materials and Methods ...... 29 3.1 Materials ...... 29 3.1.2 Selection of Wastewaters ...... 29 vi

3.1.3 Selection of Micropollutant Compounds ...... 29 3.1.4 Selection of Ecotoxicological Assessment Methods ...... 31 3.1.5 Selection of Coagulants ...... 32 3.2 Experimental Protocols ...... 33 3.2.1 Coagulation Experiments ...... 33

3.2.3 Advanced Oxidation (UV/H2O2) Experiments ...... 33 3.2.4 Background Scavenging Capacity Experiments ...... 35 3.2.5 Quality Control for Advanced Oxidation ...... 37 3.2.6 Sampling Bottles ...... 41 3.3 Analytical Methods ...... 42 3.3.1 Analysis of Micropollutant Compounds ...... 42 3.3.2 Anion Analysis ...... 49 3.3.3 Overall Background Scavenging Capacity ...... 49 3.3.4 Liquid Chromatography Organic Carbon Detection (LC-OCD) analysis ...... 54 3.3.5 Residual Hydrogen Peroxide ...... 55 3.3.6 Incident Irradiance of the LP Lamp using Iodide/Iodate Actinometry ...... 56 3.3.7 Incident Irradiance of the MP Lamp using Ferrioxalate Actinometry ...... 56 3.3.8 Fluorescence Excitation Emission Matrices Analysis ...... 57 3.3.9 Total Organic and Inorganic Carbon Analysis ...... 58 3.3.10 Alkalinity ...... 60 3.3.11 Acute Toxicity ...... 60 3.3.12 Genotoxicity ...... 61 3.3.13 Estrogenicity ...... 61 3.4 References ...... 63

Chapter 4 - A Comparative Study of the Hydroxyl Radical Scavenging Capacity of Secondary Wastewater Effluents ...... 69 Abstract ...... 69 4.1 Introduction ...... 70 4.2 Materials and Methods ...... 72 4.2.1 Wastewater Effluents ...... 72

4.2.2 Measurement of Scavenging Capacity and kOH,EfOM ...... 72 vii

4.2.3 Analytical Equipment and Methods ...... 76 4.2.4 Chemicals and Reagents ...... 77 4.3 Results and Discussion ...... 77 4.3.1 Assessing the Contributors to Overall Scavenging Capacity ...... 77 4.3.2 Comparing MBR and AS Scavenging Characteristics ...... 80

4.3.3 Identifying Components that Influence EfOM Scavenging and kEfOM ...... 82 4.4 Conclusions ...... 85 4.5 References ...... 87

Chapter 5 - Coagulation Pretreatment of Secondary Wastewater Effluent to

Improve UV/H2O2 Efficiency ...... 93 Abstract ...... 93 5.1 Introduction ...... 94 5.2 Materials and Methods ...... 95 5.2.1 Wastewater Samples ...... 95 5.2.2 Experimental Approach ...... 96 5.2.3 Advanced Oxidation Experiments ...... 97 5.2.4 Equipment and Methods ...... 99 5.2.5 Chemicals and Reagents ...... 100 5.3 Results and Discussion ...... 101 5.3.1 Effects on Wastewater Quality ...... 101 5.3.2 Effects on Micropollutant Degradation ...... 104 5.3.3 Effects on Electrical Energy Requirements ...... 110 5.3.4 Effects on Treatment Costs ...... 112 5.4 Conclusions ...... 114 5.5 References ...... 116

Chapter 6 - UV/H2O2 Treatment of Municipal Wastewater: Ecotoxicological Effects ...... 124 Abstract ...... 124 6.1 Introduction ...... 125 6.2 Materials and Methods ...... 127 6.2.1 Wastewater ...... 127 viii

6.2.2 Pretreatment ...... 127

6.2.3 Photolysis and UV/H2O2 Advanced Oxidation ...... 128 6.2.4 Analytical Methods - Bioassays ...... 129 6.2.5 Chemicals and Materials ...... 130 6.3 Results and Discussion ...... 130 6.3.1 Effects on Acute Toxicity ...... 130 6.3.2 Effects on Genotoxicity ...... 132 6.3.3 Effects on Estrogenicity ...... 137 6.4 Conclusions ...... 140 6.5 References ...... 141

Chapter 7 - UV/H2O2 Treatment of Secondary Wastewater Effluents – Effect on Effluent Organic Matter Characteristics ...... 149 Abstract ...... 149 7.1 Introduction ...... 150 7.2 Materials and Methods ...... 151 7.2.1 Wastewater Samples ...... 151

7.2.2 UV/H2O2 Advanced Oxidation Experiments ...... 152 7.2.3 Analytical Equipment and Methods ...... 153 7.3 Results and Discussion ...... 154 7.4 Conclusions ...... 160 7.5 References ...... 161

Chapter 8 - Conclusions and Recommendations for Future Research ...... 167 8.1 Conclusion...... 167 8.2 Research Contributions ...... 169 8.3 Recommendations for Future Research ...... 170

Appendix A: Endocrine Disruptors (EDCs) ...... 172 Appendix B: Pharmaceutical and Personal Care Products (PPCPs) ...... 173 Appendix C: Occurrence of some EDCs and PPCPs in wastewater effluents globally ...... 175 Appendix D: Anion calibration graphs ...... 177 ix

Appendix E: Micropollutant calibration graphs ...... 178 Appendix F: Effluent quality characteristics of the additional 23 plants sourced from the literature ...... 180 Appendix G: Correlation between EfOM scavenging capacity and DOC concentration for AS and MBR wastewaters ...... 181 Appendix H: Dose-response curves for the coagulants ...... 182 Appendix I: Micropollutant degradation plots ...... 183 Appendix J: Absorption spectra for wastewater samples and pharmaceutical compounds ...... 194 Appendix K: Loading plots and percentage variability of AS1 principal components ...... 196 Appendix L: Loading plots and percentage variability of AS2 principal components ...... 197

List of Tables

Table 3-1: Selected micropollutant compounds for research study ...... 30 Table 3-2: Photochemical characteristics of the selected compounds ...... 31 Table 3-3: Degradation rates of MB at different treatment scenarios ...... 37 Table 3-4: Methylene blue degradation rates in the presence of chloride anions ...... 40 Table 3-5: Methylene blue degradation rates in the presence of nitrate ...... 41 Table 3-6: Reagents for LC-MS/MS analysis of the micropollutants ...... 43 Table 3-7: Instrument operating conditions for LC-MS/MS analysis ...... 44 Table 3-8: Sample extraction protocol for LC-MS/MS Analysis ...... 45 Table 3-9: Surrogates and internal standards for LC-MS/MS analysis...... 48 Table 3-10: Preparation of standard solutions for micropollutant analysis ...... 48 Table 3-11: Anion Method Detection Limits at 1 mg/L spike concentration ...... 49 Table 3-12: Description of the LC-OCD components ...... 55 Table 3-13: List of reagents for hydrogen peroxide analysis ...... 55 Table 3-14: Sample volumes used for hydrogen peroxide analysis ...... 55 Table 4-1: Water quality of the wastewater effluents used in the study ...... 73 Table 4-2: Contribution of known scavengers to overall background scavenging capacity ...... 78 Table 4-3: Contribution of known scavengers in 33 secondary wastewater effluents ...... 79 Table 4-4: Scavenging characteristics of the secondary wastewater effluents ...... 81

Table 4-5: Spearman’s correlation coefficients for EfOM scavenging and kEfOM ...... 83 Table 5-1: Water quality characteristics of the secondary wastewater ...... 96 Table 5-2: Changes in wastewater quality with coagulation treatment ...... 103 Table 5-3: Percentage change in wastewater quality with coagulation treatment relative to no pretreatment ...... 103 Table 5-4: Photolytic degradation rate constants at different treatment conditions ...... 104

Table 5-5: Degradation rate constants with coagulation and UV + 10 mg/L H2O2 ...... 106

Table 5-6: Degradation rate constants with coagulation and UV + 20 mg/L H2O2 ...... 106 Table 5-7: EEO values for 1-log micropollutant degradation using UV photolysis alone ...... 111

Table 5-8: EEO values for 1-log micropollutant degradation using UV + 10 mg/L H2O2 ...... 111

Table 5-9: EEO values for 1-log micropollutant degradation with UV + 20 mg/L H2O2 ...... 112 Table 6-1: Water quality characteristics of the secondary wastewater ...... 128 xi

Table 6-2: 48h-LC50 values for Daphnia magna bioassay of the spiked wastewater ...... 131

Table 6-3: 96h-LC50 values for rainbow trout bioassay of the spiked wastewater ...... 132 Table 7-1: Characteristics of the secondary wastewater effluents used in the study ...... 152 Table 7-2: Classification of principal components for the secondary wastewater effluents ...... 154 Table 7-3: Reduction rates of the fraction volumes with UV alone ...... 157 Table 7-4: Fluorescence characteristics of tryptophan and tyrosine ...... 158

Table 7-5: Reduction rates (k) of the fraction volumes at UV + 10 mg/L H2O2 ...... 159

Table 7-6: Reduction rates (k) of the fraction volumes at UV + 20 mg/L H2O2 ...... 159

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List of Figures

Figure 3-1: Calgon Carbon Rayox® Advanced Oxidation Batch Pilot Reactor ...... 34 Figure 3-2: Calgon Carbon UV-collimated beam apparatus ...... 36 Figure 3-3: Calibration graph for methylene blue (MB) ...... 38 Figure 3-4: Methylene blue degradation in ultra-pure water in the Rayox at different

H2O2 concentrations ...... 39 Figure 3-5: Methylene degradation in the presence of chloride anions ...... 40 Figure 3-6: Methylene blue degradation with nitrate anions ...... 41

Figure 3-7: Typical degradation plot for methylene blue at different H2O2concentrations ...... 53

Figure 3-8: Typical plot of 1/ROH,UV versus 1/H2O2 ...... 53 Figure 3-9: Typical calibration graph for total organic carbon concentration ...... 59 Figure 3-10: Typical calibration graph for total inorganic carbon concentration ...... 59 Figure 3-11: Colour change in the 96-well plate for estrogenicity analysis ...... 62 Figure 3-12: Dose-response curve for estrogenicity analysis ...... 62 Figure 4-1: Distribution of the EfOM organic constituents in the wastewater effluents ...... 82 Figure 5-1: Calgon Carbon Advanced Oxidation Batch Reactor ...... 98 Figure 5-2: Change in organic composition with pretreatment of the wastewater ...... 102

Figure 5-3: Degradation rate for caffeine with coagulation and UV/H2O2 ...... 108

Figure 5-4: Degradation rate for carbamazepine with coagulation and UV/H2O2 ...... 108

Figure 5-5: Cost of acheving 1-log carbamazepine degradation with a UV/H2O2 system at different H2O2 concentrations and % UVT ...... 113 Figure 6-1: SOS Chromotest IF values for the spiked wastewater treated with UV photolysis (3200 mJ/cm2) alone ...... 134 Figure 6-2: SOS Chromotest IF values of the spiked wastewater with

UV/H2O2 treatment ...... 134 Figure 6-3: Genotoxicity (4-NQO TEQ) of the spiked wastewater with UV photolysis (3200 mJ/cm2) alone ...... 135 Figure 6-4: Genotoxicity (4-NQO TEQ) of the spiked wastewater with

UV/H2O2 treatment ...... 136 xiii

Figure 6-5: Change in relative estrogenicity of the spiked wastewater at

0 mg/L H2O2 with UV exposure time and coagulation ...... 138 Figure 6-6: Change in relative estrogenicity of the spiked wastewater at

10 mg/L H2O2 with UV exposure time and coagulation ...... 139 Figure 6-7: Change in relative estrogenicity of the spiked wastewater at 20 mg/L with UV exposure time and coagulation ...... 139 Figure 7-1: Reduction rates of the FRI fractions with UV photolysis alone ...... 155

Figure 7-2: Reduction rates of the FRI fractions with UV + 10 mg/L H2O2 treatment ...... 155

Figure 7-3: Reduction rates of the FRI fractions with UV + 20 mg/L H2O2 treatment ...... 156 xiv

Nomenclature

% Percent

p Quantum yield of the pollutant compound

Hydrogen peroxide molar absorption coefficient oC Degree Celsius 4-NQO 4-Nitroquinoline 1-oxide Al Aluminium AS Activated sludge ANOVA Analysis of variance AOP Advanced oxidation process APHA American Public Health Association CBZ Carbamazepine CCME Canadian Council of Ministers of the Environment

CH3OH Methanol Cl- Chloride

Carbonate DOC Dissolved organic carbon Es Einstein

Ea Average fluence rate E2 17β-estradiol EE2 17α-ethinylestradiol EC European Commission

EC50 Effective median concentration EDC Endocrine disrupting compound

EDTA-Na2 Ethylenediaminetetraacetic acid disodium salt dihydrate

EEM Electrical energy per mass EEO Electrical energy per order EfOM Effluent organic matter EU European Union Ex/Em Excitation-emission xv

fp Ratio of light absorbed by the pollutant (P) to that absorbed by other components Fe Iron

FeCl3 Ferric chloride FEEM Fluorescence excitation emission matrices FRI Fluorescence regional integration

H2O2 Hydrogen peroxide

H2SO4 Sulphuric acid

Bicarbonate HFBA Heptafluorobutyric acid Hg Mercury HLB Hydrophilic-lipophilic balance HOCl Hypochlorous acid HSD Honestly significant difference GAC Granular activated carbon

Incident flux of the radiation ID Internal diameter IF Induction factor

∑ks[S] Overall background scavenging capacity

Observed degradation rate constant due to UV and H2O2

Degradation rate constant due to direct UV photolysis only kOH,EfOM Reaction rate constant of effluent organic matter kDa Kilo-dalton kOH Hydroxyl radical reaction rate constant kW Kilowatt kWh/m3/order Kilowatt hour per cubic metre per order L Litre

LC50 Median lethal concentration LC-MS/MS Liquid chromatography tandem mass spectrometry LC-OCD Liquid chromatography organic carbon detection log Kow Octanol water partition coefficient LMW Low molecular weight xvi

LP Low-pressure M-1s-1 Per mole(s) per second -1 -1 Mc s Per mole of carbon per second MB Methylene blue

MBo Initial methylene blue concentration MBR Membrane bioreactor MDL Method detection limit mg/L Milligram per litre mL Milli-litre mJ/cm2 Millijoules per square centimetre MOECC Ministry of the Environment and Climate Change mol/Es moles per Einstein MP Medium-pressure MW Molecular weight mW/cm2 Milliwatt per square centimetre MQ MilliQ NaCl Sodium chloride NaOH Sodium hydroxide NF Nanofiltration

. Nitrate anion NOM Natural organic matter NPnEC Nonylphenol carboxylates NPnEO Nonylphenol ethoxylates ng L-1 Nanogram per litre

O3 Ozone OH Hydroxyl radical • [ OH]ss Steady-state concentration of hydroxyl radicals ON Ontario PACl Polyaluminium chloride pCBA Parachlorobenzoic acid PCA Principal component analysis PhACs Pharmaceutically active compounds xvii

PMT Photomultiplier voltage PODR Point-of-diminishing returns PPCP Pharmaceutical and personal care products  ROH,UV OH exposure per UV fluence REP Relative estrogenic potency r(P) Rate of oxidation of the pollutant SUVA Specific UV absorbance at 254 nm SRT Sludge retention time SMZ Sulphamethoxazole SPE Solid phase extraction TEQ Toxicity equivalent unit TOC Total organic carbon TIC Total inorganic carbon µg L-1 Microgram per litre UF Ultrafiltration USEPA United States Environmental Protection Agency USFDA United States Food and Drug Administration

UV254 UV absorbance at 254 nm UV/VIS Ultraviolet light/visible light WHO World Health Organisation WWTP Wastewater treatment plant YES Yeast estrogenicity screening

Chapter 1 Introduction

1.1 Background

Advances in analytical methodologies for the analysis of trace concentrations (ng/L) of organic compounds have led to the identification of a wide range of emerging contaminants/micropollutants in the aquatic environment. These micropollutants, which may include endocrine disrupting compounds (EDCs) and pharmaceuticals and personal care products (PPCPs), originate mainly from anthropogenic activities, and conclusive findings of the deleterious effects on aquatic organisms have been the impetus for investigating suitable treatment options for their removal. The findings have also spurred concerns for the potential impacts on human health. The primary point source for these compounds in the environment is municipal wastewater treatment plants which are not specifically designed for treating these compounds (Mompelat et al., 2009; Radjenovic et al., 2007). In the absence of regulations regarding the discharge of these compounds to the environment and in light of on-going concerns, research has been focused on a number of key areas. These areas include the identification of pollutants, understanding reaction pathways and by-products, identifying potential toxicological effects on the environment and organisms, mechanisms for removal, and treatment technologies that can be effectively used for their removal.

Advanced oxidation process (AOP) is one technology that has been shown to be very effective for treating micropollutants, but its use in the industry, particularly in wastewater treatment, has been limited due to its relatively high cost and energy requirements. AOPs involve the generation of hydroxyl radicals (•OH), and have been shown to achieve more than 90% degradation of some micropollutants (Andreozzi et al., 2003 and 2004; Rosenfeldt et al., 2004; Vogna et al., 2004; Yuan et al., 2009). It is also of particular interest since it does not entail the partitioning of the pollutants from one phase to another and no additional waste is generated that requires further treatment as is typical for other treatment options. Studies have focused on evaluating the efficiency with which certain classes of micropollutants can be removed using different oxidation processes, identification of the oxidation kinetics and rate constants for certain pollutants, as well as the identification of the specific oxidation by-products and

2 transformation products. However, a major limitation to its effectiveness is the presence of hydroxyl scavengers which impair the oxidation efficiency of target compounds by reducing the concentration of •OH available for oxidising these compounds. Low •OH concentrations will increase the AOP dose and energy required for effective treatment of a water matrix. These scavengers in wastewaters (mainly effluent organic matter (EfOM), carbonates, bicarbonates and nitrite) exert an oxidant demand that slows degradation rates, and increases energy and chemical requirements and associated costs. Given that EfOM is considered to be the major hydroxyl radical scavenger (Dong et al., 2010), reducing its concentration in wastewaters prior to the application of an AOP could improve oxidation efficiency.

EfOM is complex in nature, and its characteristics, composition, and reactivity with the •OH radical could vary from one matrix to another with possibly similar implications on the effectiveness of AOP treatment. Furthermore, EfOM reactivity may be site-specific and the effluent from one type of treatment system, or with certain characteristics or composition may be more amenable to AOP treatment due to a lower background scavenging capacity. Hence, reducing the EfOM may reduce the scavenging potential of the matrix. Organic matter removal is typically achieved in the potable water industry using coagulation and activated carbon adsorption; as such, it is feasible that applying these processes to wastewaters could reduce the effluent organic matter (EfOM) concentration, overall background scavenging capacity, and the overall oxidant demand. This approach of pretreatment of the wastewater prior to AOP could improve the efficiency of using AOPs in terms of dose and energy requirements. However, the impact of these pretreatment methods on the oxidation efficiency of advanced oxidation processes is an area that has not been evaluated for UV/H2O2 and requires further study. This is important as although AOPs such as UV/H2O2 have been shown to be effective for micropollutant removal, the relatively high energy and cost requirements compared to other treatment are usually prohibitive, therefore a practical approach of addressing this concern would benefit the wastewater treatment industry.

1.2 Thesis Format

This thesis is presented as a compilation of papers which focused on different aspects of the overall objective. A general literature review is provided in Chapter 2. An overview of the materials and methods used for all experiments is in Chapter 3. The papers are in Chapter 4 – 7 with overall conclusions and recommendations for future research in Chapter 8.

1.3 Objectives

The overall objective of this thesis was to examine modifications to the wastewater treatment process upstream of an AOP system that would optimise the efficiency and effectiveness of using UV/H2O2 AOP for treating micropollutants in secondary wastewater effluents. The specific objectives of this thesis were to:

1. Compare and evaluate the water quality and scavenging capacity characteristics of secondary wastewaters to identify components that contribute to the background scavenging capacity of the matrices. Effluent organic matter (EfOM) is considered to be the largest contributor to the scavenging potential in all wastewater matrices irrespective of the treatment process. However, its reactivity with the OH radical and its scavenging potential will vary from one effluent to another such that one or more components or characteristics of EfOM may be responsible for its scavenging capability. Chapter 4 compares the scavenging capacity of secondary wastewaters from activated sludge (AS) and membrane bioreactor (MBR) treatment systems. The purpose of this comparison is to assess whether the wastewater from one type of treatment may be more amenable to AOP treatment versus the other. Hydroxyl radical reaction rate constants for the bulk EfOM and the EfOM constituent that is most likely responsible for EfOM scavenging capacity are identified.

2. Evaluate the impact of EfOM removal on the efficiency of UV/H2O2 treatment of micropollutants in secondary wastewaters. Pretreatment of the secondary effluent prior to the application of an AOP will reduce EfOM concentration. This is expected to also reduce the overall background scavenging capacity of the effluent, thereby improving degradation

of the micropollutants while reducing the energy required for treating these compounds. Chapter 5 evaluates the impact of reducing EfOM concentration using coagulation as a

pretreatment method prior to UV/H2O2, in terms of degradation rates of target micropollutant compounds, and on AOP energy requirements and costs.

3. Assess the impact of UV/H2O2 treatment on the potential ecotoxicological effects of the micropollutants in secondary wastewaters. Advanced oxidation treatment can effectively reduce potential toxicological effects in the wastewater due to the removal of contaminants, but it is also possible that toxicity may increase with treatment due to oxidation by-products, transformation products or daughter compounds. Chapter 6 assesses the effect

of UV/H2O2 treatment on the acute toxicity, genotoxicity and estrogenicity of wastewater spiked with target micropollutants.

4. Evaluate the changes that occur in the EfOM characteristics during UV/H2O2 treatment.

During oxidation with UV/H2O2, changes in EfOM characteristics and composition could influence its scavenging capabilities. Given the complex nature of EfOM, one or more organic components may be significantly affected during oxidation. Chapter 7 evaluates the

changes that occur in the organic constituents of EfOM when treated with UV/H2O2 at a range of hydrogen peroxide concentrations.

Chapter 2 Literature Review

2.1 Micropollutants in the Environment

2.1.1 Classification

Trace natural and synthetic organic compounds, also referred to as micropollutants, have been identified in surface waters and groundwaters, and in effluent streams of wastewater treatment plants. These compounds, which can include endocrine disrupting compounds (EDCs) and pharmaceuticals and personal care products (PPCPs), have concentrations in the ng L-1 to µg L-1 range (Synder et al., 2003; Ikehata et al., 2006).

EDCs are exogenous substances that can have deleterious effects on the functioning of the endocrine system, affecting growth, reproduction and development of organisms (Scruggs et al., 2004; Esplugas et al., 2007; Bolong et al., 2009). These include pesticides, organochlorines and organohalogens, heavy metals, industrial chemicals, phytoestrogens, as well as synthetic and natural hormones or steroids (Falconer et al., 2006; Lishman et al., 2006; Esplugas et al., 2007). Some EDCs are commonly used in packaging materials, paints, flame retardants in electronics, textiles and furniture (Caliman et al., 2009). PPCPs refer to analgesics and antiflammatory drugs, antiepileptic drugs, beta-blockers, blood lipid regulators, cytostatic drugs, oral contraceptives, antiseptics, surfactants/detergents, musk fragrances, and sunscreen agents (Ikehata et al., 2008; Bolong et al., 2009; Caliman et al., 2009). A number of these compounds can be further classified into sub-groups. These subgroups are shown in Appendix A and Appendix B, which outlines the compounds that have been frequently detected in wastewater treatment plants and to a lesser extent, surface waters. Notably, some pharmaceuticals and personal care products can also be regarded as endocrine disruptors based on the compound’s potential to affect the functions of the endocrine system (Scruggs et al., 2004).

2.1.2 Occurrence

EDCs and PPCPs primarily originate from residential, industrial and agricultural activities and direct and indirect releases of the compounds are responsible for their presence in the environment. Direct releases or point sources occur from effluent discharges of wastewater treatment facilities (industrial and/or municipal), and effluents from farms and landfills. Indirect or non-point sources for these compounds are runoff from agricultural and livestock areas, paved surfaces, recreational areas such as golf courses and parks, and residential gardens (Kemper 2008; Bolong et al., 2009; Mompelat et al., 2009). These indirect sources are mainly responsible for the presence of pesticides and veterinary-based compounds.

Municipal wastewater treatment plants (WWTPs) are the main collection point for micropollutants and the primary point source for those compounds present in the aquatic environment (Daughton et al., 2001; Mompelat et al., 2009; Phillips et al., 2007). Pharmaceutical products, such as antibiotics, lipid regulators, contraceptives, and veterinary medications, are ingested by humans and animals, and are absorbed, metabolised and excreted as waste (Ikehata et al., 2006; Mompelat et al., 2009). The original compound and metabolites from the human body are constituents of the influent to wastewater treatment plants (Radjenovic et al., 2007). Household practices of disposing unused or expired medications via toilets and/or sinks are another contributing factor to the presence of these compounds in the influent streams (Kotchen et al., 2009). Micropollutants that have been identified in the effluent streams of wastewater treatment facilities in North and Central America, and Europe are shown in Appendix C.

Aerobic biological treatment systems, typical of municipal WWTPs, are considered to be a cost- effective and relatively efficient approach for treating micropollutants (Klatt et al., 2003; Scruggs et al., 2004). However, current design and operation of these WWTPs do not facilitate complete mineralisation and removal of the micropollutants (Mompelat et al., 2009; Radjenovic et al., 2007; Scruggs et al., 2004). As such, effluent streams typically contain traces of the parent compounds, metabolites from the human body, as well as metabolites and transformation products from the biological degradation process (Bolong et al., 2009).

2.1.3 Potential Impacts

Micropollutants in aquatic systems, and the persistence and bioaccumulative capabilities of some compounds (e.g., pesticides and some pharmaceuticals) are sources of concern for potential adverse environmental and human health impacts (Bolong et al., 2009; Ikehata et al., 2008; Mompelat et al., 2009). Limited information on the toxicity effects of these compounds and the actual levels of risks that are posed to human health are other contributory factors. Of the classes/categories of micropollutants identified, pharmaceutical compounds and their associated metabolites, endocrine disrupting compounds, and surfactants are of particular interest due to the increasing variety, usage, and complexity of new pharmaceutical products being manufactured to address new diseases, impacts on aquatic organisms by endocrine compounds, and the toxic effects and persistence of surfactants (Ikehata et al., 2004; Sanchez- Polo et al., 2008).

A number of studies have conclusively demonstrated that some micropollutants can have adverse effects on aquatic organisms. Reported impacts on fish species are reduced fertility, masculinisation of female fish, feminization of male fish, and alteration in spawning patterns (Kramer et al., 1998; Batty and Lim, 1999; Esplugas et al., 2007). Purdom et al. (1994), Folmar et al. (1996) and Harries et al. (1996) reported abnormalities of increased levels of vitellogenin in males, and changes in sex steroids of fishes that were living immediately downstream of the final discharge point for wastewater treatment plants. These observed changes were mainly due to estrogenic compounds (17β-estradiol (E2) – a human hormone, and 17α-ethinyl estradiol (EE2) – a synthetic birth control) whose concentrations typically found in the wastewaters were sufficient to result in these impacts (Kramer et al., 1998; Panter et al., 1998; Snyder et al., 2001). Trace concentrations of pesticides were also deemed the cause of multiple limbs or missing limbs in amphibian organisms (Hayes et al., 2002).

Despite the observed ecological impacts, there is no conclusive evidence to date that micropollutants will have adverse effects on human health (WHO/IPCS 2002). It is hypothesised that lower sperm counts and increased occurrences of testicular and prostate cancer are due to the consumption of EDCs in potable water supplies. However, trace concentrations of these compounds and the differences in exposure levels for humans and

8 aquatic organisms are considered inadequate to result in any adverse effects on human health. While aquatic organisms are subject to continuous exposure to these compounds, humans ingest considerably smaller concentrations through the consumption of potable water (Snyder et al., 2003). Direct human health effects of these compounds, as well as pharmaceuticals, have only occurred in instances of accidental exposure or dosing (WHO/IPCS 2002).

2.1.4 Current Status of Regulations

There are no regulations regarding EDCs, PPCPs, or associated by-products for the wastewater and water industry worldwide. Current legislation regarding EDCs focuses on regulating the industries that manufactures different compounds, and does not address wastewater effluent discharges or potable water supplies (Snyder et al., 2003). Nevertheless, some initiatives have been made with endocrine disruptors.

In 1996, the United States (U.S.) established an Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC) to develop analytical methods for EDCs and to regulate these chemicals. A two-tiered system for identification, screening, testing, and hazard assessment of these compounds was developed and an amendment was made to the United States Safe Drinking Water Act and the Food Quality Protection Act (Caliman et al., 2009). This amendment mandated screening for endocrine activity in substances that could contaminate drinking water or food supplies (Snyder et al., 2003). In 1998, an Endocrine Disruptor Screening Program (EDSP) was established by the United States Environmental Protection Agency (USEPA) and uses a two-tiered approach for identifying possible endocrine disruptors and the potential impacts (Falconer et al., 2006). In 2001, an Endocrine Disruptor Methods Validation Subcommittee was established to examine validation methods for the analysis of these compounds, with an aim of standardising the analytical protocols (Snyder et al., 2003). Currently, only in situations where it is anticipated that the concentration of a pharmaceutical in water may exceed an environmental concentration of 1 µg L-1, is ecological testing and evaluation required by the United States Food and Drug Administration (US FDA) (Snyder et al., 2003; Bolong et al., 2009). In the State of California in the U.S., monitoring programs have been established for EDCs and PPCPs where water reuse programs are established.

The European Union (EU) has established an EDC Priority List of 564 chemicals as part of the EU Strategy for Endocrine Disruptors adopted in 1999. In 2007, the European Commission implemented Regulation (EC) No. 1907/2006 which addresses industrial chemicals with endocrine disrupting potential. Naturally occurring EDCs are addressed in the European Union Water Framework Directive for surface waters issued in 2000. In Canada, the Canadian Council of Ministers of the Environment (CCME) Wastewater Systems Effluent Regulations establishes effluent quality standards as well as requirements for monitoring water quality and environmental effects. In the absence of regulations, three micropollutants (17- ethinylestradiol, 17-estradiol, and estrone) which are known endocrine disruptors, are among the list of compounds that are required to be measured when water quality monitoring studies are to be conducted by a wastewater plant when the condition stipulated in Section 14(1) of the Regulations is applicable.

Ongoing initiatives in Canada, the United States, Australia and elsewhere through research programs, aim to improve our understanding of the potential effects of these compounds in the environment and on human health. Organizations such as Environment Canada, the U.S. Environmental Protection Agency, and the World Health Organization (WHO) have also focused their efforts in collecting comprehensive information on analytical methodologies, occurrence and environmental fate, and the response of these compounds to different treatment strategies.

2.1.5 Treatment and Removal of Micropollutants

Aerobic biological treatment systems such as activated sludge (AS) and membrane bioreactors (MBR), membrane filtration, activated carbon adsorption and advanced oxidation processes (AOPs) are effective processes for treating micropollutants (Ikehata et al., 2008; Klatt et al., 2003; Radjenovic et al., 2007; Wintgens et al., 2002; Yoon et al., 2007). With biological systems, micropollutant removal (> 80%) is achieved by (1) Degradation resulting in biological and chemical transformations; (2) Adsorption on the activated sludge; and (3) Volatilization or air stripping (Schafer et al., 2002; Clara et al., 2005; Joss et al., 2005; Radjenovic et al., 2007; Suarez et al., 2008). However, removal efficiency is influenced by the type of treatment process,

10 solids retention time (SRT), seasonal variations, temperature, sunlight intensity, hydraulic retention time (HRT), location of the treatment system, and properties of the micropollutants (Nakada et al., 2006; Radjenovic et al., 2007; Caliman et al., 2009). As such, incomplete removal can occur where some compounds undergo transformation forming degradation/transformation products (Gonzalez et al., 2007).

Membrane filtration and activated carbon adsorption involve the mass transfer of the micropollutants to another phase (e.g., the retentate from membrane treatment and adsorption onto the carbon surface). Ultrafiltration (UF) and nanofiltration (NF) membranes are typically used in drinking water systems, while UF membranes are more common for wastewater treatment. Yoon et al. (2007) reported recoveries of greater than 75% for compounds with log

Kow < 2.8 for both UF and NF membranes where the removal mechanisms are usually hydrophobic adsorption, size exclusion, and charge exclusion such that hydrophobic micropollutants are retained. In terms of activated carbon, Westerhoff et al. (2005) reported removals of 20-69% in water for compounds such as acetaminophen, caffeine, diclofenac, naproxen, sulfamethoxazole, and atrazine using powdered activated carbon (PAC) while Snyder et al. (2007) reported 30-78% removals for pharmaceutical compounds in river water also using PAC. Li et al. (2011) reported removals of 82% and 92% for sulphamethoxazole and carbamezapine, respectively, with powdered activated carbon dose of 1.0 g/L in an MBR. However, both membrane filtration and activated carbon adsorption usually requires additional treatment of the retentate or the carbon for complete removal of the micropollutants. In contrast, advanced oxidation processes (AOP) do not require subsequent treatment steps and are also capable of achieving complete mineralisation of organic compounds, albeit at very high AOP doses.

2.2 Advanced Oxidation Processes

2.2.1 Mechanism and Reaction Kinetics

Advanced oxidation processes (AOPs) entail the in-situ generation of the hydroxyl radical (OH) for oxidising target compounds. Radical and oxygen species such as superoxide radical anions ( ), and hydroperoxyl radicals ( ) are also involved in the overall reactions;

11 however, the hydroxyl radical is the primary oxidant (Ikehata et al., 2006). The radical, which has an oxidation potential of 2.80V and a relative oxidising power of 2.05 compared with that of 1.0 for chlorine, 1.10 for hypochlorous acid, 1.31 for hydrogen peroxide and 1.52 for ozone, is a very strong oxidant (Munter 2001). The non-selectivity of the hydroxyl radical and its high rates of reaction with organic compounds in the range of 106 to 1010 M-1s-1 are responsible for the oxidation, transformation, or mineralisation of a broad range of chemical compounds (Esplugas et al., 2007; Gultekin et al., 2007; Haag et al., 1992).

Typical AOPs include ultraviolet light (UV)-based or ozone (O3)-based AOPs such as using hydrogen peroxide with UV (UV/H2O2) or ozone (O3/H2O2), ozone and UV(O3/UV), chlorine 2+ and UV(HOCl/UV), as well as the Fenton’s reagent (Fe /H2O2) and photocatalysis using titanium dioxide and UV (TiO2/UV). Using UV for final disinfection is a common practice in wastewater treatment plants. As such, a UV-based AOP would fulfil dual objectives of disinfection and micropollutant removal in the final wastewater effluent. Additionally, in Ontario, UV is one alternative for wastewater disinfection due to strict regulations for total residual chlorine in the treated effluent of less than or equal to 0.02 mg/L. Therefore, UV-based

AOPs such as UV/H2O2 would be a practical choice when considering AOP systems for wastewater treatment facilities, particularly for those already using UV disinfection.

For UV/H2O2, hydrogen peroxide undergoes direct photolysis resulting in the cleavage of the O- O bond yielding an OH in accordance with Equations 2.1 – 2.3 (Gottschalk et al., 2000; Caliman et al., 2009). The rate of photolytic degradation is dependent on the energy of the incident radiation, molar absorptivity, and quantum yield.

→ (2.1)

→ (2.2)

(2.3)

Typical hydroxyl radical reactions occur via hydrogen abstraction, radical-radical interactions, electrophilic addition, and electron transfer reactions (Ikehata et al., 2006). Initial reactions with

12 the hydroxyl radical are influenced by the nature of the organic species, and usually entail the abstraction of a hydrogen atom from water or addition of the hydroxyl radical to the pollutant (Munter, 2001). Hydrogen abstraction is typical for reactions involving alkanes or alcohol, while the addition reaction is characteristic of reaction with olefins or aromatic compounds (Munter, 2001). Ideally, the hydroxyl radical is capable of complete oxidation and mineralisation of organic molecules; however, in some instances complete mineralisation is not achieved as impractically long UV irradiation times and energy would be required. This leads to intermediate products or oxidation by-products with higher polarity and solubility (Esplugas et al., 2007; Ikehata et al., 2006; Klavarioti et al., 2009).

UV-AOP (UV/H2O2) oxidation of target compounds is achieved via direct photolysis and reactions with the hydroxyl radical where the rate of reaction can be expressed with Equation (2.4) (Parsons et al., 2004):

[ ] ( )  ( ( )) [ ][ ] (2.4)

where r(P) – rate of oxidation of the pollutant, – incident flux of the radiation, p – quantum yield of the pollutant compound (P), fp – ratio of light absorbed by the pollutant (P) to that absorbed by other components in the solution, and At – total absorbance of the solution.

Studies have shown that UV/H2O2 can achieve more than 90% removal of some micropollutant compounds such as 17α-ethinylestradiol, bisphenol A, and estradiol (Rosenfeldt et al., 2007; Zhang et al., 2010), diclofenac (Vogna et al., 2004), and clofibric acid (Andreozzi et al., 2003). For others (e.g., naproxen, carbamazepine, norfloxacin, caffeine, ibuprofen), removal efficiencies can range between 5-90% (Pereira et al., 2007; Wols et al., 2012). These variations in removal efficiencies are influenced by the chemical structure and photochemical characteristics of the specific target compound, and the quality of the water matrix (Petrovic et al., 2003; Wols et al., 2012). Most studies are performed in ultra-pure water and synthetic wastewater, but it is expected that the removal efficiencies will be lower in natural surface and wastewater matrices due to matrix effects.

2.2.2 Limitations to AOPs

The presence of hydroxyl radical scavengers is a major limitation to the efficiency of AOPs. These scavengers are organic and inorganic constituents that react with the OH radical reducing  the steady-state concentration of hydroxyl radicals ( OHss) available to react with the target compound. The oxidant demand exerted by these constituents requires larger treatment doses to achieve a desired level of degradation for the target compounds, thereby increasing energy and cost requirements for an AOP process (Dantas et al., 2012; Ikehata et al., 2006; Snyder et al., 2007). Typical hydroxyl scavengers in wastewaters are effluent organic matter (EfOM) and inorganic ions such as carbonate, bicarbonate, and nitrite (Keen et al., 2014; Rosario-Ortiz et al., 2010; Rosenfeldt et al., 2007). Of these, the major constituent of concern is the EfOM which is reported to contribute up to 85% of the overall background scavenging capacity of wastewater matrices (Keen et al., 2014; Souza et al., 2014). Given the scavenging contribution of EfOM, reducing EfOM concentration should improve the steady-state concentration of OH available for oxidation in the water matrix. However, the extent to which this may improve UV/H2O2 efficiency and the underlying factors of EfOM that influences its reactions with the hydroxyl radical have not been previously investigated.

It is also understood that the biodegradability, chemical structures, spectral characteristics, and polarity of dissolved organic matter can be altered on exposure to UV/H2O2 (Sarathy et al., 2013 and 2011). Gonzalez et al. (2013) has also shown that EfOM undergoes changes in its composition when exposed to UV/H2O2 treatment, with a reduction in large molecular weight components such as biopolymers and humics with an increase in UV fluence. These changes may alter the scavenging capabilities of EfOM and it is probable that the accuracy of the predicting the oxidation efficiency of micropollutants in wastewaters could be improved by understanding the dynamics and reaction kinetics of EfOM with the OH and the extent to which this may or may not vary from one wastewater to another.

2.2.3 Ecotoxicological Effects of UV-AOP

Micropollutant compounds can affect ecological systems through disruption in biological activity with respect to estrogenic activity, increased toxicity, genotoxicity, and antibacterial

14 activity. In biological degradation, pharmaceutical compounds are metabolised to a more hydrophilic form or adsorbed onto the sludge (Radjenovic et al., 2007; Esplugas et al., 2007; Suarez et al., 2008). Estrogenic compounds (both natural and synthetic) are excreted in an inactive conjugated form of sulphuric and glucuronic acid (Isidori et al., 2007), which are polar and water soluble (Schafer et al., 2002). In biological treatment, these compounds are re- transformed by microorganisms such Escherichia coli to an active state which is hydrophobic and unconjugated (Schafer et al., 2002). Transformation products can increase the potential for ecotoxicological effects on aquatic organisms because of probable toxicity effects. Differences in treatment processes could also contribute to variations in the type of transformation products formed as illustrated by Gonzalez et al. (2007) and Terzic et al. (2005) who noted that NPnEO and NPnEC transformation products of nonylphenol differed between AS and MBR effluents. The MBR wastewater generally had higher concentrations of the NPnEO products whereas the concentration of NPnECs, which are more toxic, was greater in the AS.

Since UV-AOPs may not necessarily result in the complete mineralization of target parent compounds or the transformation products from biological degradation, ultimately, oxidation by-products and transformation products can be formed with potential for more or less of an ecotoxicological impact (Fatta-Kassinos et al., 2011). Only a few studies have specifically examined the impact of UV/H2O2 treatment on potential ecotoxicological effects of wastewater. Reported studies have evaluated estrogenicity in surface water (Rosenfeldt et al., 2007), toxicity in municipal effluent (Andreozzi et al., 2004; Souza et al., 2013; Yuan et al., 2011), toxicity in an industrial textile effluent (Arslan et al., 2009), and genotoxicity and estrogenicity in secondary wastewater spiked with either bisphenol A, ciprofloxacin, metoprolol or sulphamethoxazole (Richard et al., 2014). In general, these studies noted that estrogenicity, acute toxicity, and genotoxicity effects were reduced with UV/H2O2 treatment, although Yuan et al. (2011) and Souza et al. (2013) reported an increase in toxicity with UV photolysis at a UV fluence of 3816 mJ/cm2.

2.3 Pretreatment Methods for EfOM

As discussed in Section 2.2.2, reducing the EfOM concentration in wastewater is expected to reduce the overall background scavenging capacity of a matrix. EfOM concentration is typically determined by measuring the dissolved organic carbon concentration in a wastewater. Coagulation and activated carbon adsorption are two effective methods for reducing the concentration of organic matter in water matrices. Optimising coagulation involves the removal of dissolved organic matter, in addition to suspended solids and colloidal material (USEPA 1999; Volk et al., 2000). Coagulants can be classified into two main groups of inorganic compounds (such as metal salts) and organic polymers (e.g., polyelectrolytes) and some commonly used inorganic coagulants are aluminum sulphate, iron salts and polyaluminum chloride (Delgado et al., 2003; Shon et al., 2006; Yan et al., 2008). The effectiveness of each coagulant is influenced by the speciation of the specific coagulant, and the efficiency with which dissolved organic matter is removed in the water matrix is dependent on the water quality characteristics (Delgado et al., 2003; Shon et al., 2006). This is evident with a study conducted by Wert et al. (2011) reporting a 10-47% decrease in EfOM concentration using ferric chloride at 10-30 mg Fe/L for three wastewaters while Xue et al. (2015) reported EfOM removals of 9- 71% for two wastewaters using 41 mg Fe/L of ferric chloride.

In wastewaters, EfOM is typically comprised of organic macromolecules and other organic compounds characterised by ranges of molecular weight and particle sizes. Coagulation using Al-based or Fe-based coagulants primarily removes the high molecular fractions such as biopolymers and humic substances (Haberkamp et al., 2007; Matilainen et al., 2010; Shon et al., 2006; Volk et al., 2000). These components also exert a coagulant demand such that wastewaters with a high concentration of dissolved organic matter would require higher coagulant doses for effective removal (Narkis et al., 1997; Delgado et al., 2003). Some removal of EDCs and PPCPs could be obtained by adsorption of the pollutants onto the particles in the water being treated, as well as the flocs formed in the flocculation process (Westerhoff et al., 2005). However, removal is limited to the organic compounds that exhibit hydrophobic properties and are generally non-polar. Given that most EDCs and PPCP compounds are polar

with a log Kow values < 3, most of these compounds are not effectively removed (Shon et al., 2006).

Activated carbon in powdered or granular form can be used to remove organic matter thereby reducing EfOM concentrations (Shon et al., 2006). It is also found to provide an added benefit of removing compounds such as pesticides, pharmaceuticals, and estrogens found in wastewater effluents and drinking water supplies (Ternes et al., 2002; Snyder et al., 2003; Snyder et al., 2007). Removal efficiency is directly influenced by the physical and chemical properties of the carbon and the compounds that will be removed. Typical carbon characteristics include the surface area, pore size distribution, surface charge and oxygen content of the carbon. In terms of the micropollutants, removals are based on hydrophobic interactions and are influenced by the shape, size, charge, and hydrophobicity of the target compound (Snyder et al., 2003). Since activated carbons are capable of removing organic matter and organic compounds present in the matrix, both constituents compete for adsorption sites on the carbon surface (Wu et al., 2001; Yu et al., 2012; Zietzschmann et al., 2014). Despite this, activated carbon was found to remove primarily the low molecular weight components of the dissolved organic matter with minimal removal of high molecular weight constituents such as biopolymers (Filloux et al., 2012; Velten et al., 2011; Yu et al., 2012). It is suggested that limited removal of biopolymers occurs as these compounds are very hydrophilic and their size limits penetration into the pores of the powdered activated carbon (Newcombe 1999). Hydrophilic compounds are poorly removed and pH also plays a significant factor such that there is a preference for the adsorption of neutral species (Shon et al., 2006).

In terms of micropollutant removal using activated carbon, activated carbons have been found to be very efficient for the removal of non-polar organic compounds (i.e., compounds with log Kow > 2.0), as well as compounds whose strength of adsorption onto carbon (represented by the Freundlich isotherm parameter, K) was greater than 200 (Snyder et al., 2003). Yoon et al. (2002) investigated the removal of 17β-estradiol (E2), 17α-ethinyl estradiol (EE2) and some other estrogenic compounds from distilled water using powdered activated carbon and reported that removals in excess of 90% was achieved. Khiari (2007) reported that granular activated carbon was effective for the removal of steroid hormones and hydrophobic micropollutants.

However, X-ray contrast media compounds (e.g., iopromide) and pharmaceuticals such as ibuprofen, meprobamate, sulfamethoxazole and diclofenac, were not effectively removed. Rowsell et al. (2009) compared the use of virgin and reactivated GAC for the removal of estrogenic compounds in wastewater effluent and noted that the reactivated GAC yielded higher efficiencies of 81% compared to the virgin GAC of 65%.

Based on the foregoing, both coagulation and activated carbon would be effective methods for pretreatment of the wastewater effluent to reduce EFOM concentration prior to UV/H2O2 treatment. Reducing the EfOM concentration would reduce the background scavenging capacity of the wastewater leading to lower dose and energy requirements for treating the micropollutants and an improvement in the degradation rates for the specific target compounds.

2.4 Energy Requirements of AOP Processes

Energy requirements are one of the main components in the operating costs of wastewater treatment plants, accounting for approximately 30% of the costs (Tchbanoglous et al., 2003; Zhelev et al., 2008; Force 2009). Energy conservation measures are usually considered by industries due to increases in operational costs, the need to fulfill environmental compliance, maintaining a competitive advantage, and the adherence to government policies and plans (Kumghare 2009). In relating this to micropollutants in the wastewater industry, at present, there are no regulations or guidelines for the effluent discharges from these facilities. However, given the possibility that regulations may be implemented in the future, energy requirements need to be considered.

Energy requirements of AOPs are commonly evaluated on the basis of the energy required to achieve 90% transformation or degradation of a particular compound (Parsons et al., 2004). Bolton et al. (2001) developed this standard “figure-of-merit” to determine the energy efficiency of AOPs, and to facilitate comparison between varying AOP processes, independent of the nature of the system. The standard figure-of-merit takes into account the overall kinetics of the organic contaminant and provides a direct correlation between the overall reaction kinetics and

18 the electrical energy consumption of the respective processes, given the electric-energy intensive nature of AOPs (Bolton et al., 2001). According to Bolton et al. (2001), the standard figure-of-merit also facilitate comparison between AOPs and conventional treatment processes, and provide data for scale-up operations.

The overall reaction kinetics of AOPs can be described as zero-order or first order on the basis of high and low concentrations of the pollutant respectively. In accounting for these conditions, two standard figures-of-merit are used: Electric energy per mass (EEM) and Electric energy per order (EEO) for zero-order and first-order reactions respectively (Bolton et al., 2001). A low value of either parameter indicates high efficiency (Bolton et al., 2001). The EEM is defined as the electrical energy that is required for the degradation of a target pollutant by one unit mass 3 (Bolton et al., 2001). The EEO (kWh/m /order) is defined as the electrical energy required for the degradation of a target pollutant by one order of magnitude in a unit volume of contaminated water (Bolton et al., 2001). For UV-based AOPs, such as UV/H2O2, EEO values are used to assess the energy requirements for treating a specific target compound. Katsoyiannis et al., 2011 3 reported EEO values of 0.17 – 0.75 kWh/m for UV/H2O2 treatment of para-chlorobenzoic acid (pCBA) in lake water. Shu et al. (2013) reported EEO values for naproxen, carbamazepine, diclofenac, gemfibrozil, ibuprofen, caffeine, 2,4-D, 2,4-DCP, and mecoprop ranging from 1.3 to 3 7.1 kWh/m with UV/H2O2 treatment using a medium UV lamp. It should be noted that EEO values are influenced by optical path length, water matrix, the fluence rate distribution, transmittance and hydraulics, water matrix and the target compound. EEO requirements are also expected to increase as the scavenging capacity of the water matrix increases (Autin et al., 2013; Katsoyiannis et al, 2011). Therefore, reducing the scavenging capacity of the wastewaters by removing EfOM should improve the energy efficiency of the AOP process.

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Chapter 3 Materials and Methods

3.1 Materials

3.1.2 Selection of Wastewaters

Secondary wastewater effluent samples from 18 municipal wastewater treatment plants across Ontario were used in this research. Two of the 18 plants were sampled on multiple occasions of which 6 were membrane bioreactor (MBR) plants and 12 were activated sludge treatment (AS) facilities. Nitrifying plants were selected to ensure the nitrite concentration of the secondary effluent was very low or negligible as nitrite can reduce the efficiency of an advanced oxidation process by scavenging the hydroxyl radicals. Plants treating municipal wastewater were selected so that any micropollutants present in the wastewater stream would be due to typical human uses of pharmaceuticals, personal care products, and other household items, rather than exhibiting a predominance of chemicals from industrial operations. All samples were collected after the secondary clarifier/membranes but prior to any filtration or disinfection. The specific characteristics of the waters used in each phase of this work as discussed in Chapters 4 – 8 will be presented in each respective chapter.

3.1.3 Selection of Micropollutant Compounds

Seven compounds (Table 3-1) were selected to be representative of different classes of micropollutants typically found in wastewaters. The criteria used for selecting the specific compounds included the reported frequency of occurrence in wastewater effluents, and the availability of photochemical information (i.e., hydroxyl radical reaction rate constant, quantum yield, and molar absorption coefficient) for each compound in the literature, as it was not within the scope of this work to identify the photochemical characteristics of the compounds (Table 3- 2). The final criterion was that the compounds could be readily analysed by liquid chromatography tandem mass spectrometry (LC-MS/MS) analysis at the University of Toronto Civil Engineering Environmental Laboratory using the protocol developed by the Ontario Ministry of the Environment & Climate Change.

Table 3-1: Selected micropollutant compounds for research study Chemical Compound Structure Classification Properties

MW = 194 g/mol Caffeine 1pKa = 10.4 Neurostimulant (C8H10N4O2)

Carbamazepine MW = 236 g/mol Analgesic, (C H N O) 1pKa =15.96 15 12 2 anticonvulsant

MW = 296.15g/mol 1pKa = 4.15 Non-steroidal anti- Diclofenac inflammatory agent (C H Cl NO ) 14 11 2 2 (NSAID)

MW = 215 g/mol Metabolite of the Clofibric acid 1pKa = 3.0 lipid regulator (C H ClO ) 10 11 3 clofibrate

Non-steroidal anti- MW = 230 g/mol inflammatory agent Naproxen 1pKa = 4.15 (NSAID) (C H O ) 14 14 3

5. Bacteriostatic MW = 253 g/mol Sulphamethoxazole antibacterial agent, 1pKa = 5.7 (C H N O S) low reactive 10 11 3 3 antibiotic

Natural hormone, MW = 272 g/mol 17B-estradiol endocrine disrupting 1pKa = 10.71 (C18H24O2) compound

1www.drugbank.ca, MW – molecular weight, pKa – acid dissociation constant

Table 3-2: Photochemical characteristics of the selected compounds Molar absorption OH reaction rate Quantum Yield Compound coefficient (M-1cm-1) constant (M-1s-1) x (mol/E) x 103 109 Caffeine 0.0018 3.92 5.90 Carbamazepine 0.0006 6.07 8.02 Diclofenac 0.29 4.77 8.38 Clofibric acid 0.27 0.927 5.03 Naproxen 0.0278 4.00 8.61 Sulphamethoxazole 0.0379 13.2 5.82 17β-estradiol 0.0550 0.403 14.1 Source: Wols et al. (2012)

3.1.4 Selection of Ecotoxicological Assessment Methods

Wastewater samples were analysed for acute toxicity using Daphnia magna and rainbow trout bioassays according to the protocol outlined by USEPA (2002), genotoxicity using the EBPI SOS-Chromo Test (EBPI, Brampton, Canada), and estrogenicity using the in-vitro yeast estrogen screening (YES) assay as previously described by Routledge et al. (1996), with minor modifications as described in Section 3.3.13. The specific bioassays used for each analysis were selected based on their common application in similar research, and each method has demonstrated a high level of sensitivity for any changes that occur in the respective sample (Lee et al., 2008; Rizzo 2011, Salste et al., 2007).

Acute toxicity analyses were conducted at the Ontario Ministry of the Environment and Climate Change Laboratory, Toronto, Canada. Genotoxicity testing was done at the University of Toronto Civil Engineering Environmental Laboratory, and the estrogenicity bioassays were performed at the Yargeau Laboratory at McGill University, Montreal, Quebec, Canada. Details of the procedures for each bioassay will be discussed in Section 3.2.

Assessing the secondary effluent for ecotoxicological effects and the extent to which these are affected by advanced oxidation treatment is an important additional analytical tool that complements chemical analyses of the effluent. Typically, the parent micropollutant compounds are monitored to gauge the efficiency of a given treatment process. However, incomplete degradation of these parent compounds can cause the formation of oxidation by-products or

32 transformation products that may or may not have a more toxic effect in the effluent. Additionally, the complexity of wastewater matrices means there may be other unknown constituents that could contribute to any negative toxicity effects. Since ecotoxicological assessment does not require prior knowledge of the individual toxic compounds, using multiple bioassays that each provides information on a specific biological effect or system is useful in determining the effect of the whole effluent sample rather than individual compounds.

Many micropollutants (pharmaceuticals and personal care products) are classified as endocrine disruptors. Studies have shown that concentrations as low as 0.1 ng/L can have negative effects on aquatic species, such as causing vitellogenesis in male fish (Aerni et al., 2004). Hence, estrogenicity testing of the secondary effluent before and after advanced oxidation treatment was conducted. Acute toxicity provides information on the effects from a single or multiple exposures to a given substance in a short period of time. Genotoxicity testing examines the genotoxic response which is an indication of the extent to which cell DNA is damaged due to exposure to a chemical substance.

3.1.5 Selection of Coagulants

+ Ferric chloride (FeCl3), aluminium sulphate (alum) and Hyper Ion (HI)705 polyaluminium chloride (PACl) were selected as the coagulants in this research for coagulation of the effluent. These products are commonly used in drinking water treatment and known to be effective for removing natural organic matter (NOM) in the water. In wastewater treatment, ferric chloride and alum are used for phosphorous control by chemical addition at different points in the treatment train. Given the familiarity of wastewater treatment plants with these two coagulants and their known effectiveness for removing organic matter, ferric chloride and alum were chosen. While PACl is not common to wastewater treatment, it was considered as an additional coagulant option given its effective use in drinking water treatment. Alum and FeCl3 were obtained from Sigma Aldrich with a purity of greater than 98%. Hyper+Ion (HI)705 polyaluminium chloride (PACl) is a high basicity (>80%) pre-hydrolyzed aluminium coagulant with an aluminium content of 11.3%–12.1% by weight and was obtained from an Ontario water treatment plant.

3.2 Experimental Protocols

3.2.1 Coagulation Experiments

The optimum doses for coagulation of the wastewater were determined using a series of bench- scale jar test experiments where the optimum doses were identified using a point-of-diminishing returns (PODR) analysis for dissolved organic carbon concentration. A point of diminishing returns is the dose for which a 10 mg/L incremental increase in the applied coagulant results in a change in dissolved organic carbon concentration removal of less than 0.3 mg/L (USEPA 1999).

A six-paddle jar testing apparatus (Phipps & Bird, VA, USA) was used where the wastewater was placed in 2L square jars (BKER2® Phipps and Bird, VA, USA). Coagulation was performed using a rapid mix of 200 rpm for 1 minute to ensure proper mixing of the coagulant, a slow mix at 30 rpm for 30 minutes to promote flocculation, and settling for a period of 30 minutes with the paddles removed. Mixing was achieved using a Cafrano BDC2002 stirrer (VWR International, Mississauga, ON). Coagulants were added using an Eppendorf pipette. Initial broad dose ranges of 0-50 mg Al/L for alum, 0-100 mg/L for FeCl3, and 0-100 mg Al/L for PACl were evaluated. Depending on the level of dissolved organic carbon removed, a narrower range of coagulant doses was selected for testing. Alum and ferric chloride coagulation were optimised at pH 6.0 with sulphuric acid (H2SO4) or sodium hydroxide (NaOH) used for pH adjustment of the effluent sample. The supernatant from each jar was collected and filtered using a 0.45 µm glass microfiber filter (Super-450 filters) for further analysis of pH, UV absorbance at 254 nm (UV254nm), and dissolved organic carbon concentration (DOC) and concentration of organic fractions in the wastewater using liquid chromatography organic carbon detection (LC-OCD) analysis. All experiments were performed in duplicate.

3.2.3 Advanced Oxidation (UV/H2O2) Experiments

UV/H2O2 advanced oxidation was conducted in a 42 L cylindrical stainless steel Calgon Carbon Rayox Advanced Oxidation Batch Pilot Reactor (Figure 3-1). The AOP reactor was equipped with a 1kW medium-pressure (MP) Hg-lamp in the centre of a quartz sleeve, a mixer, and a steel shutter that was used to control the length of time the water sample was exposed to the UV

34 light. Wastewater samples that were pretreated using coagulation were spiked with 50 µg/L of each compound (caffeine, naproxen, diclofenac, sulphamethoxazole, carbamazepine, clofibric acid, and 17-estradiol) before UV/H2O2 treatment at 0, 10 and 20 mg/L H2O2 for a maximum exposure time of 5 minutes. Sodium thiosulphate was added to quench the residual hydrogen peroxide before analysing the samples for pharmaceutical concentrations and UV254.

Figure 3-1: Calgon Carbon Rayox® Advanced Oxidation Batch Pilot Reactor

Pretreatment of the wastewater prior to UV/H2O2 was performed in a 100L stainless steel tank (Royal Industries Inc. Chicago, IL) using the optimum doses from bench-scale experiments and following the same procedure as previously described in Section 3.2.1 and Section 3.2.2.

Following pretreatment, samples were analysed for pH, DOC concentration, UV254, and hydroxyl scavenging capacity.

A volume of 40 L of wastewater was added to the reactor through one of the sample ports. The stirrer was set to a mixing speed of 6 rpm to ensure the wastewater was well-mixed for the duration of the experiment. The 1 kW MP lamp was turned on and allowed to warm up for 20 minutes before starting the experiment. During this period, hydrogen peroxide and 50 µg/L of each micropollutant compound (caffeine, carbamazepine, 17-estradiol (E2), diclofenac, sulphamethoxazole, clofibric acid, naproxen) was spiked into the wastewater. This initial concentration exceeds typical levels reported in the environment, but a high initial concentration

35 was necessary to ensure that the amount of degradation of each compound could be easily measured. After 20 minutes, 500 mL of effluent was collected as the initial sample at t = 0 minutes. The steel shutter was opened and the sample was irradiated for exposure times of 0.5, 1, 2, 3, 4 and 5 minutes following which approximately 500 mL of wastewater was collected in a 500 mL amber bottle. At each time interval, the shutter was closed and ~20 ml of wastewater was drained from the sampling port to flush the line from the previous sample before collecting the actual sample. At the end of the maximum 5 minutes exposure time, 2 L of the irradiated wastewater was also collected in a 2.5 L amber bottle and acidified to pH 2 using concentrated sulphuric acid for genotoxicity analysis. Residual hydrogen peroxide in the wastewater samples was quenched using sodium thiosulphate before any further analysis.

On completion of an experiment at a given hydrogen peroxide dose, the UV lamp and the stirrer were turned off and the effluent from the batch reactor was drained completely. Before another experiment was performed, the reactor was cleansed by filling it with 42 L of tap water and setting the stirrer at a speed of 9 rpm for approximately 3 minutes before draining the unit. This was duplicated to ensure that the reactor was properly cleansed. The wastewater samples for micropollutant concentrations were analysed using solid phase extraction (SPE) followed by liquid chromatography tandem mass spectrometry (LC-MS/MS). Samples for genotoxicity analysis were analysed using SPE followed by the SOS ChromoTestTM (EBPI, Brampton, Canada).

3.2.4 Background Scavenging Capacity Experiments

Overall background scavenging capacity was determined by monitoring the degradation of a  known OH probe compound (methylene blue) at a given hydrogen peroxide (H2O2) concentration and UV fluences using the protocol outlined by Rosenfeldt et al., (2007). A Calgon Carbon® quasi-collimated beam apparatus (Model PSI-I-120, Calgon Carbon Corporation, USA), as shown in Figure 3-2, was used for measuring the overall scavenging  capacity of the wastewater and the OH exposure per UV fluence (ROH,UV). The collimated beam apparatus can be equipped with a 40W low-pressure (LP) Hg UV lamp or a 1kW medium-

36 pressure (MP) UV lamp. When using the LP lamp, a parabolic concentrator was also used to increase the incident irradiance from the lamp.

Figure 3-2: Calgon Carbon UV-collimated beam apparatus

For the experiment, a volume of 100 mL of wastewater was spiked with 3.0 µM of methylene blue (MB). A volume of 10 mL of the spiked sample was placed in a pyrex petri-dish. A small stir bar was added to ensure that the sample was well mixed and homogeneous, without creating a vortex or disturbing the surface of the sample. Hydrogen peroxide at concentrations of 0, 5, 10, 15 and 20 mg/L were spiked into the wastewater prior to irradiation for the exposure times required to obtain UV fluences of 0 to 200 mJ/cm2 with the LP UV lamp. All samples were placed at a distance of 39.7 cm from the lamp, and had a water path length of 0.53 cm. Irradiations were performed in duplicate. The MB degradation was monitored by measuring the change in absorbance at 664 nm and applying the Beer-Lambert law. Hydrogen peroxide concentrations were determined using the triiodide method (Klassen et al., 1994). The photolytic degradation rate of methylene blue due to direct UV photolysis only (i.e., in the absence of

H2O2) was also monitored to ensure that the observed methylene blue degradation during

UV/H2O2 treatment was solely due to reactions with the hydroxyl radical.

The incident irradiance of the UV lamp was determined using iodide-iodate actinometry for the LP UV lamp (Rahn 1997) and ferrioxalate actinometry was used for the MP UV lamp (Sharpless and Linden 2003). For the 40W LP UV lamp, the Bolton® Excel Spreadsheet for fluence calculations using a low-pressure lamp with a suspension depth of less than 2 cm (Bolton 2004) with a petri factor of 0.9572 and the correction factors outlined by Bolton et al. (2003) were used to determine the average fluence rate of 0.0978 mW/cm2. For the MP lamp, the Bolton® Excel Spreadsheet for fluence calculations using a medium pressure lamp with suspension depth of less than 2 cm (Bolton 2004) with a petri factor of 0.975, and the correction factors outlined by Bolton et al., (2003) were used to determine the average fluence rate of 2.68 mW/cm2.

3.2.5 Quality Control for Advanced Oxidation

Prior to any of the advanced oxidation experiments with the wastewater, a series of quality control tests were performed using methylene blue (MB), a hydroxyl radical probe compound, in Milli-Q (MQ) water buffered to pH 8 and a wastewater sample. The objective of these tests was to verify that MB would not be easily degraded by hydrogen peroxide, other constituents in the wastewater, or UV photolysis. MB has been Six test conditions were evaluated and the corresponding degradation rates for MB are shown in Table 3-3. All experiments were performed in duplicate using the collimated beam apparatus. A typical MB calibration plot is shown in Figure 3-3.

Table 3-3: Degradation rates of MB at different treatment scenarios MB degradation rate Condition Treatment Matrix (cm2 mJ-1) -5 1 no UV or H2O2 ultra-pure water 1.20 x 10 -6 2 no UV, H2O2 only ultra-pure water 7.45 x 10 3 no UV wastewater 8.93 x 10-6 4 UV only ultra-pure water 4.72 x 10-4 -3 5 UV + H2O2 ultra-pure water 3.72 x 10 6 UV only wastewater 2.48 x 10-3

Figure 3-3: Calibration graph for methylene blue (MB)

A volume of 40 L of deionised water was added to the reactor and spiked with 2.5 µM of methylene blue. UV photolysis of the methylene blue was monitored by collecting samples at 1 minute intervals for an overall exposure time of 15 minutes. MB degradation with UV/H2O2 treatment at 10, 15, and 20 mg/L H2O2 was also performed. The methylene blue concentration at each exposure time was determined by measuring UV absorbance at 664 nm with an Agilent 8453 UV-VIS spectrophotometer. The observed MB degradation rate (Table 3-3) and the known hydroxyl radical reaction rate constants for both MB (2.1 x 1010 M-1s-1) (Buxton et al., 1988) and the micropollutant compounds (Table 3-2), were subsequently used to determine that a 5 minute exposure time would be adequate for the treatment of the target micropollutants in the wastewater samples in the Rayox reactor. Plots of the MB degradation curves are shown in Figure 3-4.

The OH probe, para-chlorobenzoic acid (pCBA) is commonly used in UV-advanced oxidation studies (Shu et al., 2013; Katsoyiannis et al., 2013; Lee et al., 2010). Nevertheless, MB was determined to be an appropriate alternative OH probe compound as its degradation rate constant in the presence of UV alone, with an order of magnitude of 10-4, was considered to be negligible (Table 3-3). MB can also be readily used, is easily detected using UV spectrometry, and requires no additional sample preparation.

UV photolysis UV+10 mg/L H2O2 UV+20 mg/L H2O2 2.5

2.0

1.5

1.0

MB Concentration Concentration MB (µM) 0.5

0.0 0 2 4 6 8 10 12 14 16 Time (mins)

Figure 3-4: Methylene blue degradation in ultra-pure water in the Rayox at different H2O2 concentrations

5 -1 -1 Nitrate ( ) anions have a low hydroxyl radical reaction rate constant (1.0 x 10 M s ) (Buxton et al., 1988) and according to Glaze et al. (1995) and Li et al. (2008) it is not considered a major OH scavenger, but its main effect is reducing the amount of photons that are absorbed by hydrogen peroxide. This subsequently reduces the steady-state concentration of hydroxyl radicals available to react with the target compounds, thereby reducing their degradation rate.  9 -1 -1 Chloride anions can react with OH (kOH = 4.3 x 10 M s ) forming radicals that subsequently reform the respective anion and OH (Liao et al., 2001). To verify that these two anions were not major OH scavengers, quality control tests using MB were performed where the anion concentration in a wastewater sample was doubled by spiking in additional chloride or nitrate anions. The degradation rates of MB in the unspiked and spiked samples were compared. For chloride, wastewater with an initial chloride anion concentration of 56 mg/L Cl-, was spiked with sodium chloride (NaCl) to yield a spiked sample of 117 mg/L . As shown in Table 3-4, doubling the chloride concentration decreased the degradation rate by 3%, but this was not considered to be significant as it was close to the relative standard deviation (2.2%) of the

40 samples. A plot of MB degradation as a function of UV fluence with chloride anions for the spiked and unspiked sample is shown in Figure 3-5.

Table 3-4: Methylene blue degradation rates in the presence of chloride anions

MB Degradation Rates (cm2mJ-1) Parameter Unspiked (56 mg/L Cl-) Spiked (117 mg/L Cl-) Sample 1 3.30 x 10-3 3.10 x 10-3 Sample 2 3.20 x 10-3 3.20 x 10-3 Average 3.25 x 10-3 3.15 x 10-3 Standard deviation 7.07 x 10-5 7.07 x 10-5 % RSD 2.2 2.2

UV Fluence (mJ/cm2) 0 200 400 600 800 1000 1200 0.0 -0.5

) -1.0 o -1.5 -2.0

-2.5 ln (MB/MB ln -3.0 -3.5 -4.0 Unspiked Spiked Figure 3-5: Methylene degradation in the presence of chloride anions

For nitrate, a similar experiment was performed where sodium nitrate (NaNO3) was spiked into a wastewater sample with an initial concentration of 62 mg/L doubling the concentration to

126 mg/L . The MB degradation rates in the unspiked and spiked nitrate samples were determined. It was found that doubling the nitrate concentration increased the degradation rate by 1.8% as shown in Table 3-5 which was not considered to be significant. The degradation plots are shown in Figure 3-6.

Table 3-5: Methylene blue degradation rates in the presence of nitrate

MB Degradation Rates (cm2mJ-1) Parameter Unspiked (62 mg/L ) Spiked (126 mg/L ) Sample 1 3.62 x 10-3 3.66 x 10-3 Sample 2 3.61 x 10-3 3.70 x 10-3 Average 3.62 x 10-3 3.68 x 10-3 Standard deviation 3.86 x 10-6 3.00 x 10-5 % RSD 0.11 0.81

UV Fluence (mJ/cm2) 0 200 400 600 800 1000 1200 0.0 -0.5

-1.0 )

o -1.5 -2.0 -2.5

ln (MB/MB ln -3.0 -3.5 -4.0 -4.5

Unspiked Spiked Figure 3-6: Methylene blue degradation with nitrate anions

3.2.6 Sampling Bottles

All the amber bottles used in the experiments were purchased pre-cleaned from Systems Plus (Edmonton, AB). After the initial use, the bottles were cleaned in the Environmental Laboratory using a MieleDisinfektor G7736 dishwasher. The bottles were cleansed using a detergent wash (LaboClean FT, Dr. Weigert, Germany) followed by an acid rinse (Neodisher acid, Dr. Weigert). Bottles were subsequently rinsed three times with distilled water and baked in the oven at 300 oC for at least six hours.

3.3 Analytical Methods

3.3.1 Analysis of Micropollutant Compounds

Analysis of the micropollutant compounds (caffeine, carbamazepine, 17-estradiol, sulphamethoxazole, clofibric acid, naproxen, and diclofenac) was conducted using the liquid chromatography tandem mass spectrometry (LC-MS/MS) method developed by the Ontario Ministry of the Environment and Climate Change (MOE 2008). All analyses were conducted in the Environmental Laboratory of the Civil Engineering Department at the University of Toronto.

The micropollutant concentrations were analysed using solid phase extraction with Oasis hydrophilic-lipophilic balance (HLB) extraction cartridges (Waters Oasis HLB 6 cc, 150 mg, 30 µm) followed by analysis with an Agilent LC (Model 212-LC) system equipped with an Agilent 500-MS IT mass spectrometer, a Pursuit XRs-C18 guard column (Metaguard 2.0 mm ID x 3 mm), and a Pursuit XRs Ultra 2.8-C18 analytical column (2.0 mm ID x 100 mm, 2.8 µm particle size)(Agilent Technologies, Mississauga, ON. Canada). The reagents and instrument operating conditions are shown in Table 3-6 and Table 3-7 respectively. The MOE Method used for the analysis is outlined in Table 3-8. The surrogates and internal standards were obtained from CDN isotopes (Pointe-Claire, Quebec) and are shown in Table 3-9. The standard solutions used in the study are shown in Table 3-10. The target micropollutant compounds were obtained from Sigma Aldrich Inc. (Oakville, Ontario) in the powdered form.

In summary, duplicate volumes of approximately 500 mL of wastewater was collected in 500 mL amber bottles with Teflon-lined caps. Samples were stored in the dark and refrigerated at 4oC until analysis which was completed within 10 days of the sample being collected. From these bottles, samples of 400 mL each were collected and deuterated compounds (Table 3-9) were added as surrogates. The duplicated samples were then extracted using the Oasis hydrophilic-lipophilic balance (HLB) extraction cartridges (Waters Oasis HLB 6 cc, 150 mg, 30 µm) followed by analysis with the Agilent LC (Model 212-LC) system. The volume of 1mL of each extracted sample (in methanol) was dried using ultra-high purity nitrogen gas and

43 reconstituted to a volume of 200 µL with an internal standard solution that contained isotopically-labelled compounds. The reconstituted samples were analysed using LC-MS/MS. The MS/MS analysis was performed using an electrospray ion source in both the positive and negative ionisation mode.

The extraction recovery for each batch of samples was determined by comparing the measured spike sample concentration with the known spike concentration for each compound in the sample. The concentration of the micropollutants in the samples was determined using a 7-point calibration curve for which the calibration standards were extracted along with the samples and included with each batch of samples. The calibration curves for the compounds are shown in Appendix E. Extracted samples were stored in the dark at 4 oC until analysis which was completed within 10 days of extraction.

Table 3-6: Reagents for LC-MS/MS analysis of the micropollutants

Reagent Supplier and Purity Sigma Aldrich,  99.9% LC-MS Acetonitrile (CH CN) 3 Chromasolv Sigma Aldrich,  99.0% LC-MS Ammonium acetate (C H NO ) 2 7 2 Ultra, Ethylenediaminetetraacetic acid disodium salt dihydrate Sigma Aldrich, 99.4-100.6% ACS (EDTA-Na2) reagent,

Heptafluorobutyric acid (HFBA) (C4HF7O2) Sigma Aldrich, 99%, Sigma Aldrich,  99.9%, LC-MS Methanol (CH OH) 3 Chromosolv, Nitrogen gas Ultra high purity Sigma Aldrich,  98% reagent Sodium hydroxide (NaOH) (pellets) grade, Sigma Aldrich, 95-98% ACS Sulphuric acid (H SO ) 2 4 reagent, Water Sigma Aldrich, LC-MS Chromasolv

Table 3-7: Instrument operating conditions for LC-MS/MS analysis Parameter Description Column Pursuit XRs Ultra 2.8 C18 100 x 2.0 mm Maximum pressure: 5000 psi Mobile Phase A: 5 mM ammonium acetate in water, pH adjusted to 7.0 ± 0.1 with (Negative Mode) 0.5M NH4OH B: Acetonitrile Mobile Phase A: 0.03% HFBA in water (Positive Mode) B: Acetonitrile Injection Method Injection volume: 100 µL Sample Loop Volume 100 uL Syringe Volume 250 uL Auto Sampler method Loop: 100 µL LC Method Negative Mode Positive Mode (Pump Program) Time (min) %A %B Time (min) %A %B 00:00 90 10 00:00 85 15 25:00 20 80 13:00 0 100 26:00 20 80 14:00 0 100 27:00 90 10 16:00 85 15 28:00 90 10 17:00 85 15

Flow rate: 200 µL/min Flow rate: 200 µL/min MS Conditions Scan Mode: SIS (Selected Ion Selection Ionization Type: Electrospray Total Run Time Negative Mode: 28 minutes Positive Mode: 17 minutes MS/MS Electrospray Negative Mode Time segments Start End Clofibric acid,clofibric acid-d4 5:00 10:00 Diclofenac, Diclofenac-d4 10:00 13:50 Bisphenol-A-d16. Estradiol, estradiol- 13:50 25:00 d2 25:00 28:00 Off Positive Mode Start End IS, Caffeine, caffeine-d3 0:00 4:50 Sulphamethoxazole, CBZ, CBZ-d10 4:50 8:24 Naproxen, naproxen-d3 8:24 15:00 Off 15:00 17:00

Table 3-8: Sample extraction protocol for LC-MS/MS Analysis

1. Samples are collected in 1 L, amber glass bottles with TeflonTM-lined caps. Sample bottles are filled to the shoulders (~ 900 mL) 2. Prior to extraction, samples are refrigerated (4° C) and stored in the dark. 3. Shake each sample to make homogeneous, and transfer 400 ± 20 mL of the sample into pre-cleaned, labeled 1L amber bottles for extraction. Keep the remaining sample in storage. 4. With each sample set, prepare one Blank and one Spike Sample. Blanks consist of 400 mL of Milli-Q® water and Spikes consist of 400mL of the wastewater spiked to a concentration of 10 µg/L. 5. Add 200 μL surrogate solution to each sample, blank, and spike.

6. Add ~2 g of EDTA-Na2 to each sample, blank and spike. Shake gently for 10 minutes to

completely dissolve the EDTA-Na2. 7. Add 10 mL of 0.25 M ammonium acetate solution into each sample, blank and spike and shake well. 8. Using the pH meter, slowly adjust the pH of each sample, blank and spike to 6.95 ± 0.05

using 50% NaOH, 10% NaOH and 10% H2SO4 solutions. 9. Place the appropriate number of SPE cartridges on the Visiprep manifold condition the SPE cartridges on the manifold using the following procedure: NOTE: Flow should be set at approximately 0.5 mL/minute or 20 drops/minute. 10. Slowly aspirate approximately 5 mL methanol through each SPE cartridge and do not allow the cartridges to go dry. 11. Close the valves on each cartridge and add approximately 5 mL Milli-Q® water to each cartridge. 12. Attach a pre-rinsed Teflon™ adaptor/Teflon™ tube to each SPE cartridge. 13. Place the free end of each Teflon™ adaptor/Teflon™ tube in a separate sample (or blank or spike) bottle, making sure the tube reaches the bottom. Label the appropriate sample name on each SPE cartridge.

Table 3-8 (Cont’d):

14. Once the blank, spike and sample bottles have been connected to the SPE cartridges, open the valves on each cartridge and apply vacuum (approximately - 34 kPa/-10mm Hg) to the Visiprep manifold. Flow rates through the SPE cartridges should be approximately 5-10 mL/minute. NOTE: Under normal conditions, it will take ~1.5 hours for a 400 mL water sample to pass through the cartridge. However, samples will finish at varying times depending on the actual vacuum at the individual SPE cartridge and the suspended solid content of the sample. 15. During this time, begin drying of controls. Also, prepare mobile phases (if analyzing on same day). 16. After all samples, blanks and spikes have passed completely through the cartridges, rinse each sample bottle with 10 mL of Milli-Q® water, at a flow rate of approximately 5-10 mL/minute. 17. Wash the cartridges with 5 mL of 5% methanol in Milli-Q® water (v/v), at 5-10 mL/minute. Vacuum dry each cartridge for 2 minutes. Once all the water has been aspirated, turn off the vacuum manifold and remove the Teflon™ adaptor/Teflon™ tubes from the SPE cartridges. 18. Place the Teflon adaptor/Teflon tubes on Kimwipes. Remove the SPE cartridges from the Visiprep manifold and place them on clean Kimwipes. Make sure the cartridges are labelled properly. 19. Using Kimwipes, dry the inside of the SPE cartridges. 20. Remove the Visiprep manifold cover. Dry any excess water from the underside of the manifold cover and place the cover on clean Kimwipes. 21. Allocate one 15 mL polypropylene centrifuge tube for each SPE cartridge. 22. Label each 15 mL centrifuge tube with the appropriate sample identification number and place the tube in the proper slots of the Visiprep collection rack. 23. Place the Visiprep collection rack in the vacuum manifold and reseat the vacuum manifold cover with the SPE cartridges. Check to ensure that the manifold cover exhaust tubes are aligned with the 15-mL polypropylene centrifuge tubes.

Table 3-8 (Cont’d):

24. Add 5 mL of methanol to each SPE cartridge. Turn ON the manifold vacuum and let methanol slowly reach the bottom of the cartridge. Turn OFF vacuum and close the valves, soak the cartridges with methanol for 3 minutes. 25. Open manifold valves and elute by gravitational force until dry. Collect the eluent in the corresponding centrifuge tube. NOTE: Some cartridges may require slight vacuum when beginning the elution to initiate a constant flow of methanol. If there is no flow, let cartridges sit for 5 more minutes before initiating elution. 26. Once all the methanol appears to be eluted, turn ON the manifold vacuum for approximately 2 minutes to aspirate remaining methanol. 27. Turn OFF the manifold vacuum. Lift the manifold cover and remove the collection rack containing the polypropylene centrifuge tubes with the sample extracts from the manifold. 28. Using disposable Pasteur pipettes thoroughly mix each final extract. Using a calibrated pipettor, transfer 1mL of each final extract to separate clean 1.5 mL vials. 29. Evaporate extracts in vials to dryness using a gentle stream of nitrogen (pressure < 2 psi; between second and third line on drying apparatus pressure gauge). 30. Reconstitute each vial with 200 μL of Internal Standard Solution and cap the vial with a Teflon™-lined septum screw cap. Rinse the walls of the vial thoroughly by gently tumbling the vial. Let sit for at least 30 minutes before transferring the concentrated extract into a vial insert and recapping. 31. Store in freezer (-15±10°C) until required for LC/MS/MS analysis.

Table 3-9: Surrogates and internal standards for LC-MS/MS analysis. Compound Supplier Surrogates Caffeine-d3 CDN Isotopes Carbamazepine-d10 CDN Isotopes Diclofenac-d4 (phenyl-d4-acetic) CDN Isotopes Sulphamethoxazole-d4 CDN Isotopes Naproxen-d3 CDN Isotopes 17β-estradiol-16,16-d2 CDN Isotopes Clofibric-d4 acid CDN Isotopes Internal Standards Bisphenol-A-d16 CDN Isotopes 13 C6-sulphamethazine CDN Isotopes

Table 3-10: Preparation of standard solutions for micropollutant analysis

Solutions Preparation Steps Blank 400 mL of ultra-pure Milli-Q water Stock Micropollutants 10 g/L solution of each micropollutant compound prepared by adding 1g of each compound to 100 mL methanol.

Surrogates 1000 mg/L solution prepared by adding 0.010g of each surrogate to 10 mL of methanol

Internal standards 1000 mg/L solution prepared by adding 0.010 g of each standard to 10 mL of acetonitrile Intermediate Micropollutants 100 mg/L solution prepared by adding 1 mL of each stock solution to 100 (mixture) mL of methanol

Dosing Micropollutants 10 mg/L dosing solution prepared by adding 5 mL of intermediate (mixture) solution to 50 mL methanol

Surrogates 50 mg/L solutions prepared by adding 500 µL of the stock solution to 10 (mixture) mL of methanol

Internal standard 200 ng/L solution prepared by adding 100 µL of the stock solution to 25 (positive and mL of ultra-pure Milli-Q water for each standard negative)

3.3.2 Anion Analysis

Anion concentrations (chloride, nitrite, nitrate, sulphate and phosphate) in the wastewater were measured in accordance with the USEPA Method 300. Analysis was performed using a ThermoScientific Dionex Ion Chromatography system comprised of a Dionex ICS-5000+DP, ICS-5000+DC conductivity detector, Dionex IonPac AS18 (2 mm x 250 mm) analytical and guard IC column, and a Dionex AS-AP autosampler. Sodium chloride, sodium nitrate, sodium nitrite, dihydrogen phosphate, and potassium sulphate were used as the calibration standards for chloride, nitrate, nitrite, phosphate and sulphate respectively. All reagents were sourced from Sigma Aldrich, Canada. Stock solutions (1000 mg/L) of each anion were prepared in ultra-pure Milli-Q water. For each anion, a 6-point calibration graph, which was prepared for each batch of samples, was used to determine the anion concentrations in the wastewater. Sample calibration graphs for each anion are shown in Appendix D. The effluent sample was filtered using a 0.45 µm filter before analysis to ensure particulates were removed. The MDL for each anion, as shown in Table 3-11, was determined using the standard deviation of 8 replicates of 1 mg/L concentrations of each compound prepared in Milli-Q water and the Student’s t-value for a 99% confidence interval.

Table 3-11: Anion Method Detection Limits at 1 mg/L spike concentration Compound MDL (mg/L) Chloride 0.68 Nitrate 0.04 Nitrite 0.08 Sulphate 0.27 Phosphate 0.07

3.3.3 Overall Background Scavenging Capacity

Overall background scavenging capacity (∑ks[S]) is defined as the sum of the individual scavenging potentials exerted by hydroxyl radical scavengers in a wastewater, and can be determined by monitoring the degradation of a hydroxyl radical probe compound at different hydrogen peroxide doses and UV fluences, according to the protocol outlined by Rosenfeldt et

50 al. (2007). Methylene blue (MB) was used as the probe compound as its concentration can be easily measured using UV spectrometry, and it is not easily degraded by UV photolysis. UV irradiation experiments were conducted using the Calgon Carbon® collimated beam apparatus equipped with a 40W low-pressure UV lamps and parabolic concentrator. The scavenging  characteristics of the wastewater influences the OH exposure per UV fluence (ROH,UV) in the wastewater matrix.

  ROH,UV, defined as the OH exposure per UV fluence, is the steady-state concentration of OH radicals present in the wastewater per unit dose of UV for an initial concentration of hydrogen peroxide. Hence, wastewater with a higher background OH scavenging will have a lower  steady-state concentration of OH per unit UV (i.e.,, a lower ROH,UV). Therefore, measuring the

ROH,UV in the wastewater by monitoring the degradation of MB is used to determine the overall background scavenging capacity (∑ks[S]). The degradation of methylene blue (MB) can be described by Equation 3.1

[ ] ( [ ] )[ ] (3.1)

Integrating Equation 3.1 and converting the time-based rate constants to fluence-based rate constants yields Equation 3.2

∫[ ] (3.2)

where is the observed degradation rate constant of MB at a given H2O2 concentration, is the degradation rate constant of MB due to direct UV photolysis only, is the MB hydroxyl radical reaction rate constant (2.1 x 1010 M-1s-1) (Buxton et al., 1988), and H is UV 2 fluence (mJ/cm ). Figure 3-7 shows a typical degradation plot for MB at different H2O2 concentrations.

From Equation 3.2, the units of the ROH,UV parameter can be expressed as shown in Equation 3.3. Based on the assumption that the concentration of the hydroxyl radicals is at steady-state

• ([ OH]ss) (Glaze et al., 1995), a mathematical expression for ROH,UV can be written as in Equation 3:4.

(3.3)

⁄

[ ] (3.4)

The steady-state concentration of hydroxyl radicals is the ratio of the rate of formation of hydroxyl radicals due to photolysis of hydrogen peroxide to the rate of consumption of the generated hydroxyl radicals by scavenging constituents in the water matrix. This can be mathematically expressed by Equation 3.5 (Rosenfeldt et al., 2007):

[ ] [ ] (3.5) {∑ [ ] [ ] [ ] }

-2 where Ea is the average fluence rate (mW cm ), U254 is the energy per 1 mol of photons at -1 5 254nm (J Es ) which is 4.72 x 10 J/mol, is the hydrogen peroxide molar absorption -1 -1 coefficient, which is equal to 18.7 M cm at 254 nm (Bolton et al., 1994), OH is the quantum yield of the production of •OH radicals from hydrogen peroxide photolysis which is 1.0 mol E-1,

∑ [ ] is the background scavenging capacity of the matrix, kMB,OH and kH2O2,OH are the hydroxyl radical reaction rate constants for methylene blue (2.1 x 1010 M-1s-1) and hydrogen 10 -1 -1 peroxide (2.7 x 10 M s respectively) (Buxton et al., 1988), and [MB] and [H2O2] are the molar concentrations of methylene blue and hydrogen peroxide.

The background scavenging capacity of the effluent can be determined by incorporating Equation 3.5 into Equation 3.4 and rearranging to form Equation 3.6.

[ ] [ ] (3.6) {∑ [ ] [ ] [ ] }

Inverting Equation 3.6 gives the following:

∑ [ ] [ ] [ ] (3.7) [ ]  [ ]

Equation 3.7 can be rewritten as follows:

(∑ [ ] [ ] ) (3.8) [ ]  [ ] 

Equation (3.8) can be rewritten as follows:

(∑ [ ] [ ] ) (3.9)  [ ] 

From Equation 3.10, a plot of versus should yield a linear plot with a gradient [ ] (m). This gradient (m) is equivalent to Equation 3.10, which can be rearranged to calculate the overall background scavenging capacity (∑ [ ] ) of the wastewater using Equation 3.11.

(∑ [ ] [ ] ) (3.10) 



∑ [ ] [ ] (3.11)

where m is the gradient of a plot of versus , is the hydrogen peroxide molar [ ] -1 -1 absorption coefficient, which is equal to 18.7 M cm at 254 nm (Bolton et al., 1994), OH is the quantum yield of the production of •OH radicals from hydrogen peroxide photolysis which is 1.0 -1 -1 mol E (Rosenfeldt et al., 2007), U254 is the energy per 1 mol of photons at 254 nm (J Es ) 5 which is 4.72 x 10 J/mol (Bolton 2001), is the MB hydroxyl radical reaction rate constant (2.1 x 1010 M-1s-1) (Buxton et al., 1988) and [MB] is the molar concentration of MB. An example of a typical versus plot is shown in Figure 3-8. [ ]

UV Fluence (mJ/cm2) 0 50 100 150 200 250 0.10

0.00

-0.10

-0.20

-0.30 ln [MB]/[MBo] ln

-0.40

-0.50

-0.60

0 mg/L H2O2 5 mg/L H2O2 10 mg/L H2O2 15 mg/L H2O2 20 mg/L H2O2

Figure 3-7: Typical degradation plot for methylene blue at different H2O2concentrations

4.5E+13

4.0E+13 y = 5.8807E+09x + 2.8903E+11 R² = 9.9385E-01 3.5E+13

3.0E+13

2.5E+13 OH,UV

1/R 2.0E+13

1.5E+13

1.0E+13

5.0E+12

0.0E+00 0 1000 2000 3000 4000 5000 6000 7000 8000

-1 1/[H2O2] (M )

Figure 3-8: Typical plot of 1/ROH,UV versus 1/H2O2

Using the overall scavenging capacity, the scavenging capacity exerted by the effluent organic matter (EfOM) in the wastewater was estimated by subtracting the sum of the scavenging potential of known scavenging constituents present in the matrix, other than EfOM, from the measured overall background scavenging capacity. The main known scavengers in wastewaters, other than EfOM, include bicarbonate, carbonate and nitrite (Keen et al., 2014; Rosario-Ortiz et al., 2010; Rosenfeldt et al., 2007); hence, Equation 3.12 was used.

[ ] ∑ [ ] ( [ ] [ ] [ ]) (3.12)

10 -1 -1 6 -1 -1 where = 1.0 x 10 M s (Buxton et al., 1988), = 8.5 x 10 M s (Buxton et al., 8 -1 -1 1988), and = 3.9 x 10 M s (Buxton et al., 1988).

Using the EfOM scavenging capacity, the hydroxyl radical reaction rate constant of EfOM

(kEfOM,OH) was determined by dividing the EfOM scavenging capacity by the molar organic carbon concentration in the matrix. EfOM rate constants are reported on a per mol of carbon -1 -1 basis (Mc s ) assuming 12 g per mol of C.

3.3.4 Liquid Chromatography Organic Carbon Detection (LC-OCD) analysis

Liquid chromatography organic carbon detection (LC-OCD) is a technique used to quantitatively characterise the components of organic matter into five constituents based on molecular weight distribution. These constituents are biopolymers, humic acids, building blocks of humics, low molecular weight (LMW) acids and LMW neutrals as shown in Table 3-12. This analysis is conducted using the procedure outlined by Huber et al. (1992) and an LC-OCD system. The system is comprised of a weak cation exchange chromatographic column on a polymethacrylate basis (250 mm × 20 mm, TSK HW 50S, 3000 theoretical plates) along with an annular UV reactor, UV detector, organic nitrogen detector, and an organic carbon detector as developed by Huber et al. (2011). Samples were filtered through 0.45µm polyethersulfone (PES) membrane filters. A phosphate buffer mobile phase (pH 6.8) at a flowrate of 1.1 mL/min was delivered to an autosampler with a 1 mL injection volume to the chromatographic column. The DOC of the samples are separated using liquid chromatography followed UV detection and

55 organic carbon detection. Organic carbon detection, as required for this study, is achieved using a Grantzel gravity flow-thin film reactor as the detector. The ChromCalc software was used for processing the LC-OCD data (DOC-Labor Dr. Huber).

Table 3-12: Description of the LC-OCD components Molecular Weight LC-OCD Constituent Description Range (g/mol) Polysaccharides, proteins and amino Biopolymers > 20,000 sugars Humics ~ 1000 Humics substances Building blocks of 300 - 500 Breakdown products of humics humics Mono-oligosaccharides, ketones, LMW neutrals < 350 alcohols, aldehydes LMW acids < 350 Monoprotic organic acids

3.3.5 Residual Hydrogen Peroxide

The initial and residual hydrogen peroxide concentrations after UV irradiation were determined using the triiodide method outlined by Klassen et al., (1994). The reagents required for the analysis are listed in Table 3-13. The volume of sample used for analysis of each hydrogen peroxide concentration is listed in Table 3-14.

Table 3-13: List of reagents for hydrogen peroxide analysis Reagent Grade and Supplier Potassium iodide, KI ACS grade, Sigma Aldrich Sodium hydroxide, NaOH ACS grade, Sigma Aldrich Ammonuim molybdate tetrahydrate (NH4)6Mo7O24•4H2O ACS grade, Sigma Aldrich Potassium hydrogen phtalate, C8H5KO4 ACS grade, Sigma Aldrich

Table 3-14: Sample volumes used for hydrogen peroxide analysis Hydrogen Peroxide Concentration (mg/L) Sample Volume (mL) 5 1 10 0.5 15 0.4 20 0.3

The required sample volume for each hydrogen peroxide concentration was added to an amber vial with 2.5 mL of each reagent solution A and B. A blank sample was prepared using MilliQ water at a volume equivalent to the sample volume. Samples were allowed to stand for at least one minute before measuring the UV absorbance at 351 nm. The hydrogen peroxide concentration was determined using the Equation 3.13.

[ ( ) ] [H2O2] (mg/L) = (3.13)

3.3.6 Incident Irradiance of the LP Lamp using Iodide/Iodate Actinometry

Iodide/Iodate actinometry was used to determine the incident irradiance of the 40W low- pressure UV lamp with the parabolic concentrator installed in the collimated beam apparatus. Experiments were performed in duplicate using the procedure outlined by Rahn (1997). A 10 mL aliquot of iodide/iodate solution was placed in a petri-dish at 39.70 cm from the lamp. The petri factor was 0.974 and the water path length was 0.53 cm. A volume of 1 mL was collected from the irradiated solution for analysis and measurement of UV absorbance at 510 nm. Experiments were performed in duplicate. The incident irradiance of the LP lamp was determined to be (0.12 ± 0.006) mW/cm2

3.3.7 Incident Irradiance of the MP Lamp using Ferrioxalate Actinometry

Incident irradiance of the 1kW medium pressure UV lamp in the collimated beam apparatus was determined using ferrioxalate actinometry for the wavelength range of 200 – 300 nm according to the procedure outlined by Bolton et al. (2011) and Goldstein et al. (2008). A 10 mL aliquot of the ferrioxalate solution was placed in a petri-dish at 39.70 cm from the lamp. The petri factor was 0.974 and the water path length was 0.53 cm. The top of the petri-dish was covered with an opaque petri-dish cover containing a small hole in the centre with a diameter of 1.5 cm as described by Bolton et al. (2011). This smaller aperture was used to ensure that the UV light enters the petri-dish uniformly thereby improving the accuracy of the measurements (Bolton et al., 2011). The sample was exposed for 2 minutes. A volume of 1 mL was collected from the irradiated solution for analysis and measurement of UV absorbance at 510 nm. Experiments

57 were performed in duplicate. The incident irradiance of the MP lamp was determined to be (3.94 ± 0.03) mW/cm2

3.3.8 Fluorescence Excitation Emission Matrices Analysis

Fluorescence excitation-emission (FEEM) matrices of the samples were collected using a Perkin Elmer LS-50B Fluorescence Spectrofluorometer. The Stokes Shift method as outlined in Peiris et al., (2009) was used to optimise the instrument parameters for FEEM analysis. In this method, an untreated effluent sample is analysed on the spectrofluorometer to obtain a 3D FEEM spectra. The maximum peak of the spectra was identified. The intensity values versus the excitation and emission wavelengths for the peak maxima were plotted. The Stoke’s Shift is the wavelength difference between the peaks for the emission and excitation spectra plot.

The instrument conditions are: photomultiplier voltage (PMT) = 775 V, scan rate = 600 nm/min, and an excitation/emission slit width = 10 nm. Samples are analysed by scanning 301 individual emission spectra (300 – 600 nm) at sequential 10 nm increments of excitation wavelengths between 200 – 500 nm using a quartz cuvette with 4 optical windows and a 1 cm pathlength.

Given the complex nature of wastewater samples, it is possible that constituents in the matrix may reduce the fluorescence intensity readings (quenching effects) of the organic matter. In verifying that quenching effects would not be a concern during FEEM analysis, serial dilutions of the wastewater effluent were prepared and the single intensity scans of these diluted samples were performed. A linear correlation between the intensity values and the organic carbon concentration of the diluted sample was obtained verifying that there were no quenching effects.

The FEEM matrices were used to generate 3-D spectra to provide qualitative information on the humic acid, fulvic acid, protein-like materials and colloidal/particulate matter present in the EfOM of the samples based on fluorescence intensities (Peiris et al., 2010). Raman scattering was eliminated and other background noise was reduced by subtracting the fluorescence spectra for ultra-pure Milli-Q water from all the sample fluorescence spectra. Analysis of the FEEM spectra and qualitative identification of the different EfOM components were done using two

58 multivariate data analysis method: principal component analysis (PCA) and fluorescence regional integration (FRI). These are statistical decomposition methods used for analysing multi-dimensional data (Baghoth et al., 2011; Bro 1997). The MATLAB 7.3.0 computational software was used for data processing with the PLS Toolbox 3.5 for PCA analysis. PCA allows the extraction of key principal components from the data sets which can be used to describe any trends in the original spectra data (Peiris et al., 2010).

Fluorescence regional integration (FRI) was used to quantify the peak maxima by delineating the matrices into four defined excitation-emission (Ex/Em) regions and quantifying the volume under the peak maxima for the four fractions. The regions were identified based on the excitation-emission (Ex/Em) of the peak maxima ±5 nm for both excitation and emission. Regions were classified as humic acid (Ex/Em = 320/425) (Chen et al., 2003), fulvic acid (Ex/Em = 240/425) (Chen et al., 2003), tryptophan (Ex/Em = 280/350) (Murphy et al., 2008) and tyrosine (Ex/Em = 280/305) (Murphy et al., 2008).

3.3.9 Total Organic and Inorganic Carbon Analysis

The total organic and inorganic carbon concentrations of the wastewater samples were determined using the Standard Method 5310 D: Wet-Oxidation Method on the TOC Analyzer (O.I. Analytical Aurora Model 1030 with auto-sampler Model 1088) (APHA 2005). Triplicate analyses were done for each sample on the equipment. A 6-point calibration graph was used to determine the TOC and TIC concentration as shown in Figure 3-9 and Figure 3-10 respectively. A 3 mg/L check standard of potassium hydrogen phthalate was used for quality control in each experimental run.

60000

50000 Area Count = 4669.8 [TOC] (mg/L) + 892.86 R² = 0.9996 40000

30000

20000 Average Area Average Count for TOC 10000

0 0 2 4 6 8 10 12 Concentration (mg/L) Figure 3-9: Typical calibration graph for total organic carbon concentration

50000

40000 Area Counts = 4665 [TIC](mg/L) - 353.37 R² = 0.9986

30000

20000 Area Counts Counts Area for TIC

10000

0 0 2 4 6 8 10 Concentration (mg/L) Figure 3-10: Typical calibration graph for total inorganic carbon concentration

3.3.10 Alkalinity

The total alkalinity of the effluents was determined using the Standard Method 2320 B (APHA, 2005). A sample volume of 100 mL was placed in a 250 mL beaker with a stir bar. Fifteen drops of bromocresol green indicator were added and this was titrated with 0.02 N H2SO4 until the colour changed from blue to yellow. Alkalinity tests were performed in triplicate. The total alkalinity of the sample was determined using Equation 3.14.

( ) (3.14)

where A – volume of H2SO4 titrated (mL), N – normality of H2SO4 (0.02N), V – volume of the sample used (100 mL)

3.3.11 Acute Toxicity

Acute toxicity was measured using 96h-LC50 Rainbow trout bioassay and 48h-LC50 Daphnia magna bioassay. For both tests, a volume of 80 L of wastewater was collected in plastic sampling containers lined with food-grade polyethylene bags which were filled and sealed without any headspace within 24 hours of sample collection. All testing commenced within 48 hours of sample collection.

48h-LC50 Daphnia magna analysis was performed in accordance with the protocol developed by

MOE (1988). For the 48h-LC50 Daphnia magna analysis, culture conditions of moderately hard water (120 – 250 mg/L), pH between 6.6 – 8.5, and a temperature of 20 oC was used. Healthy neonates of less than 24 hours old were used. Samples were prepared in quadruplicates with serial dilutions of the wastewater. The 48-h LC50 values were determined as the concentration (dilution) that caused 50% mortality of the neonates over the 48 hour period.

The 96h-LC50 Rainbow trout analysis is a 96-h static test conducted in accordance with the Biological Test Method outlined by Environment Canada (2007). Test fish are added to 100% concentration of the wastewater and the percentage mortality after 96-h is determined.

3.3.12 Genotoxicity

Genotoxicity of the effluents were determined using the SOS-Chromo Test™ kit (EBPI, Brampton, Canada). The test uses the genetically engineered Escherichia coli (E. coli) PQ37 strain to measure the SOS response of the bacterial promoter gene that occurs in the cells due to changes in the cell DNA when exposed to chemicals. This promoter gene is responsible for - galactosidase (β-gal) production, where β-gal induction is proportional to the amount of DNA damage that occurs and can be assessed using a chromogenic substrate to form a blue colour. For analysis of the wastewater, a 1L sample was filtered through a 0.45 µm PES membrane filter and acidified to pH 2 using concentrated sulphuric acid. Solid phase extraction (SPE) with the Oasis hydrophilic-lipophilic balance (HLB) extraction cartridges (Waters Oasis HLB 12 cc, 500 mg) were used to concentrate the sample. Serial dilutions of the SPE concentrated effluent samples were set-up in a 96-well plates with a known carcinogenic positive control (4-NQO) and a negative control (10% DMSO saline) in each plate. Plates were incubated for 2 hours at 37oC following which chromogen for β-gal and alkaline phosphatase (AP) were added to each well, and the plates incubated for an additional 1 hour. The optical density of the plates was measured at 605 nm (β-gal activity) and 420 nm (AP activity) using a Tecan Infinite M200 Plate reader (Morrisville, NC). Induction factors (IF) and toxicity equivalency values (TEQ) relative to the genotoxicity of 4-NQO were used in assessing the genotoxicity of the samples.

3.3.13 Estrogenicity

The estrogenicity of the samples was measured using the Yeast Estrogenicity Screening (YES) assay at the Yargeau Laboratory located at McGill University, Montreal, Canada. The YES assay was conducted as previously described (Routledge and Sumpter 1996) with minor modifications. The 96-well plates were incubated for 4 days before measuring the absorbance at 540 nm with a correction for turbidity at 630 nm using a BIO RAD Benchmark Plus equipped with Microplate Manager 5.2.1. The degree of colour change induced by successive dilutions of the test chemicals provided a measure of their estrogenic potency. The validity of the assay was confirmed by the response of the reference compound used (E2), that showed a median effective concentration (EC50) of 2.16 x10-10 M (n=13). Replicates of the samples were evaporated and re-dissolved in assay medium and tested (n=3) in a range of 12 dilutions (1:2)

62 with a row of ethanol blanks and E2 standards in each plate. Figure 3-11 shows one example of the colour change that occurred in the 96-well plate, and Figure 3-12 is an example of a dose- response curve for one sample.

Figure 3-11: Colour change in the 96-well plate for estrogenicity analysis

Figure 3-12: Dose-response curve for estrogenicity analysis

3.4 References

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Autin, O., Hart, J., Jarvis, P., MacAdam, J., Parsons, S.A., Jefferson, B. (2013) The impact of background scavenging capacity and alkalinity on the degradation of the pesticide metaldehyde by two advanced oxidation process: UV/H2O2 and UV/TiO2. Water Research 47(6): 2041 – 2049.

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Bolton, J.R., Stefan, M.I., Shaw, P-S., Lykke, K.R. (2011) Determination of the quantum yield of the ferrioxalate and KI/KIO3 actinometers and a method for the calibration of radiometer detectors. Journal of Photochemistry and Photobiology. A, Chemistry. 222: 166 – 169.

Bolton, J.R., Cater, S.R. (1994) Homogeneous photodegradation of pollutants in contaminated waters. In: Aquatic and Surface Photochemistry. Boca Raton, Florida: Lewis Publishers, pp. 467 – 490.

Bolton, J.R., Valladares, J.E., Zanin, J.P., Cooper, W.J., Nickelson, M.G., Kajdi, D.C., Waite, T.D., Kurucz, C.N. (1998) Figures-of-merit for advanced oxidation technologies: a comparison of homogeneous UV/H2O2, heterogeneous TiO2 and electron beam processes. Journal of Advanced Oxidation Technologies 3(2): 174 – 181.

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Chen, P-J., Rosenfeldt, E.J., Kullman, S.W., Hinton, D.E., Linden, K.G. (2007) Biological assessments of a mixture of endocrine disruptors at environmentally relevant concentrations in water using UV/H2O2 oxidation. Science of the Total Environment 376(1-3): 18 – 26.

Dong, M., Mezyk, S.P., Rosario-Ortiz, F.L. (2010) Reactivity of effluent organic matter (EfOM) with hydroxyl radical as a function of molecular weight. Environmental Science & Technology 44(15): 5714 – 5720.

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Goldstein, S., Rabani, J. (2008) The ferrioxalate and iodide-iodate actinometers in the UV region. Journal of Photochemistry and Photobiology A: Chemistry 193(1): 50 – 55.

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Rosario-Ortiz, F.L., Wert, E.C., Snyder, S.A. (2010) Evaluation of UV/H2O2 treatment for the oxidation of pharmaceuticals in wastewater. Water Research 44(5): 1440 – 1448.

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Rosenfeldt, E.J., Linden, K.G., Canonica, S., von Gunten, U. (2006) Comparison of the efficiency of •OH radical formation during ozonation and the advanced oxidation process

O3/H2O2 and UV/H2O2. Water Research 40(20): 3695 – 3704.

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Vogna, D., Marotta, R., Napolitano, A., Andreozzi, R., d'Ischia, M. (2004) Advanced oxidation of the pharmaceutical drug diclofenac with UV/H2O2 and ozone. Water Research 38(2): 414 – 422.

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Chapter 4 A Comparative Study of the Hydroxyl Radical Scavenging Capacity of Secondary Wastewater Effluents

Abstract

Hydroxyl radical scavengers can reduce the effectiveness of UV-based advanced oxidation processes (UV-AOPs) when treating contaminants in municipal wastewaters. This study evaluated the scavenging characteristics of wastewater from five membrane bioreactor (MBR) and five activated sludge (AS) treatment systems. The MBR and AS waters were found to be significantly different at a 90% confidence interval in terms UV254 absorbance, alkalinity, and biopolymer concentration. Effluent organic matter (EfOM), with an average kOH,EfOM of (2.75 ± 1.04) x 108 M-1s-1, was identified as the primary hydroxyl scavenger contributing to more than 70% of the background scavenging, except in cases with nitrite concentrations greater than 0.3 - mg NO2 -N/L. In general, the overall scavenging capacity, EfOM scavenging capacity, and the EfOM reaction rate constant of the AS wastewater exceeded that of the MBR. However, due to the small sample size (n=5) and considerable variability in scavenging characteristics among the MBR wastewaters, the differences in EfOM reactivity between the two wastewater types were not found to be statistically significant at a 90% confidence interval. Nevertheless, these preliminary findings imply that MBR wastewaters may be more amenable to treatment using UV-AOPs. This ease of treatment may be related to the retention/rejection of biopolymers on the membrane surface as EfOM scavenging capacity was found to have a strong positive correlation with biopolymer concentration.

4.1 Introduction

Advanced oxidation processes (AOPs) involve the generation of hydroxyl radicals (OH), which are very effective for oxidising trace organic components in water and wastewaters due to their highly reactive and non-selective nature (Andreozzi et al., 2003 and 2004; Rosenfeldt et al., 2004; Vogna et al., 2004; Yuan et al., 2009). OH radicals, however, are scavenged by non- target components which reduce the effectiveness of an AOP as it reduces the steady-state concentration of OH radicals available for oxidising the target compounds. This is of particular concern for wastewater treatment, where AOPs have not been readily implemented due to the high dose and energy requirements. Given the known effectiveness of AOPs for treating recalcitrant organic compounds such as micropollutants that can have deleterious effects on aquatic organisms and ecosystems, understanding the factors that influence scavenging capacity is important. The main scavenging components in wastewater are expected to include effluent organic matter (EfOM), nitrite, bicarbonate and carbonate ions (Keen et al., 2014; Rosario-Ortiz et al., 2010; Rosenfeldt et al., 2007).

Previous studies on wastewaters have evaluated the hydroxyl radical reaction rate constant of bulk EfOM and isolated EfOM fractions, and factors that influence bulk EfOM reactivity (Dong et al., 2010; Keen et al., 2014; McKay et al., 2011; Rosario-Ortiz et al., 2008; Souza et al., 2014). EfOM’s reactivity with OH is influenced by bulk physico-chemical characteristics such as molecular weight, specific UV absorbance, fluorescence index, and polarity. Keen et al. (2014) and Souza et al. (2014) also evaluated the scavenging capacity of effluents from activated sludge treatment systems noting that EfOM is the primary hydroxyl radical scavenger. Given the impact of EfOM as an OH scavenger and the potential variability in its reactivity, an understanding of the components or characteristics of EfOM that are responsible for its scavenging capabilities is important to reduce the influence of EfOM as a scavenger and to improve the effectiveness of AOPs.

Activated sludge (AS) and membrane bioreactor systems (MBR) are two systems used for treating municipal wastewaters. Previous comparative studies have shown that MBR effluents often have lower concentration of organic carbon and an overall higher removal efficiency of

71 micropollutants (Klatt et al., 2003; Kimura et al., 2005; Radjenovic et al., 2007; DeWever et al., 2007). Influent wastewater characteristics and plant operating conditions will differ for every wastewater system resulting in variability among effluents irrespective of the type of treatment process. Nevertheless, given the potential influence of EfOM on scavenging capacity, there may be a useful generalization that could be made about the suitability of MBR systems relative to AS systems for subsequent AOP treatment to remove micropollutants.

Ultraviolet (UV) light is increasingly being used for final disinfection of wastewater effluents, replacing chlorine as a disinfectant due to the formation of disinfection by-products. For plants using UV disinfection, implementing UV-AOP for dual disinfection and treatment of micropollutants could be a viable transition. Therefore, understanding the variability in the scavenging capacity and EfOM reactivity, and identifying the particular characteristics or components that are responsible for the scavenging potential exerted in wastewaters would be useful in assessing effluent types that may be more amenable to UV-AOP treatment.

The objectives of this study were to determine the contribution of known OH scavengers to the overall background scavenging capacity of wastewater matrices; to evaluate and compare the water quality and scavenging characteristics of effluents from AS and MBR systems to determine whether the wastewater from one type of treatment may consistently be more easily treated using UV-based AOPs; to determine the EfOM reaction rate constants for bulk EfOM; and to evaluate which specific EfOM components or water quality characteristic can be directly correlated to the scavenging capacity of EfOM. To address these objectives, the OH radical scavenging characteristics and EfOM composition of secondary effluents from five AS and five MBR systems located in Ontario, Canada were measured.

4.2 Materials and Methods

4.2.1 Wastewater Effluents

Samples from the five AS and five MBR municipal wastewater plants were collected after the secondary clarifier or membranes, but prior to any further filtration/disinfection for all the plants. Samples were collected in pre-cleaned 20 L polypropylene containers and stored at 4oC until the experiments were conducted. Comprehensive characterization of the wastewater was done within 48 hours of collection where the samples were analysed for pH, temperature, UV absorbance at 254 nm (UV254), specific UV absorbance (SUVA), dissolved organic carbon concentration (DOC), total organic carbon (TOC), total inorganic carbon (TIC) concentration, anion concentrations (chloride, nitrate, nitrite, sulphate, phosphate) and alkalinity. The water quality characteristics of the ten wastewaters are shown in Table 4-1.

The EfOM was also characterised using liquid chromatography organic carbon detection (LC- OCD) into five organic fractions of biopolymers, humics, building blocks of humics, low molecular weight (LMW) neutrals and LMW acids, according to methods described in the following sections.

4.2.2 Measurement of Scavenging Capacity and kOH,EfOM

The overall scavenging capacity, EfOM scavenging and hydroxyl radical reaction rate constants with EfOM were determined for each wastewater. The overall scavenging capacity was determined by monitoring the degradation of a known OH probe compound, methylene blue, at a given hydrogen peroxide (H2O2) concentration and UV fluence (Rosenfeldt et al., 2007). The EfOM scavenging and reaction rate constant were subsequently determined by subtracting the scavenging potential of known scavengers from the overall measured scavenging, and dividing EfOM scavenging by the molar mass of organic carbon (12 g C per mole), respectively.

Table 4-1: Water quality of the wastewater effluents used in the study

Parameter AS1 AS2 AS3 AS4 AS5 MBR1 MBR2 MBR3 MBR4 MBR5 pH 7.4 7.3 7.0 6.8 6.6 7.7 6.7 7.7 7.4 7.1

o Temperature ( C) 21 19 20 18 17 15 20 18 19 21 UVA @ 254 nm (cm-1) 0.117 0.119 0.125 0.13 0.15 0.064 0.114 0.099 0.076 0.13

-1 % UVT (cm ) 76 81 75 76 71 86 77 80 84 73 SUVA (L/mg-m) 2.54 2.20 2.36 1.80 1.78 2.03 1.95 1.89 2.16 2.01 254 TOC (mg/L-C) 4.32 4.41 4.12 6.84 6.29 3.39 6.08 4.54 2.95 6.74 DOC (mg/L-C) 4.26 4.13 3.83 6.67 5.63 3.15 5.78 4.54 2.95 6.66

TIC (mg/L-C) 39.4 26.5 34.2 24.4 16.9 17.3 7.10 27.5 14.8 15.4 2- Carbonate (mg/L CO3 ) 0.14 0.08 0.06 0.02 0.01 0.17 0.01 0.22 0.07 0.03 - Bicarbonate (mg/L HCO3 ) 126 88 115 78 52 70 28 86 51 49

Total Alkalinity 207 144 188 128 86 115 45 144 85 81 (mg CaCO3/L) a a a a a a a a Nitrite (mg/L-N) < 0.02 < 0.02 0.06 < 0.02 < 0.02 < 0.02 0.27 < 0.02 < 0.02 < 0.02 Nitrate (mg/L-N) 26 8 24 15 17 16 27 13 37 10 a a a a a a a a a Phosphate (mg/L) < 0.07 < 0.07 < 0.07 < 0.07 < 0.07 < 0.07 0.30 < 0.07 < 0.07 < 0.07 Sulphate (mg/L) 136 84 75 67 61 72 43 34 158 26 Chloride (mg/L) 382 76 283 293 435 70 136 148 337 180 aMethod detection limit Values are the averages of n= 2 or n= 3 samples

A Calgon Carbon® quasi-collimated beam apparatus (Model PSI-I-120, Calgon Carbon Corporation, USA) equipped with a 40W low-pressure mercury lamp and a parabolic concentrator (Calgon Carbon Corporation, Pittsburgh, PA, USA) was used. Incident irradiance (mW/cm2) was determined using iodide-iodate actinometry (Rahn 1997). The average fluence rate was determined using the correction factors outlined by Bolton et al. (2003) and the Bolton® Excel Spreadsheet for fluence calculations using a low-pressure lamp with a suspension depth of less than 2 cm (Bolton 2004).

Wastewater samples were spiked with 3.0 µM of methylene blue (MB). A volume of 10 mL of the spiked sample was placed in a petri-dish. A small stir bar was added to ensure that the sample was well mixed and homogeneous, without creating a vortex or disturbing the surface of the sample. Hydroxyl radical production was achieved by spiking hydrogen peroxide at concentrations ranging from 5 mg/L to 20 mg/L into the sample prior to irradiation. Samples were subsequently irradiated using the collimated beam apparatus for the exposure times required to obtain UV fluences ranging from 0 to 200 mJ/cm2. The MB and the hydrogen peroxide concentrations were determined using UV absorbance at 664 nm and the Klassen et al. (1994) triiodide method respectively. All irradiations were performed in duplicate. The photolytic degradation rate of methylene blue due to direct UV photolysis only (i.e.,, in the absence of H2O2) was also monitored to ensure that the observed methylene blue degradation during UV/H2O2 treatment was solely due to reactions with the hydroxyl radical.

The ROH,UV concept outlined by Rosenfeldt et al. (2007) was used for determining the  scavenging capacity of the wastewaters. The ROH,UV is defined as the OH exposure per UV fluence, which is the steady-state concentration of OH present in the wastewater per unit dose of UV for an initial concentration of hydrogen peroxide. Hence, a wastewater with a higher background OH scavenging will have a lower steady-state concentration of OH per unit UV

(i.e., a lower ROH,UV). ROH,UV is measured by monitoring the degradation of methylene blue (MB) and applying Equation 4.1:

∫[ ] (4.1)

where is the observed degradation rate constant of MB at a given H2O2 concentration, is the degradation rate constant of MB due to direct UV photolysis only (i.e., at 0 mg/L H2O2), 10 -1 -1 is the MB hydroxyl radical reaction rate constant (2.1 x 10 M s ) (Buxton et al., 1988), and H is UV fluence (mJ/cm2).

Using the ROH,UV values and the Glaze et al. (1995) UV/H2O2 kinetic model based on steady-  state concentration of OH radicals, the overall background scavenging capacity (∑ [ ] ) of each wastewater is determined using Equation 4.2.



∑ [ ] [ ] (4.2)

where m is the gradient of a plot of versus , is the hydrogen peroxide molar [ ] -1 -1 absorption coefficient, which is equal to 18.7 M cm at 254 nm (Bolton et al., 1994), OH is the quantum yield of the production of •OH radicals from hydrogen peroxide photolysis which is 1.0 -1 mol E (Rosenfeldt et al., 2007), U254 is the energy per 1 mol of photons at 254 nm which is 5 4.72 x 10 J/mol (Bolton 2001), is the MB hydroxyl radical reaction rate constant (2.1 x 1010 M-1s-1) (Buxton et al., 1988) and [MB] is the molar concentration of MB.

The EfOM scavenging capacity is estimated by subtracting the sum of the scavenging potential of known scavenging constituents present in the matrix other than EfOM from the measured overall background scavenging capacity, as shown in Equation 4.3.

[ ] ∑ [ ] ( [ ] [ ] [ ]) (4.3)

10 -1 -1 6 -1 -1 where = 1.0 x 10 M s (Buxton et al., 1988), = 8.5 x 10 M s (Buxton et al., 8 -1 -1 1988), and = 3.9 x 10 M s (Buxton et al., 1988).

The known scavengers in the wastewaters, other than EfOM, were considered to be bicarbonate, carbonate and nitrite (Keen et al., 2014; Rosario-Ortiz et al., 2010; Rosenfeldt et al., 2007). Other anions (nitrate, sulphate, chloride and phosphate) can react with the OH radical, however their contribution to the overall scavenging capacity for the wastewaters in this study was calculated to be less than 1%, and therefore these anions were not considered (Buxton et al., 1988). Chloride anions can react with OH, but the reaction mechanism reportedly leads to the re-formation of the chloride and OH (Liao et al., 2001). Control tests (data not shown) confirmed the lack of chloride scavenging at concentrations similar to those in the wastewaters tested. Nitrate is only a concern when medium-pressure UV lamps are used, as direct UV photolysis at low wavelengths (< 250 nm) converts nitrate to nitrite (Sharpless et al., 2003) thereby significantly increasing the background scavenging capacity of the effluent. In this work, low-pressure UV lamps were used to avoid potential nitrite formation.

The hydroxyl radical reaction rate constant of EfOM ( ) was subsequently determined by dividing the EfOM scavenging capacity by the molar organic carbon concentration in the -1 -1 matrix. EfOM rate constants are reported on a per mol of carbon basis (Mc s ) assuming 12 g per mol of C.

4.2.3 Analytical Equipment and Methods pH was measured using a calibrated pH meter (Thermo Scientific Orion Star A111). Dissolved organic carbon (DOC), total organic carbon (TOC) and total inorganic carbon (TIC) concentrations were measured using a TOC analyser (O.I. Analytical Aurora Model 1030 with auto-sampler Model 1088) in accordance with Standard Method 5310C (APHA 1998). The anion concentrations (chloride, nitrate, nitrite, sulphate and phosphate) in the effluent were measured in accordance with USEPA Method 300.1 (USEPA 1997) using a Dionex ion chromatography system equipped with an RFIC IonPac (Dionex AS9-HC, 4 x 250 mm) analytical column and a Dionex AG9-HC 4 x 50mm guard column, a CD20 conductivity detector, GP40 gradient pump, and an AS40 automated sampler. UV absorbance was measured using an Agilent 8453 UV-VIS spectrophotometer. Alkalinity was measured according to Standard Method 2320B (APHA, 1998). The organic constituents of the effluent organic matter

(biopolymers, humics, building blocks of humics, low molecular weight (LMW) neutrals and LMW acids) were characterised using liquid chromatography organic carbon detection analysis based on the protocol developed by Huber et al. (2011). Initial hydrogen peroxide concentrations were measured using the triiodide method (Klassen et al., 1994).

4.2.4 Chemicals and Reagents

All the chemical reagents (sodium persulphate, phosphoric acid, sodium chloride, sodium nitrate, sodium nitrate, potassium sulphate, potassium dihydrogen phosphate, methylene blue, 30 wt.% hydrogen peroxide) used in the study were obtained from Sigma Aldrich, Canada, with a purity of greater than 98%. Stock solutions were prepared as required in Milli-Q water generated from a Milli-QUV Plus Ultrapure Water System.

4.3 Results and Discussion

4.3.1 Assessing the Contributors to Overall Scavenging Capacity

EfOM, measured using dissolved organic carbon (DOC) concentration as a surrogate, was the primary OH scavenger in 9 of the 10 wastewaters accounting for 68%-96% of the overall scavenging capacity (with an average of 87%) (Table 4-2). This average is similar to the 80% reported by Souza et al. (2014). Ranking of the scavenging contributors showed that EfOM and bicarbonate were the major scavengers such that these two alone contributed to 98%-99.9% of the observed overall scavenging capacity in these 9 wastewaters. On average, the order of the scavenging components was EfOM (87%) > bicarbonate (10%) > nitrite (3%) > carbonate (0.4%) > nitrate (0.1%). However, nitrite can supplant EfOM as the primary scavenger in the presence of high nitrite concentrations as observed for MBR2 (Table 4-2) where nitrite scavenging was approximately 3 times that of EfOM (approximately 75% versus 25%).

Nitrite concentration in the effluents of nitrifying wastewater treatment plants is typically - negligible. However, both AS3 and MBR2 had high concentrations of 0.06 mg NO2 -N/L and - 0.27 mg NO2 -N/L, respectively, which may be due to operational problems in the plants at the

78 time of sampling. The elevated concentrations coupled with the nitrite hydroxyl radical reaction rate constant that is two orders of magnitude greater than that of EfOM contributes to the high scavenging potential exerted by nitrite in these instances. This observation is important as it demonstrates that nitrite can be a significant hydroxyl scavenger if its concentration is elevated due to plant operations. It also underscores a potential risk with using medium pressure (MP) UV lamps for treatment. MP lamps emit light in the 195 – 250 nm range, which can convert nitrate to nitrite (Goldstein et al., 2007; Sharpless et al., 2003). This suggests that low pressure UV would be preferred for UV-AOP treatment of wastewaters if nitrate may be present, or MP lamps with quartz sleeves that are treated to block emission of the lower wavelengths, assuming that all other factors are equal.

Table 4-2: Contribution of known scavengers to overall background scavenging capacity

kEfOM Contribution to Overall Background Scavenging Capacity (%) Plants (x 108) -1 -1 (Mc s ) EfOM Nitrate Nitrite Carbonate Bicarbonate AS1 3.27 87 0.1 0.0 0.1 13 AS2 4.33 92 0.0 0.0 0.0 7.6 AS3 3.85 68 0.1 23 0.2 8.8 AS4 2.14 91 0.1 0.0 0.1 8.4 AS5 1.61 91 0.1 0.0 0.1 8.8 MBR1 2.55 84 0.2 0.0 1.6 14 MBR2 1.38 25 0.1 74 0.0 1.4 MBR3 2.35 87 0.1 0.0 1.4 12 MBR4 2.06 87 0.5 0.0 0.1 12 MBR5 2.99 96 0.0 0.0 0.1 4.0

In order to supplement the observations for the 10 wastewaters used in our study, water quality data for an additional 23 secondary wastewaters were sourced from the literature (Keen et al., 2014; Rosario-Ortiz et al., 2010) (Appendix F) and combined for a total of 33 plants (Table 4-3). The additional 23 waters are secondary effluents from activated sludge treatment plants that were collected after secondary clarification, but prior to any filtration or disinfection.

Table 4-3: Contribution of known scavengers in 33 secondary wastewater effluents

Contribution to Overall Background Scavenging Capacity (%) Plant EfOM Nitrate Nitrite Carbonate Bicarbonate 1 63 0.0 34 0.1 3.4 2 95 0.1 0.0 0.0 5.0 3 80 0.0 16 0.0 4.1 4 83 0.0 12 0.0 4.8 5 88 0.0 7.5 0.0 4.9 6 96 0.1 0.0 0.0 3.7 7 93 0.0 0.0 0.0 7.1 8 95 0.0 0.0 0.0 5.4 9 77 0.0 19 0.0 4.1 10 76 0.0 19 0.0 5.3 11 85 0.0 12 0.0 2.8 12 91 0.1 0.0 0.0 9.4 13 81 0.0 9.2 0.0 9.8 14 78 0.0 11 0.0 10 15 97 0.1 0.0 0.0 2.5 16 77 0.0 21 0.0 1.9 17 65 0.0 29 0.0 6.2 18 96 0.0 0.0 0.0 4.1 19 97 0.0 0.0 0.0 2.6 20 96 0.0 0.0 0.0 3.6 21 75 0.0 18 1.9 5.3 22 45 0.0 54 0.0 1.6 23 29 0.0 68 0.3 2.8 24 87 0.1 0.0 0.1 13 25 92 0.0 0.0 0.0 7.6 26 68 0.1 23 0.2 8.8 27 91 0.1 0.0 0.1 8.4 28 91 0.1 0.0 0.1 8.8 29 84 0.2 0.0 1.6 14 30 25 0.1 74 0.0 1.4 31 87 0.1 0.0 1.4 12 32 87 0.5 0.0 0.1 12 33 97 0.0 0.0 0.1 3.0 Note: Plants 1-20 were sourced from Keen et al. (2014); plants 21-23 are sourced from Rosario-Ortiz et al. (2010); plants 24–28 and 29-33 are the AS and MBR plants respectively of this study

Using literature OH reaction rate values for bicarbonate, carbonate, nitrite and nitrate, and an average value of the OH reaction rate constants for EfOM from the 10 wastewaters used in our study, the contribution of each component to the overall scavenging capacity in the additional 23 wastewaters was determined (Table 4-3). Similar observations were noted for all 33 plants with EfOM accounting for 87%-97% of the overall •OH scavenging for wastewaters where the nitrite concentration was negligible (i.e., 0 mg/L or below method detection limits), and 25%- 88% for effluents where the nitrite concentration was measurable.

4.3.2 Comparing MBR and AS Scavenging Characteristics

Overall scavenging capacity is influenced by both organic and inorganic constituents in the wastewater (Table 4-2). However, the inorganics, particularly the bicarbonate anions of the AS and MBR wastewaters, were not significantly different based on a Mann Whitney U test at a 90% confidence interval. Furthermore, the bicarbonate anions account for only 9% of the overall scavenging capacity on average for both the MBR and AS wastewaters. Hence, the similarity in the average contribution of the bicarbonate ions for both wastewaters indicates that potential differences in the overall scavenging capacity of the MBR and AS wastewaters are likely not due to the inorganics, but the organic composition and characteristics.

The AS and MBR wastewaters had similar overall average scavenging capacities (to within 6% of each other), however, these results may be skewed by the very high nitrite scavenging in the MBR2 plant (Table 4-4). When considering only the contribution of the EfOM to the scavenging capacity, the average scavenging in the AS wastewaters exceeded that of the MBR by 22%. Although, dissolved organic carbon concentration (DOC) is used as a surrogate for EfOM concentration, this study found that EfOM scavenging capacity was poorly correlated with DOC (R2 = 0.2) for both the AS and MBR wastewaters (Appendix G). This poor correlation indicates that EfOM scavenging capacity reflects the inherent characteristics and reactivity of EfOM with the OH radical, rather than bulk DOC.

The EfOM hydroxyl radical reaction rate constants (kOH,EfOM) for all 10 wastewaters ranged 8 -1 -1 8 -1 -1 from 1.38 x 10 Mc s to 4.33 x 10 Mc s which is similar to values from previous studies

(Dong et al., 2010; Katsoyiannis et al., 2011; Rosario-Ortiz et al., 2008). The average for the AS 8 -1 -1 8 -1 -1 wastewaters was 3.0 x 10 Mc s , while the MBRs had an average of 2.47 x 10 Mc s which were both lower than some of the reaction rate constants reported for natural organic matter in 8 -1 -1 8 -1 -1 the environment of 3.6 x 10 Mc s (Westerhoff et al., 1999), 3.8 x 10 Mc s (Li et al., 2008) 8 -1 -1 and 3.0 x 10 Mc s (Rosenfeldt et al., 2006).

This study therefore shows that the EfOM of the AS wastewaters had a higher average scavenging capacity by virtue of its higher OH reactivity when compared to the MBR wastewaters, rather than a higher EfOM concentration. It must be cautioned, however, that given the small sample size of five random plants for each type of wastewater, these differences were not found to be statistically significant based on a Mann-Whitney U test at a 90% confidence interval. Nonetheless, the data suggests that AS wastewaters may generally have a greater scavenging capacity than MBR wastewaters, but this is dependent on the EfOM reactivity. Overall background scavenging capacity and EfOM scavenging capacity are not typically reported in the literature. Hence, evaluating the scavenging characteristics of additional plants would be useful in verifying if the observed trend holds true. This would be useful as it can indicate with more certainty whether MBR wastewaters may be more amenable to AOP treatment than AS wastewaters due to a lower level of scavenging.

Table 4-4: Scavenging characteristics of the secondary wastewater effluents

a c Overall Effluent b kOH,EfOM EfOM Scavenging 8 Plant Scavenging 5 -1 (x 10 ) 5 -1 (x 10 ) s -1 -1 (x 10 ) s Mc s AS1 1.23 ± 0.01 1.18 ± 0.30 3.27 ± 0.85 AS2 1.63 ± 0.01 1.59 ± 0.09 4.33 ± 0.24 AS3 1.77 ± 0.22 1.32 ± 0.02 3.85 ± 0.06 AS4 1.25 ± 0.01 1.22 ± 0.14 2.14 ± 0.24 AS5 0.862 ± 0.16 0.846 ± 0.16 1.61 ± 0.31 MBR1 0.748 ± 0.04 0.721 ± 0.004 2.55 ± 0.02 MBR2 2.67 ± 0.01 0.701 ± 0.42 1.38 ± 0.82 MBR3 0.931 ± 0.04 0.889 ± 0.18 2.35 ± 0.47 MBR4 0.527 ± 0.08 0.507 ± 0.11 2.06 ± 0.45 MBR5 2.26 ± 0.01 2.24 ± 0.01 2.99 ± 0.024 a,b,c Average ± standard deviation of duplicate experiments at 5 different H2O2 concentrations (n = 10)

4.3.3 Identifying Components that Influence EfOM Scavenging and kEfOM

In an effort to identify more specific information on how EfOM characteristics may affect radical scavenging, the EfOM samples were analysed by LC-OCD. A comparison of the five organic constituents (biopolymers, humics, building blocks of humics, LMW neutrals, LMW acids) showed that the biopolymer concentration on average was greater in the AS (0.44 mg/L) when compared to the MBR (0.078 mg/L) by a factor of 5.6. This difference was found to be significant at a 90% confidence interval using the Mann-Whitney U test.

3.5

3.0 AS1 AS2 2.5 AS3 2.0 AS4 AS5

1.5 MBR1 Concentration (mg/L) Concentration MBR2 1.0 MBR3 0.5 MBR4 MBR5 0.0 Biopolymers Humics Building blocks LMW neutrals LMW acids Figure 4-1: Distribution of the EfOM organic constituents in the wastewater effluents

Table 4-5: Spearman’s correlation coefficients for EfOM scavenging and kEfOM aCorrelation Coefficients Parameters EfOM Scavenging kOH,EfOM UV254 0.624 (p = 0.054) 0.079 DOC 0.345 -0.309 SUVA 0.212 0.721 (p = 0.019) Biopolymers 0.650 (p = 0.042) 0.201 Humics 0.345 -0.224 Building blocks 0.527 -0.055 LMW neutrals 0.333 0.139 LMW acids 0.467 0.006 aThe p values in parenthesis shows the correlations that are significant at 90% C.I.

Figure 4-1 shows that four constituents (humics, building blocks of humics, LMW neutrals and LMW acids) had a similar distribution in all ten effluents with differences ranging from 4-13%. Only the biopolymer component was significantly different between the two types of wastewaters. The low biopolymer content in the MBR wastewaters is expected as the MBR plants in this study all operated with ultrafiltration (UF) membranes. UF membranes typically reject biopolymers as the large molecular size of biopolymers usually exceeds the typical nominal pore size (0.01 – 0.10 µm) of the UF membrane (Haberkamp et al., 2011; Henderson et al., 2011; Zheng et al., 2010 and 2014). Another factor for lower biopolymer concentration in MBR wastewaters is the sludge retention times (SRT). MBR plants compared to AS plants usually have longer sludge retention times (SRTs) which has been shown to improve removal of high molecular weight components of EfOM (Gao et al., 2014). The observations from this study imply that the biopolymer constituent may be a major contributor to EfOM scavenging capacity. This correlates with the findings reported by Gonzalez et al. (2013) where the biopolymer component of the EfOM was preferentially and significantly degraded during

UV/H2O2 treatment over an exposure time of 30 minutes, such that degradation of other organic constituents (humics, building blocks of humics, LMW neutrals, LMW acids) was only observed after the biopolymer concentration had been reduced. Hence, it can be deduced that the lower concentration of biopolymers in the MBR effluent may be responsible for the lower EfOM scavenging capacity. This suggests that MBR effluents compared to AS effluent could be more easily treated using UV-AOPs for micropollutant degradation.

In identifying which of the bulk water quality parameters (UV254, DOC, SUVA) and/or the five LC-OCD organic constituents of EFOM may influence EfOM scavenging capacity and the EFOM hydroxyl radical reaction rate constant, a Spearman rank order correlation at 90% C.I. was performed. Overall EfOM scavenging was correlated only to UV254 and biopolymers, while the EfOM scavenging rate constant (kOH,EfOM) was correlated only to SUVA (Table 4-5). For the first correlation (EfOM with UV254), it is known that UV254 is an indicator of the electron density of the organic matter in a water matrix such that increasing UV254 values are an indication of an high concentration of organic constituents with electron-rich functional groups which can absorb a large number of photons. These electron-rich functional groups are usually aromatic conjugated double bonds and/or rings (Owen et al., 1995; Singer, 1999). The OH is known to preferentially react with the aromatic structures with conjugated double bonds, therefore a high concentration of these constituents in EfOM will result in increased reactivity or scavenging of OH.

EfOM scavenging was also found to be correlated to biopolymers. Biopolymers, as classified in LC-OCD analysis, refer to polysaccharides, proteins, and amino sugars, which are polymeric compounds that contain covalently bonded atoms with alternating single and double bonds, and molecular weights exceeding 20 kDa (Huber et al., 2011). The OH can preferentially react with conjugated double bonds through OH addition (Minakata et al., 2009). Since these OH reactions are diffusion controlled, there will also be a preferential initial reaction with large molecular weight constituents that contain conjugated double bond structures, as was demonstrated in the work conducted by Sarathy et al. (2007) and Gonzalez et al. (2013). Therefore, although the biopolymer concentration in EfOM is low in relation to other organic constituents, its molecular size and functional groups surpasses that of the other organic components increasing its reactivity with the OH, and will directly influence EfOM scavenging capability as scavenging potential is a function of both concentration and reaction rate. This also explains why the biopolymers were correlated with EfOM scavenging capacity while the same was not observed for kOH,EfOM. As the biopolymer was the only component that differed between the two wastewaters in this study, with the AS wastewater having both a higher biopolymer

85 concentration and EfOM scavenging capacity, it illustrates that for the wastewaters in this study, biopolymers are the influencing factor.

The third correlation between kOH,EfOM and SUVA is in agreement with the work conducted by Rosario-Ortiz et al. (2008) and Keen et al. (2014) who both identified SUVA as one of the bulk water quality parameters that is strongly correlated with kOH,EfOM. High SUVA values are indicative of a high concentration of aromatic compounds, and the hydroxyl radical will react with the electron-rich sites on the carbon typical of aromatic compounds through OH addition (Minakata et al., 2009; Sudhakaran et al., 2013). These addition reactions occur preferentially at close to diffusion-controlled rates (Minakata et al., 2009). Hence, one of the driving forces for the rate of reaction between EfOM and the OH is the prevalence of certain functional groups or chemical structures (i.e., conjugated double bonds or aromatic compounds). The implication of this study is that wastewaters with low concentrations of high molecular weight constituents such as biopolymers and less conjugated double bonds or aromatic structures demonstrated by low SUVA values should have the lowest hydroxyl scavenging capacity and may be most amenable to UV-AOP treatment.

4.4 Conclusions

Ten secondary wastewaters (from five activated sludge plants and five membrane bioreactor plants) were evaluated in this study to understand scavenging characteristics, differences between the water types, and the factor(s) that influence the scavenging potential observed in the wastewater matrices. The study showed that EfOM was normally the primary hydroxyl scavenger in wastewater matrices accounting for more than 70% of the overall background scavenging capacity in effluents from nitrifying plants, except in cases with high nitrite - concentrations (e.g., 0.3 mg NO2 -N/L in this study). While no effluent nitrite standards exist in Canada and the United States, nitrite in wastewater effluents are usually negligible, and any detected nitrite levels are usually indicative of operational problems. Other findings of the study are as follows:

1. On average, the overall scavenging capacity, the EfOM scavenging capacity, and the kEfOM of the AS wastewaters exceeded that of the MBR wastewaters. Nevertheless, due to considerable variability among the wastewaters, particularly the MBRs, the differences between the waters were not found to be statistically significant at a 90% confidence interval. 2. The hydroxyl radical reaction rate constants for EfOM of both the AS and MBR effluents 8 -1 -1 8 -1 -1 ranged from 1.38 x 10 Mc s to 4.33 x 10 Mc s based on the molar carbon concentration in the wastewater. 3. The biopolymer component of EfOM was strongly correlated to EfOM scavenging capacity suggesting that this component may be a main contributor to the EfOM scavenging capabilities due to its molecular size and electron density. The EfOM hydroxyl radical

reaction rate (kOH,EfOM) had a significant positive correlation with specific UV absorbance (SUVA), but to none of the other parameters measured, indicating preferential reactions with conjugated double bonds or aromatic structures. 4. Although all the effluents had relatively similar composition and water quality characteristics, the MBR and AS wastewaters differed significantly in biopolymer concentration. The significantly lower concentration of biopolymers in the MBR wastewater was expected as these high molecular weight components exceed the nominal pore size of the membrane and can also be readily broken down due to relatively longer sludge retention time in the MBR system. Since the study demonstrated that biopolymers may be the main scavenging component of EfOM, the lower biopolymer concentration and scavenging capacity of the MBR wastewater suggests that coupling MBR systems with UV- AOPs would be more feasible than AS systems.

The observations of this study indicate that wastewaters from MBR systems may be more easily treated using UV-AOP systems than AS systems due to the characteristics of the effluent organic matter in these wastewaters, although this is based on a small sample size (10 plants in total). Wastewaters with both low SUVA values and a low concentration of high molecular weight constituents may also be more amenable to UV-AOP treatment.

4.5 References

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Chapter 5 Coagulation Pretreatment of Secondary Wastewater Effluent to Improve UV/H2O2 Efficiency

Abstract

Coagulation using ferric chloride, aluminium sulphate, and polyaluminium chloride were evaluated as pretreatment methods for secondary wastewater effluent prior to UV/H2O2 treatment to reduce the background scavenging capacity of the wastewater. The study assessed whether pretreatment of the wastewater would improve micropollutant degradation, and reduce the energy and cost requirements of UV/H2O2. Coagulation reduced the concentration of effluent organic matter (33-40%), UV254 absorbance (28-39%), and the overall background scavenging capacity of the wastewater (15-20%). Subsequent UV/H2O2 treatment of the pretreated wastewater showed improvements in the micropollutant degradation rates by factors ranging from 1.03-2.59. Reduction in energy requirements varied with the compound and treatment ranging from 0.8-72%. Improving the UV transmittance of the wastewater with coagulation significantly reduced (30-67%) the net present worth (capital and operating costs) of the UV/H2O2 system and these savings outweighed coagulation expenses.

5.1 Introduction

Micropollutants in the environment are of concern as some are known to adversely affect aquatic organisms (Esplugas et al., 2007; Kidd et al., 2007; Snyder et al., 2003). All micropollutants are not readily removed during conventional wastewater treatment (Bolong et al., 2009; Caliman et al., 2009; Kotchen et al., 2009; Mompelat et al., 2009; Radjenovic et al., 2007), and as such, effluents from municipal wastewater treatment plants are a primary point source for the release of these compounds into the environment. While the health effects on humans are inconclusive or not fully understood, micropollutants can potentially impact our drinking water supplies and research is ongoing to evaluate different treatment options.

Membrane filtration, activated carbon adsorption, and advanced oxidation processes are identified as effective technologies for removing micropollutants (Ikehata et al., 2008; Yoon et al., 2007; Yu et al., 2008). However, membrane filtration and activated carbon involve the mass transfer of the compounds to another phase (e.g., the retentate from membrane treatment and adsorption onto the carbon surface) and requires additional treatment for complete removal from the environment. In contrast, advanced oxidation processes (AOPs), which involve the in-situ generation of the highly reactive hydroxyl radical (OH) can destroy the contaminants in the aqueous phase (Dantas et al., 2012; Ikehata et al., 2006; Rosenfeldt et al., 2004; Shu et al., 2013). AOPs in wastewater treatment systems, however, can be relatively expensive in comparison to other common forms of treatment due to its energy requirements. These requirements are primarily a function of the hydroxyl scavenging capacity exerted by scavengers in the wastewater.

Typical hydroxyl radical scavengers in wastewaters are bicarbonate, carbonate, nitrite and effluent organic matter (EfOM) (Keen et al., 2014; Rosario-Ortiz et al., 2010; Rosenfeldt et al., 2007). Of these, EfOM is considered the primary scavenger where as demonstrated in Chapter 4, it accounted for more than 70% of the overall scavenging capacity of 10 wastewaters. Also, Souza et al. (2014) reported that EfOM accounted for 85% of the overall scavenging, while Keen et al. (2014) showed that it constituted 45-60% of the overall background scavenging capacity in a study of 28 wastewater samples. Due to the scavenging potential and oxidant

95 demand exerted by these constituents, larger treatment doses would be required for effective oxidation of target compounds (Autin et al., 2013; Dantas et al., 2012; Souza et al., 2014). Therefore, reducing EfOM concentration, as the major scavenger, should improve the availability of OH radicals and AOP efficiency.

Coagulation is a process by which particulates and organic matter are removed by chemical addition. This technique can be optimised to achieve targeted removal of organic matter. As such, optimising coagulation to reduce effluent organic matter (EfOM) concentration in wastewater is expected to make AOP treatment more effective due to a lower oxidant demand. With this improved effectiveness, operating costs and the energy requirements of the AOP process should also be reduced. However, the effect of coagulation on subsequent AOP efficiency, and specifically UV/H2O2 treatment, has not yet been reported.

The key objective of this study was to assess the effect of coagulation on the overall background scavenging capacity, EfOM scavenging capacity, and the subsequent efficiency of UV photolysis and UV/H2O2 treatment of micropollutants in wastewater. The coagulants used were ferric chloride, aluminium sulphate, and polyaluminium chloride based on their prevalence in the water and wastewater industry. One secondary wastewater source was used for all the experiments to maintain consistency in the effluent quality when comparing the effect of the coagulation. The effects of coagulation pretreatment on UV/H2O2 efficiency were evaluated based on the degradation rates of the micropollutants and the electrical energy requirements of

UV/H2O2. A cost analysis was also conducted to determine the extent to which pretreatment of the wastewater may affect capital and operating costs when using the net present values (NPV) for a 20 year lifespan of a new wastewater plant using UV/H2O2.

5.2 Materials and Methods

5.2.1 Wastewater Samples

Municipal wastewater was collected from an Ontario extended aeration wastewater treatment plant immediately after the secondary clarifier but prior to filtration or disinfection. Samples were collected in pre-cleaned 20 L polypropylene containers and stored at 4oC for a maximum

96 period of six weeks until the experiments were completed. The samples were tested at the start of each experiment for UV transmittance (UVT) at 254 nm to verify their stability during storage, with the relative standard deviation found to be less than 2%. Sample characteristics are shown in Table 5-1.

Table 5-1: Water quality characteristics of the secondary wastewater Parameter Secondary Effluent pH 7.08 Temperature (oC) 20 -1 UVA254 (cm ) 0.130 % UVT (cm-1) 64.1 Conductivity (µS/cm) 1086

SUVA254 (L/mg-m) 2.21 TOC (mg-C/L) 9.2 DOC (mg-C/L) 8.1 TIC (mg-C/L) 37.7 2- Carbonate (mg CO3 /L) 0.07 - Bicarbonate (mg HCO3 /L) 113.2

Total Alkalinity (mg CaCO3/L) 186 Nitrite (mg/L-N) 0.21 Nitrate (mg/L-N) 20.5

5.2.2 Experimental Approach

The secondary wastewater was treated using coagulation with ferric chloride (FeCl3), aluminium sulphate (alum), and Hyper+Ion (HI)705 polyaluminium chloride (PACl). PACl is a high basicity (>80%) pre-hydrolyzed aluminium coagulant with an aluminium content of 11.3%–

12.1% by weight. Alum and FeCl3 were obtained from Sigma Aldrich with a purity of greater than 98%. Following each pretreatment, the pretreated sample was spiked with the target micropollutants (caffeine, carbamazepine, clofibric acid, sulphamethoxazole, naproxen, diclofenac, and 17-estradiol) and subjected to UV/H2O2 treatment. A wastewater sample which underwent no pretreatment was used as a control.

Micropollutant compounds representative of different classes of micropollutants typically found in wastewaters were selected based on their reported occurrence in wastewater effluents, availability of photochemical information, and the ease with which the compounds could be analysed at the University of Toronto Civil Engineering Environmental Laboratory. An elevated initial spike concentration (50 µg/L), which is approximately 100 times the typical concentration for these compounds in wastewater effluents, was used to ensure the degradation of the compounds during treatment could be readily monitored.

Initial bench-scale experiments were performed to identify the optimum doses for pretreatment. Coagulation was performed with a six-paddle jar testing apparatus (Phipps & Bird, VA, USA) using a rapid mix of 200 rpm for 1 minute to ensure proper mixing of the coagulant, a slow mix at 30 rpm for 30 minutes to promote flocculation, and settling for a period of 30 minutes with the paddles removed. Alum and ferric chloride coagulation were optimised at pH 6.0 using sulphuric acid (H2SO4) or sodium hydroxide (NaOH) for pH adjustment. The optimum coagulant doses were identified based on a point-of-diminishing return (PODR) analysis for dissolved organic carbon (DOC) as a surrogate measurement for EfOM. The point of diminishing returns was determined as the dose for which a 10 mg/L incremental increase in the applied dose of coagulant resulted in a change in DOC removal of less than 0.3 mg/L (USEPA 1999). The observed PODRs were: aluminium sulphate (alum) (12 mg Al/L), polyaluminum chloride (PACl) (16 mg Al/L), and ferric chloride (FeCl3) (21 mg Fe/L or 60 mg/L FeCl3). The dose-response curves for the coagulants are shown in Figure 1 in Appendix H.

5.2.3 Advanced Oxidation Experiments

UV/H2O2 advanced oxidation was conducted in a 42 L cylindrical stainless steel Calgon Carbon Rayox Advanced Oxidation Batch Pilot Reactor (Figure 5-1). The AOP reactor was equipped with a 1kW medium-pressure (MP) Hg-lamp in the centre of a quartz sleeve, a mixer, and a steel shutter that was used to control the length of time the water sample was exposed to the UV light.

Pre-treated wastewater samples (40L) were spiked with 50 µg/L of each compound (caffeine, naproxen, diclofenac, sulphamethoxazole, carbamazepine, clofibric acid, and 17-estradiol), as well as H2O2 at concentrations of 0, 10, and 20 mg/L, before UV/H2O2 treatment. The power to the MP UV lamp was 1.026 kW, as measured using a 600A AC/DC True RMS Clamp meter inductive ammeter, Model MA640 (ExTech Instruments Corporation, NH,USA). This value was used to calculate the electrical energy per order (EEO) requirements for the micropollutant degradation, as well as the UV fluence (dose), using UVCalc software (Bolton Photosciences Inc., Edmonton, Canada). This software also required the wastewater absorbance, the relative lamp emittance, and the quartz sleeve transmittance in the range 200 – 300 nm as input parameters. The water samples were exposed to UV light for a maximum exposure time of 5 minutes, which was calculated to deliver a UV fluence of 3,200 mJ/cm2. Duplicate experiments for micropollutant degradation were performed and 500 ml of sample was collected at 1 minute intervals for analysis. Sodium thiosulphate was added to quench the residual hydrogen peroxide before analysing the samples for pharmaceutical concentrations and UV254 absorbance.

A 100L stainless steel tank was used for coagulating the samples using the optimum doses from the jar tests, and using the same coagulation/flocculation protocol. Samples were analysed for pH, DOC, UV254 and hydroxyl radical scavenging capacity.

Figure 5-1: Calgon Carbon Advanced Oxidation Batch Reactor

5.2.4 Equipment and Methods

A Calgon Carbon® quasi-collimated beam apparatus (Model PSI-I-120, Calgon Carbon Corporation, USA) equipped with a 1 kW medium-pressure (MP) Hg-lamp was used for  measuring OH exposure per UV fluence (ROH,UV) and the overall scavenging capacity of the wastewater according to the protocol outlined by Rosenfeldt et al. (2007). The ROH,UV is the steady-state concentration of OH radicals present in the wastewater per unit dose of UV for an initial concentration of hydrogen peroxide. Hence, wastewater with a higher background OH scavenging will have a lower steady-state concentration of OH per unit UV (i.e., a lower

ROH,UV). Therefore, measuring the ROH,UV in the wastewater is used to determine the overall background scavenging capacity. To determine the ROH,UV, the degradation of methylene blue (MB), a hydroxyl radical probe compound, was monitored in the collimated beam tests by measuring the change in absorbance readings at 664 nm and applying the Beer-Lambert law. Hydrogen peroxide concentration was measured using the triiodide method (Klassen et al., 1994). The incident irradiance of the medium-pressure (MP) UV lamp in the collimated beam apparatus was determined using ferrioxalate actinometry (Sharpless and Linden 2003). Average fluence rate was determined using the Bolton® Excel Spreadsheet for fluence calculations using a medium pressure lamp with suspension depth of less than 2 cm (Bolton 2004). Samples were irradiated in the collimated beam apparatus for the time required to achieve a UV fluence ranging from 0–800 mJ/cm2.

Using a MP lamp for UV irradiation of wastewater can lead to the formation of nitrite due to nitrate photolysis. Nitrite is a significant hydroxyl scavenger with a hydroxyl radical reaction rate constant of 1.0 x 1010 M-1s-1 (Buxton et al., 1988). Measurement of nitrite formation as a function of UV fluence during irradiation with the medium pressure lamp found that the initial molar concentration of nitrite approximately doubled, leading to a 14.9% increase in the overall   OH scavenging, and in turn a 10.4% decrease in the OH exposure per UV fluence at a H2O2 dose of 10 mg/L. In the presence of 20 mg/L H2O2, overall scavenging capacity increased by 13.9%, along with a 9.3% decrease in the OH exposure per UV fluence. Since the overall scavenging capacity of a wastewater is the sum of the scavenging potential of the known scavengers, the scavenging potential of the nitrite formed at the maximum exposure time was

100 subtracted from the measured overall scavenging capacity to determine the “true” scavenging capacity and OH exposure of the wastewater samples.

The wastewater was analysed for pH using a calibrated pH meter (Thermo Scientific Orion Star A111). UV absorbance readings were determined using a 1 cm quartz cuvette in an Agilent 8453 UV-VIS spectrophotometer. Dissolved organic carbon (DOC), total organic carbon (TOC) and total inorganic carbon (TIC) were measured using a TOC analyser (O.I. Analytical Aurora Model 1030 with auto-sampler Model 1088) in accordance with Standard Method 5310C (APHA 1998). The concentration of the anions (nitrate, nitrite) were determined according to USEPA Method 300.1 (USEPA 1997) using a Dionex Ion Chromatography system equipped with a RFIC IonPac (Dionex AS9-HC, 4 x 250 mm) analytical column and a Dionex AG9-HC 4 x 50mm guard column, a CD20 conductivity detector, GP40 gradient pump, and an AS40 automated sampler. The alkalinity of the effluent was determined using Standard Method 2320B (APHA 1998). The constituents of EfOM were characterised into five organic fractions of biopolymers, humics, building blocks of humics, low molecular weight (LMW) neutrals and low molecular weight acids using Liquid Chromatography Organic Carbon (LC-OCD) analysis following the protocol developed by Huber et al. (2011). The micropollutants (caffeine, naproxen, diclofenac, sulphamethoxazole, carbamazepine, clofibric acid, and 17-estradiol) were analysed using solid phase extraction with Oasis hydrophilic-lipophilic balance (HLB) extraction cartridges (Waters Oasis HLB 6 cc, 150 mg, 30 µm) followed by analysis using liquid chromatography-tandem mass spectrometry (LC-MS/MS) (MOE, 2008). Analyses were performed using an Agilent LC (Model 212-LC) system equipped with an Agilent 500-MS IT mass spectrometer, a Pursuit XRs-C18 guard column (Metaguard 2.0 mm ID x 3 mm), and a Pursuit XRs Ultra 2.8-C18 analytical column (2.0 mm ID x 100 mm, 2.8 µm particle size) (Agilent Technologies, Mississauga, Canada).

5.2.5 Chemicals and Reagents

The pharmaceuticals (caffeine, naproxen, diclofenac, sulphamethoxazole (SMZ), carbamazepine (CBZ), 17-estradiol (E2) and clofibric acid) were obtained from Sigma Aldrich (Toronto, Canada) in powdered forms with a purity of greater than 98%. The compounds were selected to

101 be representative of different classes of micropollutants and on the basis of their common occurrence in wastewater effluents as reported in the literature (Metcalfe et al., 2003; Miao et al., 2004). Other chemicals such as 30 wt.% hydrogen peroxide, methylene blue, acetonitrile (LC-MS grade), methanol (LC-MS grade), sodium hydroxide, sulphuric acid, sodium thiosulphate, and sodium acetate were also obtained from Sigma Aldrich. Milli-Q water was generated using a Milli-QUV Plus Ultrapure Water System and was used for preparing the stock solutions.

5.3 Results and Discussion

5.3.1 Effects on Wastewater Quality

Coagulation pretreatment of the wastewater significantly reduced the DOC concentration (33-

40%), UV254 (28-39%), overall background scavenging capacity (12-16%) and EfOM scavenging capacity (15-21%) based on a paired sample t-test at a 90% confidence interval (C.I.) (Table 5-2 and Table 5-3). It was also found that the reduction in DOC concentration was strongly correlated to a reduction in the overall scavenging capacity (R2 = 0.937) and increased OH exposure (R2 = 0.879). Given that EfOM is usually the primary OH scavenger in wastewater where it can account for more than 70% of the overall scavenging capacity as discussed in Chapter 4 and by Souza et al. (2014), this observation is as expected.

Although a significant reduction in OH scavenging capacity was achieved using coagulation, this reduction was approximately only half of that observed for DOC and UV254 removal. In other words, the coagulants were approximately twice more effective at removing bulk DOC  and UV254 than at removing organic OH radical scavengers. The research in Chapter 4 suggested that organic fractions with aromatic structures or a high concentration of conjugated double bonds were effective radical scavengers. Organics with these same structures are also well removed by coagulation as reported by Haberkamp et al. (2007) and Matilainen et al. (2010). The current research also demonstrated that coagulation resulted in the removal of such compounds as discussed in Chapter 4, and as demonstrated by LC-OCD results (Figure 5-2) which showed that coagulation removed the high molecular weight organic components (biopolymers and humics) and the building blocks of humics with reductions ranging from 17%

102 to 58%, but did not affect the low molecular weight (LMW) neutrals and LMW acids. As such, these results suggest that there are fractions of organic matter which do not react well with the OH radicals, yet are still effectively removed by coagulation.

Raw effluent PACl FeCl3 Alum 3.0

2.5

2.0

1.5

1.0

Concentration Concentration (mg/L) 0.5

0.0 biopolymers humics building LMW neutrals LMW acids blocks Figure 5-2: Change in organic composition with pretreatment of the wastewater (Error bars are the standard deviation of duplicate samples)

The coagulants all led to reductions in overall scavenging capacity (Table 5-2 and Table 5-3), but there was no statistically significant difference between the three coagulants based on a one- way analysis of variance (ANOVA) at a 90% C.I. As such, a specific coagulant could not be identified as performing significantly better than the others. Previous studies which have focused on using coagulation as a pretreatment step for specifically improving UV/H2O2 treatment of wastewater could not be found in the literature to date. Hence, a comparison to the observations in this current study could not be performed.

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Table 5-2: Changes in wastewater quality with coagulation treatment

Overall a,b a,b Overall Scavenging EfOM ROH,UV @ a Optimum DOC Scavenging Capacity Scavenging 15 mg/L H2O2 UVT Treatment UV pH Dose (mg/L) Capacity (excluding Capacity (× (x 10-14) 254 (%/cm)

5 -1 5 -1 2 -1 (× 10 ) s Nitrite) 10 ) s M s cm mJ 5 -1 (× 10 ) s No pretreatment n/a 7.77 17.15 ± 0.21 14.33 14.17 3.33 ± 0.00 0.123 75 6.90

FeCl3 60 mg/L 4.75 14.95 ± 2.33 12.12 11.96 3.75 ± 0.59 0.079 83 6.32 Alum 12 mg Al/L 4.61 14.40 ± 4.38 11.51 11.35 4.17 ± 1.18 0.075 84 6.19

PACl 16 mg Al/L 5.17 15.05 ± 2.33 12.19 12.03 3.75 ± 0.59 0.088 82 6.95 aAverages ± standard deviation of duplicate experiments bValues were determined using a 1kW medium pressure lamp.

Table 5-3: Percentage change in wastewater quality with coagulation treatment relative to no pretreatment

% Reduction % Reduction in Overall % Reduction % Reduction Optimum % DOC of Overall Scavenging in EfOM % Increase in Treatment in UV254 Dose Removal Scavenging Capacity Scavenging ROH,UV -1 (cm ) Capacity (excluding Capacity Nitrite) FeCl3 60 mg/L 38.9 12.8 15.4 15.6 12.5 35.6 Alum 12 mg Al/L 40.7 16.0 19.7 19.9 25.0 39.0 PACl 16 mg Al/L 33.5 12.2 15.0 15.1 12.5 28.4

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5.3.2 Effects on Micropollutant Degradation

In general, coagulation of the wastewater prior to UV photolysis increased the photolytic degradation rates of the seven compounds (i.e., in the absence of H2O2) by factors ranging from 1.14 to 2.59 as shown in Table 5-4, with some exceptions. For each compound, the degradation rates were determined using the gradient of the logarithmic plots of pharmaceutical concentration versus UV exposure time (Appendix I). Caffeine, carbamazepine, clofibric acid and diclofenac showed an increase in the photolytic degradation rates irrespective of the coagulant applied, however only some rate increases were considered to be statistically significant based on a one-way ANOVA followed by Tukey’s Honestly Significant Difference (HSD) post hoc test at a 90% confidence interval (C.I.) (Table 5-4).

Table 5-4: Photolytic degradation rate constants at different treatment conditions aPhotolytic Degradation Rate Constants (min-1) with Different Treatments Micropollutant No pretreatment PACl FeCl3 Alum Caffeine 0.027 ± 0.016 0.030 ± 0.005 0.039 ± 0.017 0.031 ± 0.002 Carbamazepine 0.035 ± 0.009 0.059 ± 0.002b 0.059 ± 0.002b 0.077 ± 0.002b Naproxen 0.232 ± 0.005 0.256 ± 0.008 0.221 ± 0.011 0.314 ± 0.035b 17β-estradiol 0.215 ± 0.029 0.295 ± 0.0002b 0.118 ± 0.014 0.370 ± 0.013b Sulphamethoxazole 0.606 ± 0.053 0.368 ± 0.008 0.379 ± 0.003 0.631 ± 0.014 Clofibric acid 0.313 ± 0.002 0.363 ± 0.002 0.609 ± 0.023b 0.812 ± 0.132b Diclofenac 0.85 ± 0.106 1.85 ± 0.217b 1.23 ± 0.096 2.27 ± 0.208b aAverage ± standard deviation of duplicate experiments bIndicates degradation rates that were significantly higher compared to the no pretreatment sample based on a 1- way ANOVA followed by Tukey’s HSD post hoc test at 90% C.I.

Fluctuations in the degradation rates of naproxen, 17β-estradiol, and sulphamethoxazole were observed after coagulation, with the degradation rates increasing normally with pretreatment, as expected, but occasionally decreasing when using ferric chloride. The reason for this counterintuitive observation is unknown, but there appeared to be greater variability in the data from the ferric chloride experiments (Appendix H). Given the variability in the degradation rates, a specific coagulant could not be identified as consistently outperforming the others. Nevertheless, the weight of evidence from this test demonstrated that coagulation pretreatment of wastewater prior to UV photolysis will generally improve the removal rates of the micropollutant compounds.

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Although an increase in oxidation rates was observed for most of the compounds, it should be noted that the increase in the degradation rates for caffeine and carbamazepine (both of which are not readily photolyzed by UV alone), in the absence of hydrogen peroxide, is most likely due to oxidation by hydroxyl radicals generated from nitrate photolysis and UV irradiation of the bulk EfOM. Hydroxyl radical generation via these two mechanisms have been reported by Dong et al. (2012) and Lee et al. (2013) with formation rates ranging from 0.96 × 10-10 to 4.8 × 10-10 M s-1 in wastewater.

Reducing the organic content of the wastewater using coagulation improved the wastewater quality in terms of scavenging capacity and UV transmittance (UVT). Improvements in the UVT of the wastewater from 75%/cm to an average of 83%/cm (Table 5-2) means that the specific rate of light absorption would increase, resulting in higher UV fluences for a given exposure time. In this study, the UV fluence values increased by 20-49% with coagulation of the wastewater for the same exposure time. This higher UV fluence will lead to an increase in the degradation rates of the target compounds, nitrate photolysis, and the production of OH from the UV irradiation of EfOM.

Similar to UV photolysis, coagulation prior to UV/H2O2 treatment generally increased the degradation rates of the compounds irrespective of the hydrogen peroxide concentration when compared to UV/H2O2 with no pretreatment. There were exceptions primarily associated with

FeCl3 where the rates counterintuitively decreased. Apart from these isolated cases, the degradation rates of all seven compounds increased by factors ranging from 1.03 to 2.70 at 10 mg/L H2O2 (Table 5-5). Similarly, when treatment with 20 mg/L H2O2 was applied, the degradation rates normally increased by factors of 1.05 to 1.61 relative to no pretreatment (Table 5-6). These increases were expected as the reduced scavenging capacity of the wastewater, specifically EfOM scavenging capacity, increased the OH exposure (12-25%) indicating that a higher concentration of OH radicals would be available in the wastewater to react with the target compounds. These findings are analogous to the study reported by Wert et al. (2011) where reducing the DOC concentration using ferric chloride coagulation led to a subsequent increases in the ozonation degradation rate of 13 trace organic contaminants

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(atenolol, carbamazepine, DEET, diclofenac, dilantin, gemfibrozil, ibuprofen, meprobamate, naproxen, primidone, sulfamethoxazole, triclosan, and trimethoprim).

Table 5-5: Degradation rate constants with coagulation and UV + 10 mg/L H2O2 aDegradation Rate Constants (min-1) with UV + 10 mg/L H O Micropollutant 2 2 No pretreatment PACl FeCl3 Alum Caffeine 0.046 ± 0.010 0.071 ± 0.006b 0.061 ± 0.009 0.075 ± 0.001b Carbamazepine 0.103 ± 0.004 0.110 ± 0.003 0.093 ± 0.006 0.129 ± 0.006b Naproxen 0.264 ± 0.012 0.287 ± 0.0004 0.366 ± 0.089 0.368 ± 0.023 17β-estradiol 0.375 ± 0.062 0.418 ± 0.049 n.d. 0.881 ± 0.790 Sulphamethoxazole 0.481 ± 0.011 0.496 ± 0.012 0.406 ± 0.028 0.847 ± 0.064b Clofibric acid 0.320 ± 0.022 0.365 ± 0.005 0.757 ± 0.073b 0.707 ± 0.061b Diclofenac 0.807 ± 0.066 1.71 ± 0.174 1.21 ± 0.055 2.18 ± 0.977 aAverage ± standard deviation of duplicate samples; bIndicates degradation rates that were significantly higher compared to the no pretreatment sample based on a 1- way ANOVA followed by Tukey’s Honestly Significant Difference (HSD) post hoc test at 90% C.I. n.d. – no data due to lost sample

Table 5-6: Degradation rate constants with coagulation and UV + 20 mg/L H2O2

aDegradation Rate Constants (min-1) with UV + 20 mg/L H O Micropollutant 2 2 No pretreatment PACl FeCl3 Alum Caffeine 0.104 ± 0.021 0.109 ± 0.004 0.128 ± 0.037 0.126 ± 0.001

Carbamazepine 0.129 ± 0.005 0.181 ± 0.003 0.168 ± 0.044 0.203 ± 0.0002b

Naproxen 0.294 ± 0.014 0.313 ± 0.020 0.250 ± 0.014 0.490 ± 0.075b 17β-estradiol 0.443 ± 0.026 0.504 ± 0.130 n.d. 0.470 ± 0.089

Sulphamethoxazole 0.481 ± 0.001 0.544 ± 0.019 0.533 ± 0.028 0.936 ± 0.061b Clofibric acid 0.245 ± 0.005 0.394 ± 0.008 n.d. 1.28 ± 0.351 Diclofenac 1.74 ± 0.918 1.58 ± 0.034 1.30 ± 0.132 2.45 ± 0.484 aAverage ± standard deviation of duplicate samples bIndicates degradation rates that were significantly higher compared to the no pretreatment sample based on a 1- way ANOVA followed by Tukey’s Honestly Significant Difference (HSD) post hoc test at 90% C.I. n.d. – no data due to lost sample

As mentioned, however, in some isolated cases pretreatment was followed by a decrease in the rate of degradation for certain compounds in the wastewater coagulated using ferric chloride (Table 5-5 and Table 5-6). At pH of less than 8.0, the soluble species present following 3+ + 3+ coagulation with ferric chloride are usually Fe , Fe(OH)2 and FeOH (Gabelich et al., 2002). 3+ Of these species, Fe can catalyse the decomposition of H2O2 reducing the concentration

107 available for OH generation (De Laat et al., 1999; Dunford 2002; Lin et al., 1998) and could be a possible reason for the lower oxidation rates, as the pH of the coagulated samples ranged from pH 6.2-7.0 (Table 5-2). Another possibility may be experimental errors that could have occurred during sample preparation for analysis of the compounds. It is unlikely that residual coagulant would have reduced the UVT of the wastewater as the UV absorbance of the wastewater treated by ferric chloride was less than the untreated wastewater at all UV and visible wavelengths (Appendix J), indicating an improvement in the wastewater quality. Nitrite scavenging of the OH radicals may have been a possibility due to the medium–pressure UV lamps; however, as discussed in Section 5.3.1, the concentrations formed were insufficient to substantially scavenge the OH radicals required for oxidising the target compounds. Overall, however, the lower degradation rates were considered to be exceptions. An improvement in the wastewater quality by reducing the background scavenging capacity, and improving the UVT and OH exposure using coagulation pretreatment is very likely to improve the degradation rates as observed for most of the compounds.

Comparing the low (10 mg/L) and high (20 mg/L) H2O2 concentrations in UV/H2O2 treatment of the wastewater, it was observed that the degradation rates of the seven compounds were greater at 20 mg/L H2O2 concentration compared to 10 mg/L H2O2, regardless of whether the water was pretreated by coagulation. This rate increase was most apparent with the caffeine and carbamazepine where an increase in hydrogen peroxide concentration significantly increased the degradation rates of both caffeine (Figure 5-3) and carbamazepine (Figure 5-4) by factors of 1.5 to 4.1, irrespective of the coagulation option applied, based on a two-way ANOVA at a 90% C.I. followed by Tukey’s HSD post hoc test.

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No pretreatment PACl FeCl3 Alum 0.18

0.16

0.14

0.12

) ) 0.10

1 -

0.08 (min 0.06

0.04

0.02

Caffeine Degradation Rate Constants Rate Degradation Caffeine 0.00 0 mg/L H2O2 10 mg/L H2O2 20 mg/L H2O2

Figure 5-3: Degradation rate for caffeine with coagulation and UV/H2O2 (Error bars are the standard deviation of duplicate samples)

No pretreatment PACl FeCl3 Alum 0.25

0.20

1) -

0.15

0.10 Constants (min Constants

0.05 Carbamazepine Degradation Rate Degradation Carbamazepine

0.00 0 mg/L H2O2 10 mg/L H2O2 20 mg/L H2O2

Figure 5-4: Degradation rate for carbamazepine with coagulation and UV/H2O2 (Error bars are the standard deviation of duplicate samples)

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This demonstrates that the hydrogen peroxide concentration in UV/H2O2 treatment can also influence degradation efficiency. Additional hydrogen peroxide in a UV/H2O2 process means additional photons of light will be absorbed by the H2O2 which increases the steady-state  concentration of OH radicals (OHss) available in the water matrix, resulting in increased degradation. However, this improved performance does not occur ad infinitum, as Crittenden et al. (1999) noted that at high concentrations, the hydrogen peroxide itself can become a major OH scavenger. Similar findings were reported by Li et al. (2008) who suggested that hydrogen peroxide only becomes a significant OH scavenger at concentrations exceeding 100 mg/L.

This current study has demonstrated that coagulation of the wastewater to reduce the organic  carbon (DOC) content generally improves the steady-state concentration of OH radicals (OHss) as shown by the increase in the ROH,UV values of Table 5-2 and Table 5-3. Since an increase in

OHss can result from both H2O2 addition and DOC removal, kinetic modelling of the OHss concentration in the wastewater using Matlab™ was performed. The purpose of the model was to determine if there was a H2O2 concentration at which reducing the DOC content would have little impact on the OHss (i.e., the scavenging of DOC becomes very small and unimportant   relative the the formation of OH and the scavenging of OH by H2O2 itself). The model was used to compare the OHss concentration in a non-pretreated wastewater with similar characteristics to the wastewater used in this study, and the OHss when 50% (an arbitrary level) of the initial DOC concentration (4.26 mg/L) was removed by coagulation. Hydrogen peroxide concentrations of 5 to 2000 mg/L were evaluated for treatment in a hypothetical collimated beam unit with an assumed maximum UV fluence rate of 52.5 mW/cm2.

The model found that the benefits of pretreating the wastewater to remove DOC and increase the OHss decreased with H2O2 addition such that at concentrations exceeding 600 mg/L H2O2, the benefit of achieving higher OHss with pretreatment to remove DOC was less than 10%. A dose of 600 mg/L is unrealistically high, and as such, this model suggests that removal of DOC prior to UV/H2O2 will always offer a meaningful improvement to AOP performance, regardless of the H2O2 concentration. It was also determined that increasing the H2O2 concentration did not result in a directly proportional increase in OHss, which instead reached a maximum value and subsequently decreased. The maximum OHss concentration was obtained at 400 mg/L H2O2 and

110

500 mg/L H2O2 for the non-pretreated wastewater and the coagulated wastewater respectively.

The subsequent decrease in OHss concentration at higher H2O2 doses is due to the scavenging of  the OH radicals by the H2O2 and blocking the absorbance of the UV light entering the wastewater.

5.3.3 Effects on Electrical Energy Requirements

The electrical energy required for UV photolysis and UV/H2O2 advanced oxidation treatment is assessed using the standard figure-of-merit of electrical energy per order (EEO) (Bolton et al., 2001). The EEO is defined as the electrical energy (kWh) required to achieve 90% degradation of a compound in 1 m3 of water, and was determined for each micropollutant using Equation 5.1.

( ) (5.1) ( )

where P (kW) is the rated power of the UV lamp, t (minutes) is the exposure time for treating the volume of water, V (L) is the volume of water treated, and Ci and Cf are the initial and final concentrations of the target compounds.

For UV photolysis, coagulation of the wastewater reduced the energy required to achieve a 1- log (90%) degradation (i.e., the EEO) of clofibric, diclofenac, caffeine and carbamazepine compounds with reductions ranging from 8-64% (Table 5-7). Exceptions were observed for naproxen, 17β-estradiol, and sulphamethoxazole where similar to the compound degradation rates, the energy requirements varied depending on the coagulant applied. In the presence of

UV/H2O2 treatment with 10 mg/L H2O2 (Table 5-8) and 20 mg/L H2O2 (Table 5-9), coagulation pretreatment of the wastewater also reduced the energy required for compound degradation.

Reductions ranged from 5-59% at 10 mg/L H2O2 and from 0.8-72% at 20 mg/L H2O2. The decrease in energy consumption for UV photolysis and UV/H2O2 treatment is expected. The increase in the degradation rates due to the higher UV fluences means the targeted 90% removal of the micropollutant is achieved in a shorter time, thereby requiring less energy.

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Table 5-7: EEO values for 1-log micropollutant degradation using UV photolysis alone aElectrical energy per order (EEO) with UV Photolysis (kWh/m3/order) Micropollutant No pretreatment PACl FeCl3 Alum Caffeine 17.99 ± 9.97 14.16 ± 2.45 11.66 ± 5.45 13.97 ± 0.64 Carbamazepine 12.69 ± 3.12 7.25 ± 0.33b 7.36 ± 0.25b 5.36 ± 0.01b Naproxen 1.81 ± 0.04 1.66 ± 0.14 2.07 ± 0.16 1.32 ± 0.14b 17β-estradiol 2.11 ± 0.27 1.58 ± 0.02 3.58 ± 0.40b 0.96 ± 0.02b Sulphamethoxazole 0.73 ± 0.02 0.94 ± 0.03b 1.02 ± 0.07b 0.58 ± 0.02 Clofibric acid 1.10 ± 0.03 0.90 ± 0.04b 0.79 ± 0.02b 0.57 ± 0.07b Diclofenac 0.54 ± 0.02 0.23 ± 0.03 0.35 ± 0.03 0.19 ± 0.02 aAverages ± standard deviation of duplicate samples bEEO values which are significantly different with pretreatment compared to the no pretreatment sample based on a one-way ANOVA followed by a Tukey’s HSD post hoc test at 90% C.I.

For UV photolysis and UV/H2O2 treatment, naproxen, 17β-estradiol, sulphamethoxazole, clofibric acid, and diclofenac required less than 3.6 kWh/m3/order for 90% removal. In contrast, caffeine and carbamazepine EEOs ranged from 2-18 kWh/m3/order depending on the treatment applied. The effect of treatment was most apparent with these two compounds, where lower EEO values were obtained with an increase in the hydrogen peroxide concentration such that the lowest EEO values for these compounds were obtained in the presence of 20 mg/L H2O2.

Table 5-8: EEO values for 1-log micropollutant degradation using UV + 10 mg/L H2O2 aElectrical energy per order (EEO) values with 3 Micropollutant UV + 10 mg/L H2O2 (kWh/m /order) No pretreatment PACl FeCl3 Alum Caffeine 9.70 ± 2.2 5.75 ± 0.59b 7.18 ± 1.03 5.67 ± 0.02b Carbamazepine 4.19 ± 0.15 3.77 ± 0.00 4.38 ± 0.22 3.37 ± 0.09b Naproxen 1.68 ± 0.12 1.28 ± 0.03 1.30 ± 0.26 1.18 ± 0.06b 17β-estradiol 1.27 ± 0.34 1.20 ± 0.07 nd 0.89 ± 0.77 Sulphamethoxazole 0.88 ± 0.01 0.85 ± 0.02 1.02 ± 0.10 0.49 ± 0.01 Clofibric acid 1.16 ± 0.00 0.85 ± 0.03 0.58 ± 0.05b 0.49 ± 0.15b Diclofenac 0.53 ± 0.04 0.25 ± 0.03b 0.35 ± 0.02 0.22 ± 0.10b aAverages ± standard deviation of duplicate samples bEEO values which are significantly different with pretreatment compared to the no pretreatment sample based on a one-way ANOVA followed by a Tukey’s HSD post hoc test at 90% C.I. nd – no data due to lost samples

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Table 5-9: EEO values for 1-log micropollutant degradation with UV + 20 mg/L H2O2 aElectrical energy per order (EEO) values with 3 Micropollutant UV + 20 mg/L H2O2 (kWh/m /order) No pretreatment PACl FeCl3 Alum Caffeine 3.80 ± 0.54 3.91 ± 0.08 5.32 ± 1.58 3.42 ± 0.13b Carbamazepine 3.11 ± 0.04 2.39 ± 0.02 2.66 ± 0.63 2.15 ± 0.05b Naproxen 1.56 ± 0.06 1.04 ± 0.01b 1.71 ± 0.09 0.90 ± 0.08b 17β-estradiol 1.00 ± 0.01 0.99 ± 0.24 nd 0.97 ± 0.22 Sulphamethoxazole 0.84 ± 0.00 0.80 ± 0.04 0.74 ±0.03b 0.46 ± 0.02b Clofibric acid 1.78 ± 0.08 0.90 ± 0.10 nd 0.49 ± 0.00 Diclofenac 0.29 ± 0.15 0.27 ± 0.01 0.33 ± 0.03 0.18 ± 0.03 aAverages ± standard deviation of duplicate samples bEEO values which are significantly different with pretreatment compared to the no pretreatment sample based on a one-way ANOVA followed by a Tukey’s HSD post hoc test at 90% C.I nd. – no data due to lost samples

5.3.4 Effects on Treatment Costs

A cost analysis to assess how coagulation may affect operating costs for a UV/H2O2 advanced oxidation system and overall plant costs was based on a scenario of a newly constructed plant whose operating conditions and wastewater characteristics were identical to the wastewater plant used in this current study. Hence, the cost model was based on a new extended aeration plant operating with a peak flow of 64,500 m3/day, an average flow of 18,000 m3/day and a secondary wastewater effluent with a UV transmittance of 64%/cm that would be treated using a

UV/H2O2 system equipped with medium-pressure (MP) UV lamps. Model analysis was also based on the assumption that doped MP lamp sleeves would be used to prevent nitrite formation. A 1-log (90%) removal of carbamazepine was used as the performance standard for the

UV/H2O2 system. At operating conditions, an improvement in the UVT of the wastewater from 64%/cm to 74%/cm (based on experimental data from this study) and 84%/cm (projected value) using coagulation pretreatment of the wastewater were considered.

The Net Present Value (NPV) analysis was based on a plant life of 20 years and a 6% interest rate. The analysis was done for the new plant encompassing only primary and secondary treatment and for a separate UV/H2O2 system that would be added for tertiary treatment. Costs specifically for the UV/H2O2 system were based on the assumption that coagulation would occur upstream of the AOP unit in the secondary clarifier such that the infrastructure for

113 coagulation would already exist. The coagulants were also assumed to have the same cost. The CapDet Works Software from Hydromantis Environmental Software Solutions Inc. (Hamilton, Ontario, Canada) and collaboration with Calgon Carbon Corporation (Pittsburgh, PA, USA) were used for the cost estimations. Calgon Carbon Corporation internal models were used to predict carbamazepine degradation and EEO values at 84%/cm UVT. Higher hydrogen peroxide doses of 50 mg/L and 100 mg/L were also included in the cost analysis. Results are shown in Figure 5-5.

64% UVT 74% UVT 84% UVT $50,000,000

$45,000,000 Enhanced coagulation costs: $40,000,000 74% UVT = $2,000,000 84% UVT = $4,000,000 (estimated)

system alone alone system $35,000,000

2 2 O 2 $30,000,000

$25,000,000

(CAD$) (CAD$) $20,000,000

$15,000,000

$10,000,000

$5,000,000 year Present Worth of UV/H Present Worth year - $- 20 AOP-10mg/L AOP-20mg/L AOP-50mg/L AOP-100mg/L H2O2 H2O2 H2O2 H2O2

Figure 5-5: Cost of acheving 1-log carbamazepine degradation with a UV/H2O2 system at different H2O2 concentrations and % UVT

To achieve 1-log removal of carbamazepine when treating the water with no additional pretreatment, the 20 year cost (capital and operating) of a UV/H2O2 system would range from

$20M (using 100 mg/L H2O2) to $46M (using only 10 mg/L H2O2). With an expense of $2M on coagulant pretreatment to improve the UVT from 64%/cm to 74%/cm, the costs of UV/H2O2 are

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reduced by 30-39% (savings of $6M-$18M over the 20 years, when using H2O2 doses of 100 mg/L and 10 mg/L, respectively). This demonstrates that coagulation to improve the UVT of the wastewater can have a cost-effective benefit on UV/H2O2 costs. If a simplistic assumption is made that the relationship between UVT and coagulant dose is linear, a further improvement in

UVT from 64%/cm to 84%/cm would reduce the costs of UV/H2O2 by 51-67% (savings of $10M-$31M), at a coagulant chemical cost of only $4M, again suggesting that the expense of increasing coagulant doses is more than offset by the resulting savings in UV/H2O2 treatment costs.

These costs did not consider the impact of increased sludge handling requirements when increasing the coagulant dose, and only accounted for the chemical costs. Nonetheless, given the large savings in UV/H2O2 costs for the small increase in coagulation costs, it is expected that coagulation pretreatment would remain a cost effective method to minimize the expense of

UV/H2O2. The model, however, also demonstrated that the benefit of coagulation was relatively smaller compared to the benefit of selecting an optimum H2O2 dose (Figure 5-5). By selecting

100 mg/L H2O2 instead of 10 mg/L H2O2, the 20 year cost of the UV/H2O2 system was reduced by 36-57%($6M-$27M) depending on the UVT of the wastewater. As such, H2O2 dose optimization is likely to be one of the key drivers when designing a UV/H2O2 system, and was more important that coagulant optimization in this case.

5.4 Conclusions

This study examined the effect of coagulation pretreatment prior to UV/H2O2 on the efficiency of treating micropollutants in a secondary wastewater. Efficiency was assessed by evaluating hydroxyl radical scavenging capacity and other water quality characteristics, compound degradation rates, energy requirements, and the treatment costs. Three coagulants (PACl, alum, and FeCl3) were used in this study which demonstrated:

1. PACl, alum, and FeCl3 coagulation significantly reduced the DOC concentration, UV254nm, overall scavenging capacity and EfOM scavenging capacity of the wastewater. 2. The high molecular weight organic components (biopolymers and humics) as well as building blocks of humics were the primary components removed by all three coagulants.

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3. Micropollutant degradation rates with UV photolysis and UV/H2O2 treatment generally increased by factors ranging from 1.14 to 2.59 with coagulation of the wastewater. Some

anomalous results were observed in cases usually associated with FeCl3 whereby coagulation led to slower degradation rates. It is believed that this is an artefact of either 3+ H2O2 decomposition by residual soluble Fe species in the wastewater thereby reducing the concentration of OH radicals, or experimental variability. 4. The electrical energy per order (EEO) requirements were generally reduced with

coagulation for both UV photolysis and UV/H2O2 treatment with reductions ranging from 0.8-72%. There was no significant difference among the coagulants.

5. Coagulation prior to UV/H2O2 treatment significantly reduced the net present worth of the

UV/H2O2 system by 30-67%, and the cost savings to UV/H2O2 exceeded the chemical costs of coagulation.

6. The cost of UV/H2O2 was more sensitive to H2O2 dose than to the degree of coagulation under the conditions modelled. This is not to say that coagulation should be ignored, but

rather that H2O2 optimization should be carefully considered.

Coagulation using ferric chloride, aluminium sulphate and polyaluminium chloride can be effectively used to pretreat secondary wastewater prior to UV/H2O2 treatment to improve oxidation efficiency of micropollutants. The benefits of this approach are significant in terms of wastewater quality, oxidation efficiency, energy requirements and the UV/H2O2 system costs.

Plants considering UV/H2O2 treatment for disinfection and the treatment of micropollutants should evaluate the two options of increasing the hydrogen peroxide dose with or without coagulation of the wastewater. The effect of coagulation in this study was only assessed on a single wastewater sample, therefore it cannot be conclusively stated that coagulation would always be a cost-effective option for pretreatment of all secondary wastewaters. Each wastewater must be evaluated on a case-by-case basis.

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5.5 References

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Bolton, J.R., Linden, K.G. (2003) Standardization of methods for UV fluence (UV Dose) determination in bench-scale UV experiments. Journal of Environmental Engineering 129(3): 209 – 215.

Bolton, J.R. (2004) Germicidal Fluence (UV Dose) Calculations for a Low Pressure UV Lamp. Edmonton, Alberta, Canada: Bolton Photosciences Inc.

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Crittenden, J.C., Hu, S., Hand, D.W., Green, S.A. (1999) A kinetic model for H2O2/UV process in a completely mixed batch reactor. Water Research 33(10): 2315 – 2328.

De Laat, J., Gallard, H. (1999) Catalytic decomposition of hydrogen peroxide by Fe(III) in homogeneous aqueous solution: Mechanism and kinetic modeling. Environmental Science & Technology 33(16): 2726 – 2732.

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Dong, M.M., Rosario-Ortiz, F.L. (2012) Photochemical formation of hydroxyl radical from effluent organic matter. Environmental Science & Technology 46(7): 3788 – 3794.

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Filloux, E., Gallard, H., Croue, J-P. (2012) Identification of effluent organic matter fractions responsible for low-pressure membrane fouling. Water Research 46(17): 5531 – 5540.

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Huber, S.A., Balz, A., Abert, M., Pronk, W. (2011) Characterisation of aquatic humic and non- humic matter with size exclusion chromatography–organic carbon detection-organic nitrogen detection (LC-OCD-OND). Water Research 45(2): 879 – 885.

Ikehata, K., Naghashkar, N. J., El-Din, M. G. (2006) Degradation of aqueous pharmaceuticals by ozonation and advanced oxidation processes: a review. Ozone: Science and Engineering 28(6): 353 – 414.

Ikehata, K., El-Din, M.G., Snyder, S.A. (2008) Ozonation and advanced oxidation treatment of emerging organic pollutants in water and wastewater. Ozone: Science and Engineering 30(1): 21 – 26.

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Kidd, K.A., Blanchfield, P.J., Mills, K.H., Evans, R.E., Lazorchak, J.M., Flick, R.W. (2007) Collapse of a fish population after exposure to a synthetic estrogen. Proceedings of the National Academy of Sciences of the United States of America 104(21): 8897 – 8901.

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Klassen, N.V., Marchington, D., McGowant, H.C.E. (1994) H2O2 determination by the method and by KMnO4 titration. Analytical Chemistry 66(18): 2921 – 2925.

Lee, Y., von Gunten, U. (2010) Oxidative transformation of micropollutants during municipal wastewater treatment: Comparison of the kinetic aspects of selective (chlorine, chlorine dioxide, ferrate (VI) and ozone and non-selective oxidants (hydroxyl radical). Water Research 44(2): 555 – 556.

Lee, E., Glover, C.M., Rosario-Ortiz, F.L. (2013) Photochemical formation of hydroxyl radical from effluent organic matter: Role of composition. Environmental Science & Technology 47(21): 12073 – 12080.

Lin, S-S., Gurol, M.D. (1998) Catalytic decomposition of hydrogen peroxide on iron oxide: Kinetics, mechanism, and implications. Environmental Science & Technology 32(10): 1417 – 1423.

Matilainen, A., Vepsalainen, M., Sillanpaa, M. (2010) Natural organic matter removal by coagulation during drinking water treatment: A review. Advances in Colloid and Interface Science 159(2): 189 – 197.

Mestre, A.S., Pires, J., Nogueira, J.M.F., Parra, J.B., Carvalho, A.P., Ania, C.O. (2009) Waste- derived activated carbons for removal of ibuprofen from solution: Role of surface chemistry and pore structure. Bioresource Technology 100(5): 1720 – 1726.

Metcalfe, C.D., Miao, X-S., Koenig, B-G., Struger, J. (2003) Distribution of acidic and neutral drugs in surface waters near sewage treatment plants in the Lower Great Lake, Canada. Environmental Toxicology and Chemistry 22(12): 2881 – 2889.

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Miao, X-S., Bishay, F., Chen, C., Metcalfe, C.D. (2004) Occurrence of antimicrobials in the final effluents of wastewater treatment plants in Canada. Environmental Science & Technology 38(13): 3533 – 3541.

Mompelat, S., Le Bot, B., Thomas, O. (2009) Occurrence and fate of pharmaceutical products and by-products, from resource to drinking water. Environment International 35(5): 803 – 814.

Ontario Ministry of the Environment (MOE) (2008) The Determination of Emerging Organic Pollutants in Environmental Matrices by LC-MS-MS Analysis, Laboratory Services Branch Method EOP-E3454, Etobicoke, Ontario.

Pereira, V.J., Weinberg, H.S., Linden, K.G. (2007) UV degradation kinetics and modelling of pharmaceutical compounds in laboratory garde and surface water via direct and indirect photolysis at 254 nm. Environmental Science & Technology 41(5): 1682 – 1688.

Radjenovic, J., Petrovic, M., Barcelo, D. (2007) Analysis of pharmaceuticals in wastewater and removal using a membrane bioreactor. Analytical and Bioanalytical Chemistry 387(4): 1365 – 1377.

Rosario-Ortiz, F.L., Wert, E.C., Snyder, S.A. (2010) Evaluation of UV/H2O2 treatment for the oxidation of pharmaceuticals in wastewater. Water Research 44(5): 1440 – 1448.

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Ryan, C.C., Tan, D.T., Arnold, W.A. (2011) Direct and indirect photolysis of sulphamethoxazole and trimethoprim in wastewater treatment plant effluent. Water Research 45(3): 1280 – 1286.

Schafer, A.I., Mastrup, M., Jensen, R.L. (2002) Particle interactions and removal of trace contaminants from water and wastewater. Desalination 147(1-3): 243 – 250.

Sharpless, C.M., Linden, K.G. (2003) Experimental and model comparisons of low and medium-pressure Hg lamps for the direct and H2O2 assisted UV photodegradation of N- nitrosodimethylamine in simulated drinking water. Environmental Science & Technology 37(9): 1933 – 1940.

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United States Enironmental Protection Agency 1997. Method 300.1 – Determination of Ionorganic Ions in Drinking Water by Ion Chromatography. Revision 1.0. Cincinatti, Ohio 45268: National Exposure Research Laboratory, Office of Research and Development.

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Yoon, Y., Westerhoff, P., Snyder, S.A., Wert, E., Yoon, J. (2007) Removal of endocrine disrupting compounds and pharmaceuticals by nanofiltration and ultrafiltration membranes. Desalination 202(1-3): 16 – 23.

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Chapter 6

UV/H2O2 Treatment of Municipal Wastewater: Ecotoxicological Effects

Abstract

Chemical oxidation of trace organic compounds in municipal wastewater can potentially cause negative ecotoxicological effects depending on the treatment used and the oxidation by- products/transformation products formed. This study evaluated the effect of UV/H2O2 on the acute toxicity, genotoxicity, and estrogenicity of a secondary wastewater effluent spiked with seven micropollutants (caffeine, carbamazepine, clofibric acid, diclofenac, 17-estradiol, naproxen, sulphamethoxazole) typically found in wastewaters. UV/H2O2 was applied both directly to the water or following pretreatment using coagulation with ferric chloride, aluminium sulphate, and polyaluminium chloride. UV/H2O2 with or without pretreatment did not induce acute toxicity effects based on the 48h-LC50 Daphnia magna and 96h-LC50 rainbow trout bioassays. Additionally, UV/H2O2 was not observed to cause an increase in the genotoxic potential or the estrogenicity of the wastewater.

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6.1 Introduction

Secondary wastewater effluents are primary point sources for trace organic compounds in the environment, such as pharmaceuticals and personal care products (Daughton et al., 1999; Stalter et al., 2010). Many of these compounds are known to have deleterious effects on aquatic organisms when present at sufficient concentrations (Bolong et al., 2009). Some compounds (e.g. 17β-estradiol, 17α-ethinylestradiol) are known endocrine disruptors that can cause significant estrogenic effects (Aguayo et al., 2004; Caliman et al., 2009; Metcalfe et al., 2001; Richard et al., 2014). Furthermore, disinfection of final effluents can give rise to estrogenic, toxic, mutagenic or genotoxic effects from oxidation by-products, particularly when chemical disinfection methods such as chlorination and ozonation are used (da Costa et al., 2014; Fukushima et al., 2014; Magdeburg et al., 2012 & 2014; Pignata et al., 2011; Watson et al., 2012; Wu et al., 2013). UV-advanced oxidation processes (UV-AOPs) can serve a dual role of wastewater disinfection and treatment, and are one technology that is very effective for treating these trace compounds (Autin et al., 2013; James et al., 2014; Schulze-Hennings et al., 2013; Shu et al., 2013). Nonetheless, complete mineralisation of trace organic compounds is generally not achieved with UV-AOPs, leading to the potential of forming transformation products which may retain toxic properties. Hence, it is useful to monitor the impact of UV-AOP treatment on the toxicity of the whole wastewater instead of, or in addition to, the degradation of the parent compound.

Ecotoxicological effects in wastewaters can be evaluated using different bioassays where an organism’s response on exposure to a contaminant is monitored. Bioassays can be classified according to the trophic level being assessed (microorganisms, plants and algae, invertebrates and fishes), however given the complexity of wastewater, different assays should be used concurrently to ensure that many different potential effects are identified (Magdeburg et al., 2014; Richard et al., 2014). Some typical effects of toxic components in water matrices are acute/chronic toxicity, estrogenicity, and genotoxicity. Acute toxicity in both water and wastewater is most commonly assessed using Vibrio fischeri or Daphnia magna because of their reproductive cycle and high sensitivity to a number of compounds including transformation products (Rizzo 2011). In Canada, the 96h-LC50 rainbow trout and the 48h-LC50 Daphnia

126 magna assays are commonly used for acute toxicity testing according the methods outlined by Environment Canada (2007) and (2000). Estrogenicity or the estrogenic effect of a sample is commonly assessed using in-vitro yeast estrogen screening (YES) assays since they are sensitive to any estrogenic activity irrespective of the compound (Lee et al., 2008; Salste et al., 2007). These assays are also capable of detecting the presence of estrogenic compounds that may be below the method detection limit for chemical analyses (Cespedes et al., 2004; Gaido et al., 1997; Rehmann et al., 1999), and so are suitable for evaluating estrogenic effects due to unknown oxidation by-products. Genotoxicity can be measured using the Comet assay (Heringa et al., 2011), umuC bioassay (Magdeburg et al., 2014) or commercially available test kits such as the EBPI SOS ChromoTestTM kit which was used in our study (Kovats et al., 2013; Zani et al., 2005) .

Only a few studies have examined the impact of UV/H2O2 treatment on potential ecotoxicological effects of wastewater. Reported studies have focused on toxicity in municipal effluent (Andreozzi et al., 2004; Souza et al., 2013; Yuan et al., 2011), toxicity in an industrial textile effluent (Arslan et al., 2009), and genotoxicity and estrogenicity in secondary wastewater spiked with either bisphenol A, ciprofloxacin, metoprolol or sulphamethoxazole (Richard et al., 2014). Findings from these studies noted that estrogenicity, acute toxicity, and genotoxicity effects were reduced with UV/H2O2 treatment, although Yuan et al. (2011) and Souza et al. (2013) reported that toxicity increased with UV treatment alone with exposure up to 3816 2 mJ/cm . While these studies demonstrate the suitability of UV/H2O2 treatment for different matrices, assessments were conducted using different target compounds, treatment approaches and bioassays. Additionally, variability in wastewater quality could result in the formation of different oxidation products, based on the specific compound, matrix effects and different degradation pathways, which subsequently influences the observed ecotoxicological effect.

There is limited information in the literature on transformation products from UV/H2O2 treatment of trace compounds in wastewaters, and their potential biological impact. Hence, an assessment of UV/H2O2 treatment of secondary effluent to evaluate ecotoxicological effects due to degradation of the parent compounds and any possible oxidation/transformation products formed would be useful, and would help to ensure that using UV/H2O2 for wastewater treatment and disinfection does not inadvertently induce toxicity in the treated water.

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The objective of this study was to determine whether UV/H2O2 treatment of a secondary wastewater spiked with elevated concentrations of typical pharmaceutical compounds would increase or decrease potential ecotoxicological effects (acute toxicity, genotoxicity, estrogenicity). The effect of coagulation pretreatment of the wastewater prior to UV/H2O2 treatment was also considered, as it has been shown that coagulation prior to UV/H2O2 can improve the degradation of some micropollutants, and therefore perhaps also reduce toxicity, or conversely, increase the formation of toxic by-products. Coagulation was performed using aluminium sulphate (alum), polyaluminium chloride (PACl) and ferric chloride (FeCl3).

6.2 Materials and Methods

6.2.1 Wastewater

Ideally, a study of this nature would involve multiple wastewaters to capture variations in composition; however, this study was limited to a single source due to practical constraints. The sample was a secondary municipal wastewater effluent collected from an Ontario (Canada) extended aeration wastewater treatment plant. Grab samples were collected immediately after the secondary clarifier, but prior to filtration and UV disinfection. The samples were collected in pre-cleaned 20 L polypropylene containers and stored at 4oC for a maximum period of six weeks until the experiments were completed. The samples were analyzed for pH, total and dissolved organic carbon (DOC) concentration, UV absorbance at 254 nm (UV254), and alkalinity. The wastewater characteristics are shown in Table 6-1. At the start of each experiment, the wastewater was analysed for UV transmittance (UVT) at 254 nm to verify its stability during storage, with the relative standard deviation found to be less than 2%.

6.2.2 Pretreatment

Pretreatment of the wastewater was performed in the laboratory using coagulation with + aluminium sulphate (alum), ferric chloride (FeCl3), and Hyper Ion (HI)705 polyaluminium chloride (PACl). The PACl is a high basicity (>80%) pre-hydrolyzed aluminium coagulant with an aluminium content of 11.3–12.1% by weight. Alum and FeCl3 were obtained from Sigma

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Aldrich with a purity of greater than 98%. Alum and FeCl3 coagulation were optimised at pH

6.0 using sulphuric acid (H2SO4) or sodium hydroxide (NaOH) for pH adjustment. The optimum doses for each coagulant were previously determined using a series of bench-scale experiments and applying a point-of-diminishing return (PODR) analysis where the PODR is the dose for which a 10 mg/L incremental increase in the applied dose of coagulant resulted in a change in DOC removal of less than 0.3 mg/L (USEPA 1999). Using the optimum doses for alum (12 mg

Al/L), FeCl3 (60 mg/L as FeCl3), and PACl (16 mg Al/L), pretreatment of the secondary wastewater was conducted in a 100L stainless steel tank at room temperature of 22 0C. Coagulation of the effluent was achieved using a rapid mix of 200 rpm for 1 minute, a slow mix at 30 rpm for 30 minutes followed by settling for 30 minutes.

Table 6-1: Water quality characteristics of the secondary wastewater Parameter Secondary Effluent pH 7.08 Temperature (oC) 20 -1 UVA254 (cm ) 0.130 % UVT (cm-1) 64.1 Conductivity (µS/cm) 1086

SUVA254 (L/mg-m) 2.21 TOC (mg-C/L) 9.2 DOC (mg-C/L) 8.1 TIC (mg-C/L) 37.7 2- Carbonate (mg CO3 /L) 0.07 - Bicarbonate (mg HCO3 /L) 113.2

Total Alkalinity (mg CaCO3/L) 186 Nitrite (mg/L-N) 0.21 Nitrate (mg/L-N) 20.5

6.2.3 Photolysis and UV/H2O2 Advanced Oxidation

A 42 L Calgon Carbon Rayox Advanced Oxidation Batch Pilot Reactor equipped with a 1 kW medium pressure Hg-lamp was used for UV photolysis and UV/H2O2 treatment of the wastewater. The wastewater (40L) was spiked with 50 µg/L of each pharmaceutical compound after coagulation pretreatment, but before UV or UV/H2O2 treatment. Samples for acute toxicity

129 and genotoxicity analyses were collected after the maximum exposure time of 5 minutes, which corresponded to a UV fluence of 3200 mJ/cm2. Samples (60L) for acute toxicity were couriered to the Environmental Laboratory at the Ontario Ministry of Environment and Climate Change for analysis. Genotoxicity samples (1L) were concentrated using solid phase extraction with the Oasis hydrophilic-lipophilic balance (HLB) extraction cartridges (Waters Oasis HLB 12 cc, 500 mg) after the samples were acidified to pH 2 using concentrated sulphuric acid. For estrogenicity, samples were collected at 0, 1 and 5 minute intervals and concentrated using solid phase extraction with Oasis hydrophilic-lipophilic balance (HLB) extraction cartridges (Waters  Oasis HLB 6 cc, 150 mg, 30 µm) before analysis. The residual H2O2 in all collected samples was quenched using sodium thiosulphate immediately after UV exposure and before any analysis.

6.2.4 Analytical Methods - Bioassays

Acute toxicity was measured using the 96h-LC50 Rainbow trout bioassay (Environment Canada

2007) and the 48h-LC50 Daphnia magna bioassay (MOE 1988). Estrogenicity or the estrogenic potential of the effluent was determined using the Yeast Estrogenicity Screening (YES) assay. The YES assay was conducted as previously described by Routledge et al. (1996), with minor modifications. The 96-well plates were incubated for 4 days before measuring the absorbance at 540 nm with a correction for turbidity at 630 nm using a BIO RAD Benchmark Plus equipped with Microplate Manager 5.2.1. The degree of colour change induced by successive dilutions of the test chemicals provided a measure of their estrogenic potency. The validity of the assay was confirmed by the response of the reference compound, 17β-estradiol (E2), that showed a -10 median effective concentration (EC50) of 2.16×10 M (n=13). Replicates of the samples were evaporated and re-dissolved in the assay medium and tested (n=3) in a range of 12 dilutions (1:2) with a row of ethanol blanks and E2 standards in each plate. The estrogenic potential/activity relative to the reference compound E2 was determined using the ratio of the effective concentration of the sample (EC50,sample) to the effective concentration of E2 (EC50,E2).

Genotoxicity was measured using the SOS ChromoTestTM (EBPI, Brampton, Canada) which uses the genetically modified Escherichia coli (E. coli) PQ37 strain where the bacterial SOS

130 promoter gene is linked to a gene responsible for -galactosidase (β-gal) production. Protein synthesis is also measured photometrically using alkaline phosphatase (AP). In this bacterial colorimetric bioassay, induction of -galactosidase is proportional to the amount of DNA damage that occurs and is assessed using a chromogenic substrate to form a blue colour. Quantification of β-gal induction (induction factor) is determined by the ratio of β-gal activity to AP activity. Serial dilutions of the SPE concentrated effluent samples were set-up in 96-well plates with a known carcinogenic positive control, 4-nitroquinoline 1-oxide (4-NQO) and a negative control (10% dimethyl sulphoxide (DMSO) saline solution) in each plate. Plates were incubated for 2 hours at 37oC following which chromogen for β-gal and AP were added to each well, and the plates incubated for an additional 1 hour. The optical density of the plates was measured at 605 nm (β-gal activity) and 420 nm (AP activity) using a Tecan Infinite M200 Plate reader (Morrisville, NC). Induction factors (IF) and toxicity equivalency values (TEQ) relative to the genotoxicity of 4-NQO were used in assessing the genotoxicity of the samples.

6.2.5 Chemicals and Materials

Chemical reagents (hydrogen peroxide (30 wt.%), sodium thiosulphate, acetonitrile (LC-MS grade), acetone (LC-MS grade), methanol (LC-MS grade), sodium acetate, sulphuric acid, dimethyl sulphoxide (DMSO), 4-nitroquinoline 1-oxide (4-NQO)) were obtained from Sigma Aldrich (Oakville, Canada) with a purity greater than 98%. Seven pharmaceuticals (carbamazepine, caffeine, clofibric acid, diclofenac, 17-estradiol, naproxen, sulphamethoxazole) were also obtained from Sigma Aldrich (Oakville, Canada) in a powdered form with a purity greater than 98%.

6.3 Results and Discussion

6.3.1 Effects on Acute Toxicity

In the absence of UV or UV/H2O2 treatment, the spiked wastewater sample with and without coagulation pretreatment exhibited no acute toxicity effects using the Daphnia magna and rainbow trout bioassays. This is shown by the LC50 (v/v) values of more than 100% for the initial conditions in Table 6-2 and Table 6-3. These values mean that without dilution (100%) of

131 the wastewater sample, a 50% mortality rate of the test organisms was not observed. These observations correlate well with previous studies as the spiked concentration in this study was less than the effective concentrations (EC50) previously identified for some of the compounds.

Quinn et al. (2008) reported the effective concentration (EC50) values for naproxen (1-10 mg/L),

CBZ (10-100 mg/L), SMZ and caffeine (EC50 > 100 mg/L) based on the European Union (EU)

Directive 93/67/EEC using the cnidarian, Hydra attenuata. It was noted that since the EC50 exceeded the typical concentrations in the environment, these compounds are not expected to induce any acute toxicity effects. Furthermore, Hernando et al. (2005) evaluated the acute toxicity of the wastewater effluent from nine wastewater plants, reporting that eight had an absence of acute toxicity when using the Daphnia magna bioassay, with the ninth plant identified as having problems with their treatment process. Therefore, it is not unexpected that wastewater effluents from well-operated facilities may have no acute toxicity effects even with the addition of the pharmaceutical compounds.

Similarly, the spiked wastewater exhibited no acute toxicity effects with UV oxidation at a 2 maximum fluence of 3200 mJ/cm or UV/H2O2 treatment at 20 mg/L of the coagulated and non- coagulated samples (Table 6-2 and Table 6-3). These observations demonstrate that although a high concentration (50 µg/L) of the pharmaceutical compounds was spiked into the wastewater, exceeding typical values in the environment in order to magnify any acute toxicity effects in the wastewater or that could be induced by the formation of oxidation by-products or transformation products, no acute toxicity effects were observed at any stage despite the UV and advanced oxidation treatment.

Table 6-2: 48h-LC50 values for Daphnia magna bioassay of the spiked wastewater

LC50 values (% v/v) Treatment Initial UV + 20 mg/L UV alone Condition H2O2 No pretreatment >100 >100 n.d.

FeCl3 n.d. >100 n.d. PACl >100 >100 >100 Alum >100 >100 >100 n.d. – no data due to sample loss

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Table 6-3: 96h-LC50 values for rainbow trout bioassay of the spiked wastewater LC50 values (% v/v) Treatment Initial UV + 20 mg/L UV alone Condition H2O2 No pretreatment >100 >100 >100 FeCl3 n.d. >100 >100 PACl >100 >100 >100 Alum >100 >100 >100 n.d. – no data due to sample loss

It is important to note that using wastewater as the matrix for spiking the mixture of pharmaceutical compounds would have introduced matrix effects due to the complex nature of wastewaters. Furthermore, the wastewater may have contained unknown constituents that could have induced acute toxicity effects before or after UV and UV/H2O2 treatment. However, as discussed earlier, the wastewater did not exhibit any toxicity effects after pharmaceutical addition, therefore any subsequently observed acute toxicity effects would have been due to UV or UV/H2O2 treatment. Hence, this study demonstrated that UV and UV/H2O2 treatment did not induce any acute toxicity effects in the wastewater used in this study. The absence of any acute toxicity effects with UV/H2O2 treatment was also reported by Andreozzi et al., (2004) who reported that a mixture of six pharmaceuticals (oxflaxacin, sulphamethoxazole, propanolol, carbamazepine, clofibric acid, and diclofenac) in synthetic water using a low-pressure UV lamp with an incident irradiance of 2.51×10-6 E/s and hydrogen peroxide concentrations of 5 mM

(170 mg/L) and 10 mM (340 mg/L) H2O2 did not result in an increase in toxicity.

6.3.2 Effects on Genotoxicity

Induction factors (IF) are useful for comparing the effects of different treatments on the SOS response of the bioassay where samples with IF values greater than 2.0 are considered to be genotoxic (Quillardet et al., 1982). Following UV photolysis (3,200 mJ/cm2) of the spiked wastewater, the coagulated samples were observed to have a 12-21% increase in the genotoxic potential when compared to the non-coagulated sample (Figure 6-1). While this increase is not significant at a 90% confidence interval (C.I.) based on a one-way analysis of variance

(ANOVA), the samples treated with FeCl3 and PACl had induction factors greater than 2.0. This

133 suggests that the coagulants could increase the genotoxic potential to such a degree that a genotoxicity response may be observed.

In the presence of UV/H2O2 treatment at 10 mg/L and 20 mg/L H2O2, lower genotoxic potentials were observed in the spiked wastewater with the increase in hydrogen peroxide concentration when compared to UV photolysis alone (Figure 6-2). These reductions (28-34%) were significant at 20 mg/L H2O2, based on a one-way ANOVA at a 90% confidence interval (C.I.) followed by Tukey’s Honestly Significant Difference (HSD) post hoc test. When considering the effect of coagulation prior to UV/H2O2 treatment, there was no significant difference in the genotoxic potential of the coagulated wastewater samples when compared to the non-pretreated sample.

Toxicity equivalency values (TEQ) relative to the genotoxicity of the carcinogen, 4-NQO, with a TEQ value of 1.0, demonstrated that all the samples in the study irrespective of the treatment applied were not as genotoxic as 4-NQO, as the TEQ values were less than 1.0 (Figire 6-3 and Figure 6-4). Nevertheless, with UV photolysis, the TEQ of the spiked coagulated wastewaters increased by factors of 1.0 to 1.6 when compared to the spiked no-pretreatment sample (Figure 6-3), but this increase was not found to be statistically significant at a 90% C.I. based on a one- way ANOVA followed by Tukey’s HSD post hoc test. When UV/H2O2 treatment was applied however, lower genotoxic potentials were formed irrespective of the coagulant used with reductions ranging from 23-73% (Figure 6-4) relative to UV photolysis. These reductions were statistically signifant at both 10 mg/L (23-51%) and 20 mg/L (52-73%) H2O2 based on a one- way ANOVA at 90% C.I. followed by Tukey’s HSD post hoc test.

134

3.0

2.5

IF = 2.0 2.0

1.5

1.0

0.5 Induction Factor (IF) at 2.5 eq mL mL eq 2.5 at (IF)Factor Induction

0.0 No pretreatment + Alum + UV FeCl3 + UV PACl + UV UV Figure 6-1: SOS Chromotest IF values for the spiked wastewater treated with UV photolysis (3200 mJ/cm2) alone (Error bars are the standard deviation of duplicate samples)

0 mg/L H2O2 10 mg/L H2O2 20 mg/L H2O2 3.0

2.5

IF = 2.0 2.0

1.5

1.0

Induction Factor (IF) at 2.5 eq mL mL eq 2.5 at (IF)Factor Induction 0.5

0.0 No pretreatment + Alum + UV/H2O2 FeCl3 + UV/H2O2 PACl + UV/H2O2 UV/H2O2

Figure 6-2: SOS Chromotest IF values of the spiked wastewater with UV/H2O2 treatment (Error bars are the standard deviation of duplicate samples)

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The genotoxic potentials formed at 20 mg/L H2O2 were also significantly lower (36-69%) in the coagulated samples when compared to no-pretreatment sample at a 90% C.I. based on a one- way ANOVA at 90% C.I. followed by Tukey’s HSD post hoc test. Although all the samples had TEQ values of less than 1.0 demonstrating that all the samples are not as genotoxic as the 4- NQO, the data suggests that the type of coagulant used could increase the genotoxic potential of the spiked wastewater. However, the study showed that UV/H2O2 treatment of the spiked wastewater (with 20 mg/L H2O2) can significantly lower the genotoxic potential formed.

1.1 4-NQO TEQ = 1.0 1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2

0.1

NQO Toxicity Equivalency Equivalency Units (TEQ) Toxicity NQO - 4 0.0 No pretreatment + Alum + UV FeCl3 + UV PACl + UV UV

Figure 6-3: Genotoxicity (4-NQO TEQ) of the spiked wastewater with UV photolysis (3200 mJ/cm2) alone (Error bars are the standard deviation of duplicate samples)

136

0 mg/L H2O2 10 mg/L H2O2 20 mg/L H2O2 1.1 4-NQO TEQ = 1.0 1.0

0.9

0.8

0.7

0.6

0.5

0.4

0.3

0.2

NQO Toxicity Equivalency Equivalency Units (TEQ) Toxicity NQO -

4 0.1

0.0 No pretreatment + Alum + UV/H2O2 FeCl3 + UV/H2O2 PACl + UV/H2O2 UV/H2O2

Figure 6-4: Genotoxicity (4-NQO TEQ) of the spiked wastewater with UV/H2O2 treatment (Error bars are the standard deviation of duplicate samples)

The observed increase in the genotoxic potential due to the coagulant used may be due to residual Fe3+ and Al3+ ions. These ions have been shown to cause DNA damage and can induce genotoxicity in human cells (Lima et al., 2007 & 2011; Sanders et al., 2014). Coagulation using 3+ 3+ FeCl3 and PACl will result in elevated concentrations of Fe and Al ions in the wastewater due to the 20 mg Fe/L and 16 mg Al/L doses applied in this study. With coagulation, not all of the Fe3+ and Al3+ ions would be removed during settling of the flocs, hence, the observed genotoxic response may be due to residual ion concentrations.

Variations in the genotoxic response can also occur based on the amount of organic matter removed during coagulation and the dose of coagulant used (Petala et al., 2006b; Zhang and Wang 2000). In comparing the two Al-based coagulants (PACl and alum), the genotoxic response of alum was lower based on IF values (5%) and TEQ values (25%) when compared to PACl (Figure 6-1 and Figure 6-3). This was attributed to the alum dose being 33% less (in terms of Al) than that of the PACl. The dissolved organic matter removed by alum (41%), FeCl3

137

(39%) and PACl (34%) were relatively similar, however, the FeCl3 (60 mg/L as FeCl3 or 21 mg Fe3+/L) dose exceeded that of PACl (16 mg Al/L) by 25% which may account for the higher observed genotoxic response with FeCl3. Therefore, ranking of the coagulants on the basis of applied dose gave FeCl3 > PACl > alum; and on the basis of the genotoxicity/genotoxic response, the coagulant ranking was the same with FeCl3 > PACl > alum.

Considering the IF and TEQ values, when coagulation pretreatment is applied, depending on the type of coagulant used, the genotoxic potential can increase such that the sample will be ™ considered genotoxic according to the SOS ChromoTest . However, with UV/H2O2 treatment at 20 mg/L H2O2, the genotoxic potential formed is significantly lower than in the presence of UV alone, such that the potential for the wastewater to be genotoxic is minimal. This is likely due to an increased availability of OH radicals for oxidising the pharmaceutical compounds and other constituents in the wastewater, as the overall scavenging capacity of the wastewater was reduced with coagulation.

6.3.3 Effects on Estrogenicity

Estrogenicity or estrogenic activity is a measure of the enzymatic response of estrogen receptors in the endocrine system that occurs in the presence of steroid hormones and/or other compounds which can mimic the action of these hormones. Estrogenic activity was assessed using relative estrogenic potency (REP) where REP = EC50(sample)/EC50(E2). The EC50 value for each treatment condition is the dilution factor that gives 50% of the maximum response of the positive control (E2) from the same plate and is usually obtained from E2 dose-response curves (Beck et al., 2006; Salste et al., 2007). The REPs at 0, 10 and 20 mg/L hydrogen peroxide for each treatment condition are shown in Figure 6-5, Figure 6-6, and Figure 6-7. As observed in the Figures, there was a general decrease in the REP with exposure time/UV fluence when subjected to UV photolysis or UV/H2O2 treatment under all the treatment conditions, with some exceptions. Exceptions were observed using alum treatment with UV photolysis, and the raw effluent sample at UV+ 10 mg/L H2O2 treatment. In both instances, there was an increase in the REPs which are considered to be anomalies resulting from errors during sample processing. Despite these, the general decrease observed was found to be significantly different at 1 minute

138

(UV fluence = 640 mJ/cm2) and 5 minutes exposure (UV fluence = 3200 mJ/cm2) when compared to the initial REP at t = 0 mins, based on a two-way analysis of variance (ANOVA) followed by Tukey’s Honestly Significant Difference (HSD) post hoc test at 90% C.I. No statistically significant differences were observed among the pretreatment options in the absence or presence of hydrogen peroxide for UV photolysis and UV/H2O2 treatment respectively, based on a two-way ANOVA followed by Tukey’s HSD post hoc test at a 90% C.I.

no pretreatment PACl FeCl3 Alum

2 90

O 2 80 70 60

50 (E2) (E2) at mg/L 0 H

50 40 30

20 (sample)/EC

50 10

EC 0 0 1 5 UV Exposure Time (minutes)

Figure 6-5: Change in relative estrogenicity of the spiked wastewater at 0 mg/L H2O2 with UV exposure time and coagulation

139

no pretreatment PACl FeCl3 Alum

2 90

O 2 80

50 (E2) (E2) at 10 mg/L H

50 40

(sample)/EC 10 50

EC 0 0 1 5 UV Exposure Time (minutes)

Figure 6-6: Change in relative estrogenicity of the spiked wastewater at 10 mg/L H2O2 with UV exposure time and coagulation

no pretreatment PACl FeCl3 Alum

2 O 2 80

50 (E2) (E2) at 20 mg/L H

50 40

20 (sample)/EC

50 10 sample lost lost sample EC 0 0 1 5 UV Exposure Time (minutes) Figure 6-7: Change in relative estrogenicity of the spiked wastewater at 20 mg/L with UV exposure time and coagulation

140

The decrease in estrogenicity of the samples with exposure time, irrespective of the applied hydrogen peroxide dose or treatment option indicates that UV photolysis is responsible for the reduction in estrogenic activity. Of the pharmaceutical compounds spiked into the wastewater, 17β-estradiol (E2) is a known estrogen and the reduction in the spike concentration was on average more than 95% for all the treatment conditions. The endocrine disrupting potency of E2 and estrone considerably exceeds (~ 500 times) that of other estrogenic hormones or compounds that exhibit similar characteristics (Ngheim et al., 2004; Pelissero et al., 1993). No detectable levels of E2 were observed in the raw wastewater sample, so this implies that the estrogenicity in the wastewater at t = 0 mins would be due to the elevated spike concentration of E2. Similar reductions in estrogenicity have been reported in other studies by Rosenfeldt et al. (2007), Sarkar et al. (2014), Whidbey et al. (2012) and Zhang et al. (2010). Rosenfeldt et al. (2007) reported a 99% reduction in the estrogenicity associated with E2 using a medium-pressure lamp and a UV fluence of 4000 mJ/cm2. The wastewater samples were not analysed for the presence of any transformation or oxidation by-products, however, it is unlikely that any potential oxidation by-products would have contributed additional estrogenic activity. Whidbey et al. (2012) and Rosenfeldt et al. (2007) both noted that the estrogenic effects of any potential by- products of E2 were less than the original parent compound and would not contribute to the estrogenic activity of the treated water matrix.

6.4 Conclusions

This study evaluated the changes that occur in potential ecotoxicological effects (acute toxicity, genotoxicity, estrogenicity) of a secondary wastewater effluent treated with UV photoloysis,

UV/H2O2, and coagulation combined with UV/H2O2. Our findings show:

1. UV photolysis and UV/H2O2 treatment did not induce any acute toxicity effects to the Daphnia magna and rainbow trout in the wastewater effluent used in this study.

2. Coagulation, particularly with FeCl3 and PACl, can increase the genotoxic potential of the wastewater from levels of no measurable genotoxicity to the lowest threshold where some potential genotoxicity may be observed.

141

3. UV/H2O2 treatment of the wastewater subjected to coagulation pretreatment can significantly lower the genotoxic potential formed during exposure. 4. UV photolysis was primarily responsible for reducing the estrogenicity associated with the estrogenic compound, 17β-estradiol (E2), irrespective of the hydrogen peroxide

concentration in UV/H2O2 treatment or the coagulation option applied.

5. With UV/H2O2 treatment, there was no observed increase in the estrogenicity of the wastewater suggesting either there was no apparent formation of oxidation by-products or any products formed were not sufficient to produce an estrogenic response.

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Routledge, E.J., Sumpter, J.P. (1996) Estrogenic activity of surfactants and some of their degradation products assessed using a recombinant yeast screen. Environmental Toxicology and Chemistry 15(3): 241 – 248.

Salste, L., Leskinen, P., Virta, M., Kronberg, L. (2007) Determination of estrogens and estrogenic activity in wastewater effluent by chemical analysis and the bioluminescent yeast assay. Science of the Total Environment 378(3): 343 – 351.

Sanders, T., Liu, Y-M., Tchounwou, P.B. (2014) Cytoxic, genotoxic, and neurotoxic effects of Mg, Pb, and Fe on pheochromocytoma (PC-12) cells. Environmental Toxicology DOI: 10.1002/tox.22014.

Sarkar, S., Ali, S., Rehmann, L., Nakhla, G., Ray, M.B. (2014) Degradation of estrone in water and wastewater by various advanced oxidation processes. Journal of Hazardous Materials 278: 16 – 24.

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Shu, Z., Bolton, J.R., Belosevic, M., El Din, M.G. (2013) Photodegradation of emerging micropollutants using the medium-pressure UV/H2O2 advanced oxidation process. Water Research 47(8): 2881 – 2889.

Souza, B.S., Dantas, R.F., Agullo-Barcelo, M., Lucena, F., Sans, C., Esplugas, S., Dezotti, M.

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Stalter, D., Magdeburg, A., Oehlmann, J. (2010) Compartive toxicity assessment of ozone and activated carbon treated sewage effluent using an in vivo test battery. Water Research 44(8): 2610 – 2620.

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Chapter 7 UV/H2O2 Treatment of Secondary Wastewater Effluents – Effect on Effluent Organic Matter Characteristics

Abstract

Effluent organic matter (EfOM) is considered to be a major hydroxyl radical scavenger. However, changes in its composition and characteristics during oxidation may alter its reactivity with the hydroxyl radical. An understanding of these changes is important to identify possible factors that influence EfOM scavenging capabilities. Changes in EfOM composition were evaluated during UV/H2O2 treatment at 0 mg/L, 10 mg/L, and 20 mg/L hydrogen peroxide using fluorescence spectroscopy followed by regional integration analysis. The fluorescence intensity of the high molecular weight components (humic-like substances and protein-like components) were primarily reduced during exposure to UV and UV/H2O2. Of the humic-like substances, which consist of humic and fulvic acids, fulvic acid-like components had the highest reduction rate with UV photolysis alone. The tryptophan protein-like components were significantly reduced during UV/H2O2 treatment of the wastewaters. This suggests that optimising the removal of high molecular weight and tryptophan protein-like components could reduce EfOM  scavenging of the OH radical, thereby improving UV/H2O2 performance for targeted micropollutant degradation.

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7.1 Introduction

Effluent organic matter (EfOM) is a complex heterogeneous mixture of natural organic matter (NOM) and soluble microbial products (SMP) that can cause problems such as reduced efficiency of treatment processes due to the oxidant demand it exerts, and precursors for disinfection by-product formation (Matilainen et al., 2010; Shon et al., 2006). The oxidant demand of EfOM is a particular concern for advanced oxidation processes (AOPs) where EfOM is typically the major scavenger of OH radicals and can constitute over 80% of the overall background scavenging capacity of a wastewater (Souza et al., 2014). Interest in the use of

AOPs, such as UV/H2O2 for tertiary treatment, has been growing in recent years due to its effectiveness for treating organic micropollutants. Therefore, an understanding of EfOM behaviour or characteristics on exposure to UV/H2O2 treatment is important in improving the effectiveness of UV-AOP for treating micropollutants whose presence in the environment are known to have negative effects on aquatic organisms (Jones et al., 2005; Lishman et al., 2006). This information is important as factors that influence EfOM reactivity and scavenging capabilities could be identified to facilitate the use of appropriate treatment processes optimised to reduce or remove these factors, and subsequently improve the performance of the UV/H2O2 system for treating the target compound in terms of degradation rate, energy, and operating costs.

Different analytical techniques have been used in previous studies to evaluate the changes that occur to organic matter when subjected to UV/H2O2 treatment and include liquid chromatography organic carbon detection (LC-OCD) analysis of EfOM (Gonzalez et al., 2013) and UV-VIS absorbance ratios of natural organic matter (NOM) (Sarathy et al., 2011). LC- OCD, size exclusion chromatography, and fluorescence spectroscopy have also been used for characterisation of organic matter (Audenaert et al., 2013; Henderson et al., 2011; Vakondis et al., 2014). Fluorescence spectroscopy has been used for characterising membrane foulants or monitoring changes in organic matter to identify fouling events, specifically because of its high sensitivity and very minor sample preparation requirements that ensures the integrity of the original sample (Pieris et al., 2008; Henderson et al., 2011; Peiris et al., 2010). In this current study, fluorescence spectroscopy is used to qualitatively and quantitatively assess the changes

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that occur in the composition of EfOM during oxidation with UV/H2O2 treatment of wastewater. Previous work on activated sludge wastewater by Gonzalez et al. (2013) using LC-OCD analysis, showed that UV/H2O2 treatment primarily degrades biopolymers, while humics and all the other low molecular weight components are retained in the treated wastewater. Biopolymers in LC-OCD analysis are classified as consisting of high molecular weight polysaccharides and proteins. However, the fluorescence excitation-emission matrices (FEEM) obtained in fluorescence spectroscopy can be analysed using multivariate data analysis methods, which provides additional qualitative and quantitative information on the organic constituents. FEEM analysis classifies the organic components as humic acid-like (humics), fulvic acid-like (fulvics), and different protein components such as tryptophan-like and tyrosine-like. It has also been demonstrated that the sensitivity of the fluorescence spectroscopy technique makes it effective for monitoring changes that occurs with organic matter in the natural environment (Peuravuon et al., 2002; Zhang et al., 2011). Hence, it is expected that this current study will complement previous studies and improve the understanding of the changes that occur in EfOM during UV/H2O2 treatment.

The objective of this study was to evaluate the changes that occur in the constituents of effluent organic matter of wastewater effluents when treated with UV/H2O2 at different hydrogen peroxide concentrations. Fluorescence spectroscopy analyses were used to characterise the wastewaters, and fluorescence regional integration (FRI) of the FEEM spectra was used to assess the changes in organic matter.

7.2 Materials and Methods

7.2.1 Wastewater Samples

Grab samples of secondary wastewater effluent were collected from four Ontario activated sludge (AS) wastewater treatment plants after secondary clarification but prior to any filtration or disinfection. The characteristics of the wastewaters used in the study are shown in Table 7-1. Samples were collected in pre-cleaned 20 L polypropylene containers and stored at 4oC until the experiments were conducted. All experiments were completed within seven days of collection. The chemical reagents (hydrogen peroxide (30 wt.%), sodium persulphate, phosphoric acid,

152 sulphuric acid) were obtained from Sigma Aldrich, Canada with a purity of greater than 98%. Stock solutions were prepared as required in Milli-Q water generated from a Milli-QUV Plus Ultrapure water system.

Table 7-1: Characteristics of the secondary wastewater effluents used in the study Parameter AS1 AS2 AS3 AS4 pH 7.3 6.9 6.8 7.0 Temperature (oC) 19 19 18 20 UVA @ 254 nm (cm-1) 0.119 0.125 0.120 0.125 % UVT (cm-1) 81 75 76 75

SUVA254 (L/mg-m) 2.20 2.31 1.80 2.36 TOC (mg/L-C) 4.41 5.17 6.84 4.12 DOC (mg/L-C) 4.13 5.06 6.67 3.83 TIC (mg/L-C) 26.5 14.3 18.53 34.2

Total Alkalinity (mg CaCO3/L) 144 107 128 188 Nitrite (mg/L-N) < 0.02* <0.02* <.0.02* 0.06 Nitrate (mg/L-N) 8 21 15 24 Phosphate (mg/L) < 0.07* <0.07* <0.07* < 0.07*

7.2.2 UV/H2O2 Advanced Oxidation Experiments

A Calgon Carbon® quasi-collimated beam apparatus (Model PSI-I-120, Calgon Carbon Corporation, USA) equipped with a 1 kW medium-pressure Hg-lamp (Calgon Carbon Corporation, Pittsburgh, PA, USA) was used for all experiments. A volume of 10 mL of wastewater was placed in a glass petri-dish with a small stirbar to ensure that the sample was well-mixed and homogeneous without creating a vortex or disturbing the surface of the sample.

Samples were spiked with hydrogen peroxide (H2O2) concentrations of 0, 10 and 20 mg/L and irradiated for the time duration required to achieve a UV fluence of 100, 200, 400, 600 and 800 mJ/cm2. All experiments were performed in duplicate. After irradiation, residual hydrogen peroxide in the samples were quenched using sodium thiosulphate, and the samples subsequently analysed using fluorescence spectroscopy to assess the changes that occurred in the EfOM components during the different treatment conditions.

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7.2.3 Analytical Equipment and Methods

Dissolved organic carbon (DOC) and total organic carbon (TOC) concentration were measured using a TOC analyser (O.I. Analytical Aurora Model 1030 with auto-sampler Model 1088) according to the Standard Method 5310C (APHA 1998). pH was measured using a calibrated pH meter (Thermo Scientific Orion Star A111). UV absorbance was measured using an Agilent 8453 UV-VIS spectrophotometer. Alkalinity was measured using Standard Method 2320B (APHA, 1998).

Fluorescence spectroscopy using a Perkin Elmer LS-50B fluorescence spectrofluorometer was used to generate fluorescence excitation emission matrices (FEEM) by scanning 1 nm increments of emission spectra (300 – 600 nm) at sequential 10 nm increments of excitation wavelengths between 200 – 500 nm using a quartz cuvette with 4 optical windows and a 1 cm pathlength. Spectra were collected at a photomultiplier voltage (PMT) = 775 V, scan rate = 600 nm/min, an excitation/emission slit width = 10 nm and a temperature of 22 0C. Raman scattering was eliminated and other background noise was reduced by subtracting the fluorescence spectra for ultra-pure Milli-Q water from all the sample fluorescence spectra. Since the same instrument was used for all analyses, the samples were not corrected for the excitation response of the lamp. Fluorescence spectroscopy generated a dataset of 120 FEEM for further analysis. Principal component analysis (PCA) was used to identify and characterise the key organic components in the wastewaters. This was performed on two of the wastewaters where the key components in both waters were identified as humic-like substances and protein-like substances. Given the similarity of the constituents of the two samples and agreement with previous studies where FEEM analysis showed that the key components in organic matter of wastewater effluents are humic-like and protein-like substances (Cohen et al., 2014; Li et al., 2014), an assumption was made that all four wastewaters would contain the same substances. Loading values from PCA were used to identify major organic components represented by the principal components (PCs).

Fluorescence regional integration (FRI) was applied to all four wastewaters to quantify the organic fractions of EfOM by determining the volume of the peak maxima in FEEM spectra.

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The FEEM matrices for each wastewater were delineated into four excitation-emission (Ex/Em) regions defined using excitation and emission boundaries with ±5 nm for excitation and emission, previously reported in the literature, and identifying the peaks in both the original fluorescence spectra as well as PCA loading plots. The regions were identified as humic acid- like (Ex/Em = 320/425) (Chen et al., 2003), fulvic acid-like (Ex/Em = 240/425) (Chen et al., 2003), tryptophan protein-like (Ex/Em = 280/350) (Murphy et al., 2008) and tyrosine protein- like (Ex/Em = 280/305) (Murphy et al., 2008). The change in each of the four organic components during treatment was assessed by determining the rate of change (slope) in the volume of the peak maxima for each component as a function of UV fluence by applying a first order exponential decay equation. These slopes (k) were determined in the absence of hydrogen peroxide (UV photolysis) and in the presence of 10 and 20 mg/L H2O2 (UV/H2O2 treatment).

7.3 Results and Discussion

The EfOM of the wastewaters used in this study consisted of humic-like and protein-like components based on fluorescence analyses. With PCA analysis and using the classification outlined in Table 7-2, Error! Reference source not found.the humic-like principal component PC1) was found to be responsible for most of the variability (40-44%) observed in the FEEM spectra of the untreated wastewater, where variability in PCA analysis refers to components that exhibited the most change/variation during treatment. Plots of the principal components and their associated variability identified from PCA are shown in Appendix K and Appendix L.

Table 7-2: Classification of principal components for the secondary wastewater effluents Principal Component Excitation/Emission (nm) Organic Classification PC1 280-400/350-550 Humic-like PC2 250-320/320-380 Tryptophan protein-like PC3 200-280/280-330 Aromatic protein-like (tyrosine)

For fluorescence regional integration (FRI) analysis, the rate at which each of the four fractions (humic acid-like, fulvic acid-like, tryptophan protein-like, and tyrosine protein-like) changed as a function of UV fluence in the absence or presence of H2O2 was determined by applying a first order exponential decay equation. The average slope (k) values for the FRI fraction volume as a

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function of UV fluence were determined with UV alone (Figure 7-1), with 10 mg/L H2O2

(Figure 7-2) and 20 mg/L H2O2 (Figure 7-3) for each of four wastewaters.

humic fulvic tryptophan tyrosine

14000

) 1

- 12000

L mgC L

2 10000

8000

6000

4000

with UV alone (AU cm alone (AU UV with 2000 (k) reduction ratesfraction Organic

0 AS1 AS2 AS3 AS4 Wastewater Plant

Figure 7-1: Reduction rates of the FRI fractions with UV photolysis alone (Error bars are the standard deviation of duplicate samples)

humic fulvic tryptophan tyrosine

14000 ) )

1 - 12000

L mgC L

2 10000

8000 (AU cm (AU

2 O 2 6000

4000

2000

H mg/L + 10 UV

(k) withreduction ratesfraction Organic 0 AS1 AS2 AS3 AS4

Wastewater Plant

Figure 7-2: Reduction rates of the FRI fractions with UV + 10 mg/L H2O2 treatment (Error bars are the standard deviation of duplicate samples)

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humic fulvic tryptophan tyrosine 14000

) 1 - 12000

mgC 2 10000

8000

cm (AU 2

O 2 6000

4000

2000 H +mg/L 20 UV fraction reduction (k)withrates Organic 0 AS1 AS2 AS3 AS4 Wastewater Plant

Figure 7-3: Reduction rates of the FRI fractions with UV + 20 mg/L H2O2 treatment (Error bars are the standard deviation of duplicate samples)

The slopes (k) of the four fractions were then evaluated using a mixed analysis of variance

(ANOVA) at a 90% confidence interval (C.I.) with H2O2 dose as the within-subjects effect and the fraction volume as the between-subjects effect. Mixed ANOVA analysis showed there was a significant interaction effect between the H2O2 dose and the volume of the EfOM fractions (based on the Pillai’s trace = 0.95, F(6, 24) = 3.65, p = 0.010, multivariate η2 = 0.48) indicating that there were differences in the degradation rate of the organic fractions at different H2O2 concentrations. This interaction effect was further investigated using a mains effect analysis. In the absence of hydrogen peroxide (UV alone), FRI analysis showed that fulvic acid-like components had the highest reduction rate (Table 7-3) and this was found to be statistically significant when compared to the rates for the other three fractions based on the mixed ANOVA analysis at 90% C.I. followed by a mains effect analysis. Ranking of the fractions based on degradation rates with UV only showed that fulvic acid-like > tryptophan protein-like > humic acid-like > tyrosine protein-like.

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Table 7-3: Reduction rates of the fraction volumes with UV alone

aReduction rate (k) of organic fraction volume (AU cm2 L mgC-1) Plant Humic Fulvic Tryptophan Tyrosine acid-like acid-like protein-like protein-like AS1 2633 ± 188 7007 ± 552 3066 ± 319 633 ± 22 AS2 3474 ± 223 6642 ± 406 4374 ± 85 564 ± 104 AS3 4820 ± 92 11163 ± 2077 6458 ± 241 1278 ± 209 AS4 4205 ±228 8097 ± 818 5632 ± 704 931 ± 246 aAverage ±standard deviation of duplicate samples

Fulvic acid and humic acid are the two constituents of humic substances that both contain aromatic and aliphatic structures, but differ in terms of molecular weight and functional group composition. Fulvic acid, relative to humic acid, has a lower molecular weight and a higher composition of carboxyl and hydroxyl groups rendering it more chemically reactive than humic acid (Pettit 2004; Weng et al., 2006). The humic acid also contains a higher fraction of aromatic structures which increases the stability of the chromophores, compared to fulvic acid. Therefore the fulvic acid would be predicted to degrade at a faster rate in the presence of UV, as observed in this current study where, on average, the reduction rate of the fulvic acid-like components was twice that of the humic-acid like components. This observation is in agreement with previous studies by Kulovaara et al. (1996), Corin et al. (1996), and Buckland (1992), which have shown that fulvic acid is more susceptible to UV degradation than humic acid. Notably, Allard et al. (1994) investigated the UV degradation of humic and fulvic acids prepared as individual solutions in Milli-Q water using 16 W/m2 UV at 254 nm and reported a 95% degradation of fulvic acid after 12 hours of exposure compared to 50% humic degradation after 30 hours.

Of the two protein-like components, on average, the reduction rate of tryptophan was significantly greater than that of tyrosine by a factor of 5.7 in the presence of UV alone, where tyrosine had the lowest reduction rate of all the four organic fractions (Table 7-3). Tryptophan and tyrosine are two aromatic amino acids typically found in proteins among the basic building blocks of the molecules, but they differ in terms of quantum yield, lifetime, and molar absorptivity of the fluorophore as shown in Table 7-4 (Lakowicz 2006). Due to the differences

158 in the characteristics of the amino acids, and the higher quantum yield, shorter lifetime, and higher molar absorptivity of tryptophan, it is expected that tryptophan protein-like components will degrade at a faster rate compared to tyrosine protein-like constituents as observed in this study. Furthermore, energy absorbed by tyrosine is usually transferred to tryptophan within the same protein (Lakowicz 2006) which may account for tyrosine having the lowest degradation rate of the four fractions.

Table 7-4: Fluorescence characteristics of tryptophan and tyrosine Absorption Fluorescence aLifetime Molar Compound Wavelength Wavelength Quantum (ns) Absorptivity (nm) (nm) Yield (M-1 cm-1) Tryptophan 2.6 280 5600 348 0.20 Tyrosine 3.6 274 1400 303 0.14 aLifetime is the average time a molecule stays in the excited state before returning to the ground state Source: Lakowicz 2006

With the presence of hydrogen peroxide in UV/H2O2 treatment and in contrast to UV only, the degradation rate of the tryptophan protein fraction was significantly greater than the other three fractions irrespective of the hydrogen peroxide dose applied based on a mixed ANOVA analysis at 90% C.I. followed by a mains effect analysis. The ranking of the organic fractions based on the degradation rates was the same at 10 and 20 mg/L H2O2 such that tryptophan protein-like > humic acid-like > fulvic acid-like > tyrosine protein-like (Table 7-5 and Table 7-6). The observed degradation rate of tryptophan is in agreement with the work conducted by Davies (1987) who reported that hydroxyl radicals cause significant structural changes in protein molecules, particularly tryptophan, resulting in significant degradation of the compound which is observed as a decrease in fluorescence intensity. Of all the protein molecules investigated, tryptophan showed the highest rate of reduction in fluorescence. It is suggested that this degradation is the result of fragmentation due to the abstraction of hydrogen from the amino acid α-carbon atoms by the hydroxyl radical as well as an unfolding of the protein molecules resulting in an increase in proteolytic susceptibility (Davies 1987).

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Table 7-5: Reduction rates (k) of the fraction volumes at UV + 10 mg/L H2O2

aReduction rate (k) of organic fraction volume (AU cm2 L mgC-1) Plant Humic Fulvic Tryptophan Tyrosine acid-like acid-like protein-like protein-like AS1 2385 ± 10 2780 ± 602 2662 ± 198 362 ± 60 AS2 3504 ± 252 2883 ± 86 4849 ± 437 699 ± 22 AS3 4920 ± 233 4071 ± 104 7196 ± 315 1182 ± 3 AS4 3301 ± 633 3313 ±991 4856 ± 49 935 ± 226 aAverage ± standard deviation of duplicate samples

Table 7-6: Reduction rates (k) of the fraction volumes at UV + 20 mg/L H2O2

aReduction rate (k) of organic fraction volume (AU cm2 L mgC-1) Plant Humic Fulvic Tryptophan Tyrosine acid-like acid-like protein-like protein-like AS1 2605 ± 22 1454 ± 226 3775 ± 325 759 ± 55 AS2 3884 ± 487 2875 ± 819 5331 ± 646 722 ± 114 AS3 4611 ± 266 2340 ± 1140 7439 ± 693 1133 ± 131 AS4 3958 ± 146 1712 ± 351 4309 ± 1011 553 ± 131 aAverage ± standard deviation of duplicate samples

Furthermore, while tyrosine and tryptophan protein-like components are present in EfOM, tryptophan is usually the prevalent fraction in high molecular weight organic fractions (Wu et al., 2003; Yamashita et al., 2003). Hoekstra (2007) also noted that molecules which have a high UV-excited fluorescence typically have conjugated double bonds. Additionally, Minakata et al. (2009) and Sudhakaran et al. (2013) reported that the hydroxyl radical preferentially reacts with the electron-rich sites on the carbon typical of aromatic compounds through OH addition at diffusion controlled rates. Therefore a high rate of removal of tryptophan protein-like components compared to the other fractions will occur in the presence of the hydroxyl as was observed in this study. This observation also implies that the protein, particularly those containing the tryptophan protein-like constituents may be the component responsible for the scavenging capacity of the EfOM.

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For the other organic fractions, the humic acid-like substances were degraded at a faster rate than that of the fulvic acid-like components in the presence of hydrogen peroxide, however, the difference in the rates for these two components were not found to be significantly different based on a mixed ANOVA at a 90% C.I. followed by a mains effects analysis. This is in contrast to the scenario with UV photolysis alone where the fulvic acid-like component was the most easily reduced faction. However, in the presence of OH radicals, the observed reversal in the degradation rates of the humics and fulvics is likely due to the higher molecular weight and concentration of aromatic structures of the humic acid.

This current study demonstrated that UV photolysis of the wastewater altered the humic-like substances of EfOM, specifically the fulvic acid-like components. However, UV/H2O2 treatment of the same waters resulted in significant removal of the tryptophan protein-like constituents. These findings are important as they provide an additional understanding of the changes that occur in organic matter during UV/H2O2 treatment and insight into the characteristics or specific components of EfOM that may be responsible for its scavenging capabilities of the OH radical.

Hence, when considering optimization of the performance of UV/H2O2 processes, consideration should be given to examine suitable methods of reducing the concentration of high molecular weight protein components (e.g. biopolymers) as a means of improving oxidation efficiency.

7.4 Conclusions

This study examined the effect of UV/H2O2 on the organic constituents of effluent organic matter using fluorescence spectroscopy. The change in the organic composition was evaluated under three conditions of UV alone, UV + 10 mg/L H2O2 and UV + 20 mg/L H2O2 using fluorescence regional integration of the 3D fluorescence spectra. The findings demonstrated:

1. Humic-like substances were the dominant organic fraction in all the wastewaters. 2. The fulvic acid-like component of humic substances was most susceptible to UV treatment and is reduction rate exceeds that of the other organic constituents.

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 3. In the presence of OH radicals with UV/H2O2 treatment, proteins (i.e., tryptophan-protein like substances) were the most readily oxidised or reduced.

The study implies that high molecular weight tryptophan protein-like constituents are the components primarily responsible for EfOM scavenging of the OH radical, and consideration should be given to reducing its concentration in wastewater as a means of reducing overall background scavenging capacity and improving the efficiency of advanced oxidation processes.

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Chapter 8 Conclusions and Recommendations for Future Research

8.1 Conclusion

The objective of this thesis was to evaluate modifications to a typical wastewater treatment process upstream of an AOP system that could be used to optimise the efficiency of UV/H2O2 AOP for treating micropollutants in secondary wastewaters. The underlying hypothesis for the overall study is that reducing the concentration of effluent organic matter (EfOM) in the wastewater would improve the UV/H2O2 efficiency in terms of degradation rates, energy, and cost requirements. It was also important to understand the extent to which EfOM, which is considered to be a major hydroxyl radical scavenger, influences the scavenging capacity of a wastewater matrix based on its characteristics.

In fulfilling the overall objective, the hypotheses evaluated and the main conclusions are outlined: 1. Effluent organic matter (EfOM) contributes the largest scavenging potential in both activated sludge (AS) and membrane bioreactor plants (MBR), but its reactivity will vary among different effluents. EfOM was found to contribute more than 70% of the overall scavenging capacity, except in instances where the wastewater had a high nitrite concentration. There was considerable variability in the scavenging potential among the different wastewaters. On average, the scavenging capacity of wastewater from AS systems was 22% greater than the scavenging capacity of MBR wastewaters where the scavenging capacity is mainly influenced by the EfOM reactivity with the hydroxyl radical and not the EfOM concentration. The average EfOM reaction rate constant of the 8 -1 -1 8 AS wastewaters (3.0 x 10 Mc s ) also exceeded that of the MBR (2.47 x 10 -1 -1 Mc s ).

2. Reducing the EfOM concentration will reduce the overall background scavenging capacity of the wastewater, improving degradation of the micropollutants while

168

reducing the energy and costs required for treating the compounds. Coagulation treatment using aluminium sulphate (alum), polyaluminium chloride (PACl) and ferric chloride significantly reduced the EfOM concentration and subsequently the overall background scavenging capacity of the wastewater. This led improvements in the degradation rate of the micropollutant compounds by factors ranging from 1.14 to 2.59. Energy requirements were reduced ranging from 0.8-72%. Improved rates and lower energy consumption resulted in significant reductions in the net present worth of an

UV/H2O2 system by 30-67%. The study also showed that the cost savings to the

UV/H2O2 system outweighed the coagulation expense.

3. The scavenging potential of EfOM is directly influenced by a specific component or characteristic. An evaluation of the organic components of EfOM of a sample of 10 wastewaters found a strong positive correlation between biopolymers and EfOM scavenging potential. Biopolymers may be a main contributor to the EfOM scavenging capabilities due to their molecular size and electron density. The EfOM hydroxyl radical

reaction rate (kOH,EfOM) had a significant positive correlation with specific UV absorbance (SUVA), but to none of the other parameters measured, indicating a preferential reaction with conjugated double bonds or aromatic structures. An evaluation

of the changes that occur in EfOM during exposure to UV/H2O2 treatment also showed that tryptophan protein-like components are significantly reduced during treatment, suggesting that these components are major radical scavengers.

4. UV/H2O2 treatment can either reduce or increase potential toxicological effects in the

whole effluent. UV/H2O2 did not induce any acute toxicity effects, reduces genotoxic potential, and will not cause an increase in the estrogenicity of the wastewater. While it was found that ferric chloride or polyaluminium chloride coagulants could increase the genotoxic potential of the wastewater to a degree where the sample may exhibit lower

thresholds of genotoxicity, UV/H2O2 treatment significantly reduced this genotoxic effect.

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8.2 Research Contributions

1. The was some evidence generated to suggest that MBR wastewaters may have a lower scavenging capacity than AS wastewaters, and may therefore be more easily treated using UV-AOPs. This is important in the design of new facilities or the upgrade of existing wastewater facilities where advanced oxidation may be under consideration for tertiary treatment and the treatment of emerging contaminants. While the selection of a treatment process is influenced by the wastewater quality characteristics and the desired treatment goals, coupling UV-AOP with membrane systems should be considered.

2. The overall scavenging capacity of wastewaters is primarily driven by the reactivity of the EfOM with the hydroxyl radical which is influenced by its composition, and not necessarily the EfOM concentration. Hence, high/low dissolved organic carbon (DOC) concentrations are not a reliable indicator of the scavenging capacity and treatability of a wastewater using AOPs.

3. The high molecular weight biopolymer component and tryptophan protein-like components are the constituents of EfOM that are mainly responsible for EfOM scavenging capability. Therefore, targeted removal of these components prior to UV-AOP treatment would improve wastewater quality and UV-AOP performance in terms of degradation rates and energy requirements.

4. Coagulation using ferric chloride, aluminium sulphate or polyaluminium chloride can be effectively used for targeted removal of the high molecular weight biopolymer and tryptophan protein-like components. Using coagulation as a pretreatment option for the wastewater will reduce EfOM scavenging capacity and increase the OH exposure in the wastewater.

5. Coagulation of the secondary effluent can be an effective upstream modification to the

wastewater treatment process prior to UV/H2O2 to improve the efficiency of UV/H2O2 in terms of oxidation, energy requirements and operating costs associated with using a

170

UV/H2O2 system. Facilties should also consider increasing the hydrogen peroxide dose used

during UV/H2O2 treatment as it can similarly improve AOP performance; however, the cost benefits are relatively similar to pretreatment of the wastewater when the hydrogen peroxide dose exceeds 50 mg/L.

6. Coagulation could increase the genotoxic potential of a wastewater; however subsequent

UV/H2O2 treatment may effectively reduce any potential formed. Additionally, UV/H2O2 treatment of the wastewater is not expected to increase or result in an increase in the estrogenicity of the treated wastewater.

8.3 Recommendations for Future Research

1. Experimental study which compares the effects of coagulation as a pretreatment method for wastewaters from both a membrane bioreactor (MBR) and an activated sludge (AS) system. In this study, MBRs were identified as possibly being more amenable to UV- AOP treatment (Chapter 4) and as such the coagulation treatment was subsequently

performed with an AS wastewater to improve its treatability with UV/H2O2 (Chapter 5). It would useful to compare whether pretreatment of the AS wastewater would provide

equivalent performance to UV/H2O2 treatment of MBR wastewater not subjected to pretreatment, and also to determine whether applying pretreatment to the MBR wastewater would provide any additional benefit in AOP performance.

2. Experimental study which compares the effects of pretreatment using coagulation on the

performance of UV/H2O2 treatment using medium-pressure UV and low-pressure high

output (LPHO) UV lamps. In this study, UV/H2O2 treatment was performed using a medium-pressure UV lamp. However, some wastewater treatment plants use low- pressure UV lamps such LPHO lamps for final UV disinfection of the effluent due to its high output in the germicidal UV range. It would be useful to identify and compare the benefits of pretreatment of the wastewater on using these two systems for treating micropollutants.

171

3. Experimental study to identify the effect of coagulant dose on the scavenging capacity of the treated wastewater. The study identified that the high molecular weight components (biopolymers) and the tryptophan-protein like components are the constituents most likely responsible for EfOM scavenging capability (Chapter 4 and Chapter 7). Chapter 5 showed that coagulation primarily removes these high molecular weight components, however, the coagulant doses were optimised for DOC removal. Previous studies have reported that very low coagulant doses are optimal for biopolymer removal. A study which evaluates the effect of the coagulant dose on both biopolymers and protein content of the wastewater and the resulting effect on the overall and EfOM scavenging capacity would be useful in maximising the removal of the constituents that contribute to EfOM scavenging.

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Appendix A: Endocrine Disruptors (EDCs)

Main Class Sub-class Compounds Source

Alachlor, mezachlor, trifluralin, diquat, atrazine, Falconer et al., 2006; Herbicides cyanazine, simazine, Ikehata et al., 2008; diuron, methoxychlor

Pesticides

Carbofuran, DDT, endosulfan, lindane, Falconer et al., 2006; Insecticide diazinon, malathion,, Ikehata et al., 2008 parathion

Nonylphenol, nonylphenol Surfactants --- etoxylate, octylphenol, Esplugas et al., 2007 octylphenol etoxylate

Steriods (Synthetic) Diethylstilbestrol, 17α- --- Esplugas et al., 2007 ethynylestradiol

Natural estrogens --- Estrone, 17β-estradiol Esplugas et al., 2007

Metals --- Cadmium, mercury, lead Esplugas et al., 2007

Diadzen, genistein, Phytoestrogens --- enterodiol and Esplugas et al., 2007 enterolactone

Tributylin, PCBs, dioxins Falconer et al., 2006; --- and furans, phthalates Esplugas et al., 2007

Industrial chemicals Hexabromocyclodecane, Brominated flame poly-brominated diphenyl Esplugas et al., 2007 retardants ethers and tetrabrromobisphenol A

173

Appendix B: Pharmaceutical and Personal Care Products (PPCPs)

Main Class Sub-Class Compounds Source

Amoxicillin, penicillinm Ikehata et al., 2008; β-Lactam sultamicillin, Caliman et al., 2009 ceftriaxone

Azithromycin, clarithromycin, erythromycin, Ikehata et al., 2008; Macrolide roxithromycin, Bhandari et al., 2009 lincomycin, tylosin

ofloxacin, Quinolone Ikehata et al., 2008 enrofloxacin

Antibiotics Sulfachlorpyridazine, sulfadiazine, sulfadimethoxine, sulfamerazine, sulfamethazine, Ikehata et al., 2008; Sulfonamides sulfamethizole, Bhandari et al., 2009 sulfamethoxazole, sulfimoxole, sulfathiazole, sulfasoxazole ciprofloxacin, enrofloxacin, Fluoroquinolones Bhandari et al., 2009 norfloxacin, sarafloxacin Carbadox, tetracycline, Other Ikehata et al., 2008 primidone, trimethoprim, Carbamezapine, Anticonvulsant and Ikehata et al., 2008; --- diazepam, primidone; antidepressants Caliman et al., 2009 fluoxetine Acetaminophen, Esplugas et al., 2007; Analgesics --- paracetomol, Caliman et al., 2009 Antipyretic, Non- diclofenac, Esplugas etal., 2007; --- steroidal anti- ibuprofen, naproxen, Ikehata et al., 2008

174 inflammatory drugs salicyclic acid, (NSAID) indomethacin

Atenolol, propranolol, sotalol, Ikehata et al., 2008; β-blocker --- celiprolol, Caliman et al., 2009 metoprolol, bisprolol

Cyclophophamide, Alkylating agents ifosfamide, Ikehata et al., 2008 melphalan Cytostatic drugs Azathioprine, Anti-metabolites cytarabine, 5- Ikehata et al., 2008 fluorouracil Natural hormones: estrone, 17β- estradiol, Oral Contraceptives* --- Diethylstilbestrol, Ikehata et al., 2008 17α-ethinylestradiol (synthetic contraceptive) Bezafibrate, clofibrate (clofibric Lipid Regulators --- Ikehata et al., 2008 acid), fenofibric acid, gemfibrozil Diatrizoate, Ikehata et al., 2008; X-ray Contrast media --- iomeprol, iopamidol, Caliman et al., 2009 iopentol, iopromide Musk xylol, musk ketone,, galaxolide, --- Bolong et al., 2009; Fragrances* tonalide, celestolide, Caliman et al., 2009 nitromusks, polycyclic musks Disinfectants/ --- Triclosan Bolong et al., 2009 Antiseptics Oxybenzone, Sunscreen agents --- Bhandari et al., 2009 octocrylene Alkyl- Preservatives Parabens Bolong et al., 2009 phydroxybenzoate

Stimulant --- Caffeine Snyder et al., 2003

Insect Repellents --- DEET, Bayrepel Bhandari et al., 2009

*Indicates compounds that are also potential endocrine disruptors

175

Appendix C: Occurrence of some EDCs and PPCPs in wastewater effluents globally Limit of Influent Effluent Removal Country Type of System Compound Detection Concentration Concentration Efficiency References (µg/L) (µg/L) (µg/L) (%) Preliminary and 17β-estradiol Ternes et al., 0.001 0.015 Not detected --- secondary (E2) 1999 clarification, aeration tank, Germany phosphate 0.009(median Ternes et al., Estrone 0.001 0.027 --- elimination using value) 1999 Fe(III)Cl3 or Fe(II)Cl2 83% (activated Estrone 0.001 0.040 --- sludge effluent) Primary and 17β-estradiol >99.9 (aerator Ternes et al., secondary 0.001 0.021 --- (E2) effluent) 1999 Brazil clarification, aerator 64% (trickling tank or biological 17α-ethinyl filter effluent); Ternes et al., trickling filter 0.001 ------estradiol 78% (activated 1999 sludge effluent) Salicyclic acid 0.087 13.7 0.106 --- Ibuprofen 0.061 8.45 0.384 --- Gemfibrozil 0.077 0.453 0.246 --- Naproxen 0.074 5.58 0.452 --- Ketoprofen 0.088 0.146 0.125 --- AS, lagoon, Diclofenac 0.062 0.204 0.194 --- Lishman et al., Canada AS + media Indomethacin 0.10 0.230 0.190 --- 2006 filtration Triclosan 0.031 1.93 0.108 --- Clofibric acid 0.066 Not detected Not detected --- Fenoprofen 0.066 Not detected Not detected --- Fenofibrate 0.026 Not detected Not detected --- Celestolide 0.016 0.0372 0.025 ---

176

Limit of Influent Effluent Removal Country Type of System Compound Detection Concentration Concentration Efficiency References (µg/L) (µg/L) (µg/L) (%) Traseolide 0.013 0.168 0.045 --- Galaxolide 0.012 2.031 0.751 --- AS, lagoon, Tonalide 0.0085 0.804 0.274 --- AS + media Estradiol 0.005 0.0083 < MDL --- filtration Lishman et al., Estrone 0.005 0.0295 0.0076 --- 2006 Canada 0.003 (median Ternes et al., AS Estrone 0.001 ------value) 1999 Ternes et al., AS 17β-estradiol 0.001 --- 0.006 --- 1999 17α- 0.009 Ternes et al., AS 0.001 ------ethinylestradiol 1999 4-nonylphenol --- 0.545 – 3.022 0.126 – 1.965 --- (NP) Pothitou et al., Greece AS NP1EO --- 0.466 – 4.025 0.013 – 0.573 --- 2008 NP2EO --- 0.490 – 2.670 0.026 – 0.216 --- Bisphenol A --- 0.468 – 0.857 0.020 – 0.048 --- Lipophilic 1.2 – 2.9 (in May metabolites (NP, – June 2007) Ahel et al., Switzerland STPs ------NP1EO, NP2EO) 0.7 – 4.1 (in 2000 February) Carboxylic nonylphenolic Ahel et al., Switzerland STPs metabolites ------5 – 20 --- 2000 (NP1EC, NP2EC, NP3EC) Notes: NP- nonylphenol; NP1EO – 4-nonylphenol monoethoxylate; NP2EO – 4-nonylphenol diethoxylate; NP1EC – nonylphenoxy acetic acid; NP2EC – nonylphenoxy (ethoxy) acetic acid; NP3EC – nonylphenoxy (diethoxy) acetic acid

177

Appendix D: Anion calibration graphs

250 70 60 200 y = 1.1451x - 0.5123 50 R² = 0.9998 y = 0.5916x + 1.0078 150 40 R² = 0.9974

100 30 20

50 Peak Area Peak for Chloride Peak Area Peak for Nitrate 10 0 0 0 50 100 150 200 250 0 20 40 60 80 100 120 Concentration (mg/L) Concentration (mg/L)

7 4.0 6 3.0 5 y = 0.5617x + 0.1592 y = 0.3878x - 0.2171 R² = 0.9981 4 R² = 0.9955 2.0 3

2 1.0 Peak Area Peak for Nitrite

1 Area forPeak Phosphate 0.0 0 0 2 4 6 8 10 12 0 2 4 6 8 10 12 Concentration (mg/L) Concentration (mg/L)

Figure D-1: Calibration graphs for anions in the wastewater

178

Appendix E: Micropollutant calibration graphs

17B-estradiol clofibric acid diclofenac 14.0

12.0 y = 0.2421[clofibric acid] - 0.1554 R² = 0.9943 10.0

8.0 y = 0.091 [diclofenac] + 0.2475 R² = 0.974 6.0

4.0

Area of Analyte/Area of Analyte/Area Area of Surrogate 2.0 y = 0.0858 [17B-estradiol] + 0.1155 R² = 0.983 0.0 0 10 20 30 40 50 60 Concentration (µg/L)

Figure E2-1: Calibration graphs for compounds analysed in the negative mode

caffeine carbamazepine (CBZ) sulphamethoxazole (SMZ) 7.0

6.0 y = 0.1155 [caffeine] + 0.0506 R² = 0.999 5.0

4.0 y = 0.0548 [SMZ] - 0.0181 R² = 0.9985 3.0

2.0 y = 0.0147 [CBZ] + 0.0023 R² = 0.9988 Area of Analyte/Area of Analyte/Area Area of Surrogate 1.0

0.0 0 10 20 30 40 50 60 Concentration (µg/L)

Figure E2-2: Calibration graphs for compounds analysed in the positive mode

179

80.0

70.0 y = 1.3286 [naproxen] - 0.4351 R² = 0.998 60.0

50.0

40.0 naproxen 30.0

20.0

10.0 Area of analyte/Area of Area analyte/Area of surrogate

0.0 0 10 20 30 40 50 60 Concentration (µg/L)

Figure E2-3: Calibration graph for naproxen (positive mode compound)

180

Appendix F: Effluent quality characteristics of the additional 23 plants sourced from the literature

Alkalinity UV254 DOC Nitrate Nitrite Plant # Plant ID pH (mg (cm-1) (mg/L) (mg-N/L) (mg-N/L) CaCO3/L) 1 1-1 7.1 0.109 5.81 13.1 0.1 84 2 1-2 7.2 0.101 5.67 15.2 0.0 80 3 1-3 6.9 0.111 5.96 11.2 0.0 82 4 2-1 7.0 0.105 6.49 3.8 0.0 102 5 2-2 6.7 0.149 7.97 5.6 0.0 120 6 3-1 6.4 0.129 6.34 12.8 0.0 65 7 3-2 7.4 0.121 6.28 2.3 0.0 130 8 3-3 7.1 0.155 8.13 4.2 0.0 126 9 3-4 6.8 0.147 8.60 4.0 0.1 123 10 3-1sc 7.1 0.130 8.91 1.3 0.1 168 11 3-2sc 6.9 0.163 8.97 2.0 0.0 81 12 4-1 6.3 0.106 5.34 13.3 0.0 150 13 4-2 7.4 0.115 5.76 2.3 0.0 188 14 4-3 6.9 0.105 5.41 0.9 0.0 191 15 5-1 6.3 0.127 6.41 14.4 0.0 45 16 5-2 7.0 0.158 7.46 13.1 0.1 51 17 6-1 7.0 0.174 7.75 0.2 0.1 200 18 8-1 7.2 0.115 5.29 2.4 0.0 60 19 8-2 6.8 0.124 7.20 6.4 0.0 52 20 8-3 6.9 0.115 5.62 4.5 0.0 57 21 lvnv 8.2 0.26 6.6 0.0 0.1 128 22 rmco 7.1 0.17 10.3 13.8 0.4 101 23 pcfl 7.6 0.26 10.3 9.38 0.8 269 Note: Plants 1-20 were sourced from Keen et al. (2014), Plants 21-23 were sourced from Rosario-Ortiz et al.(2010)

181

Appendix G: Correlation between EfOM scavenging capacity and DOC concentration for AS and MBR wastewaters

AS MBR

) ) 1.8

- s

5 1.6 1.4 R² = 0.2445 1.2 1.0 0.8 0.6 R² = 0.214 0.4 0.2

EfOM Scavenging Capacity (x (x 10 CapacityScavenging EfOM 0.0 0.0 2.0 4.0 6.0 8.0 DOC concentration (mg/L)

Figure G-1: Correlations between EfOM Scavenging Capacity and DOC Concentration

182

Appendix H: Dose-response curves for the coagulants

7.0 7.0 6.0 6.0 5.0 5.0 4.0 4.0 3.0 3.0 2.0 2.0

1.0 1.0 DOC ConcentrationDOC (mg/L) DOC Concentration Concentration DOC (mg/L) 0.0 0.0 0 5 10 15 20 0 20 40 60 80 100 120 Alum Dose (mg Al/L) FeCl3 Dose (mg/L)

Figure H-1: Alum dose response curve Figure H-2: FeCl3 dose response curve

9.0 8.0 7.0 6.0 5.0 4.0 3.0 2.0

DOC Concentration Concentration DOC (mg/L) 1.0 0.0 0 5 10 15 20 25 30 PACl Dose (mg Al/L)

Figure H-3: PACl dose response curve

183

Appendix I: Micropollutant degradation plots

Time (mins)

2 0 1 2 3 4 5 6 O

2 0.0

-0.5

-1.0

-1.5

-2.0

-2.5 ) for caffeine at mg/L 0 H

o -3.0

-3.5

log (C/Clog -4.0 No pretreatment PACl FeCl3 Alum

Figure I-1: Degradation of caffeine with UV alone (0 mg/L H2O2)

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0 -0.5 -1.0

-1.5

2 O

2 -2.0 H

-2.5 ) for caffeine ) for caffeine at 10 mg/L o -3.0 -3.5 log (C/C log -4.0 No pretreatment PACl FeCl3 Alum

Figure I-2: Degradation of caffeine at UV + 10 mg/L H2O2

184

Time (mins)

0.0 1.0 2.0 3.0 4.0 5.0 6.0

0.0

2 O 2 -0.5

-1.0

-1.5

-2.0

-2.5 ) for caffeine at 20 mg/L H o -3.0

-3.5 log (C/Clog -4.0

No pretreatment PACl FeCl3 Alum

Figure I-3: Degradation of caffeine with UV+20 mg/L H2O2

Time (mins) 0 1 2 3 4 5 6

2 0.0

O 2 -0.5

-1.0

-1.5

-2.0

-2.5

-3.0

) for naproxen ) for naproxen mg/L at 0 H o -3.5

-4.0 log (C/Clog No pretreatment PACl FeCl3

Figure I-4: Degradation of naproxen with UV alone (0 mg/L H2O2)

185

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0

0.0

2 O 2 -0.5

-1.0

-1.5

-2.0

-2.5

-3.0 ) for naproxen ) for naproxen at 10 mg/L H

o -3.5

-4.0 log(C/C No pretreatment PACl FeCl3

Figure I-5: Degradation of naproxen with UV+10 mg/L H2O2

Time (mins)

0.0 1.0 2.0 3.0 4.0 5.0 6.0

0.0

O 2 -0.5

-1.0

-1.5

-2.0

-2.5 ) for naproxen ) for naproxen at 20 mg/L H o -3.0

-3.5 log (C/C log

-4.0

No pretreatment PACl FeCl3

Figure I-6: Degradation of naproxen with UV+20 mg/L H2O2

186

Time (mins) 0 1 2 3 4 5 6

2 0.0

O 2 -0.5

-1.0

-1.5

-2.0

estradiol estradiol mg/L at 0 H -

β -2.5

-3.0

) for 17 o -3.5

log (C/Clog -4.0

Figure I-7: Degradation of 17β-estradiol with UV alone (0 mg/L H2O2)

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0

-0.5

-1.0

-1.5

estradiol estradiol at 10 mg/L -

H -2.0 β

-2.5 ) for 17

o -3.0

-3.5

log (C/Clog -4.0

No pretreatment PACl FeCl3 Alum

Figure I-8: Degradation of 17β-estradiol with UV + 10 mg/L H2O2

187

Time (mins)

0.0 1.0 2.0 3.0 4.0 5.0 6.0

2 0.0

O 2 -0.5

-1.0

-1.5

-2.0

estradiol estradiol at 20 mg/L H - -2.5

-3.0

) for 17B o -3.5

log (C/C log -4.0

No pretreatment PACl FeCl3 Alum Figure I-9: Degradation of 17B-estradiol with UV + 20 mg/L H2O2

Time (mins) 0.0 0.2 0.4 0.6 0.8 1.0 1.2 0.0 -0.5 -1.0

-1.5

2 O

2 -2.0 H

-2.5 ) for diclofenac ) for diclofenac at mg/L 0 o -3.0 -3.5

log log (C/C -4.0 No pretreatment PACl FeCl3 Alum

Figure I-10: Degradation of diclofenac with UV alone (0 mg/L H2O2)

188

Time (mins) 0.0 0.5 1.0 1.5 2.0 2.5 0.0

-0.5

-1.0

-1.5 2

O -2.0

2 H -2.5

) for diclofenac ) for diclofenac at 10 mg/L -3.0 o -3.5

log (C/C log -4.0 No pretreatment PACl FeCl3 Alum

Figure I-11: Degradation of diclofenac with UV + 10 mg/L H2O2

Time (mins)

0.0 0.5 1.0 1.5 2.0 2.5

2 O

2 0.0

-0.5

-1.0

-1.5

-2.0

-2.5

) for diclofenac ) for diclofenac at 20 mg/L H -3.0 o

-3.5 log (C/Clog -4.0

No pretreatment PACl FeCl3 Alum Figure I-12: Degradation of diclofenac with UV + 20 mg/L H2O2

189

Time (mins)

0 1 2 3 4 5 6 0.0

-0.5

-1.0

-1.5

2 O

2 -2.0 H -2.5

-3.0

) for carbamazepine ) for carbamazepine at 0 mg/L o -3.5

-4.0 log (C/Clog No pretreatment PACl FeCl3 Alum

Figure I-13: Degradation of carbamazepine with UV alone (0 mg/L H2O2)

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0 -0.5

-1.0

2 O 2 -1.5 -2.0

mg/L H mg/L -2.5 ) for carbamazpine ) for carbamazpine at 10 o -3.0 -3.5

log (C/Clog -4.0

No pretreatment PACl FeCl3 Alum

Figure I-14: Degradation of carbamazepine with UV + 10 mg/L H2O2

190

Time (mins)

2 2 O 2 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0

-0.5

-1.0

-1.5

-2.0

-2.5

-3.0

) for carbamazepine ) for carbamazepine at 20 mg/L H o -3.5

-4.0 log (C/Clog

No pretreatment PACl FeCl3 Alum

Figure I-15: Degradation of carbamazepine with UV + 20 mg/L H2O2

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0

-0.5

-1.0

2 -1.5

O 2 -2.0

mg/L H mg/L -2.5

) for sulphamethoxazole at 0 -3.0 o -3.5

log (C/Clog -4.0

No pretreatment PACl FeCl3 Alum

Figure I-16: Degradation of sulphamethoxazole with UV alone (0 mg/L H2O2)

191

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0 -0.5

2 -1.0

O 2 -1.5 -2.0

-2.5

) for sulphamethoxazole at 10 10 mg/L at H o -3.0 -3.5

log (C/Clog -4.0 No pretreatment PACl FeCl3 Alum

Figure I-17: Degradation of sulphamethoxazole with UV + 10 mg/L H2O2

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0

-0.5

-1.0

-1.5 2

O -2.0

2 H -2.5

-3.0

) for sulphamethoxazole at 20 mg/L o -3.5

-4.0 log (C/Clog

No pretreatment PACl FeCl3 Alum

Figure I-18: Degradation of sulphamethoxazole with UV + 20 mg/L H2O2

192

Time (mins) 0.0 1.0 2.0 3.0 4.0 5.0 6.0 0.0 -0.5 -1.0

-1.5 2

O -2.0

2 H -2.5

-3.0

) for clofibric acid mg/L at 0 o -3.5

-4.0 log (C/Clog No pretreatment PACl FeCl3 Alum

Figure I-19: Degradation of clofibric acid with UV alone (0 mg/L H2O2)

Time (mins)

0.0 1.0 2.0 3.0 4.0 5.0 6.0 2

O 0.0 2

-0.5

-1.0

-1.5

-2.0

-2.5

-3.0

) for clofibric acid at 10 mg/L H o -3.5

-4.0 log (C/Clog

No pretreatment PACl FeCl3 Alum

Figure I-20: Degradation of clofibric acid with UV + 10 mg/L H2O2

193

Time (mins)

0.0 1.0 2.0 3.0 4.0 5.0 6.0 2

O 0.0 2

-0.5

-1.0

-1.5

-2.0

-2.5

-3.0 ) for clofibric acid at 20 mg/L H o -3.5

-4.0 log (C/Clog

No pretreatment PACl FeCl3 Alum Figure I-21: Degradation of clofibric acid with UV + 20 mg/L H2O2

194

Appendix J: Absorption spectra for wastewater samples and pharmaceutical compounds

No pretreatment FeCl3 Alum PACl 3.5

3.0

2.5

1) -

2.0

1.5

1.0 UV absorbanceUV (cm

0.5

0.0 175 195 215 235 255 275 295 315

Wavelength (nm)

Figure J-1: Absorption spectra of the untreated and the coagulated wastewater effluent

195

No pretreatment FeCl3 Alum PACl CBZ caffeine clofibric acid diclofenac 17B-estradiol sulphamethoxazole naproxen 3.5

3.0 1) - 2.5

2.0

1.5

UV UV absorbance(cm 1.0

0.5

0.0 170 190 210 230 250 270 290 310 Wavelength (nm) Figure J-2: Absorption spectra of the pharmaceutical compounds, the untreated and the coagulated wastewater samples

196

Appendix K: Loading plots and percentage variability of AS1 principal components

PC1 – 39.8%

PC2 – 22%

PC3 – 12.2%

197

Appendix L: Loading plots and percentage variability of AS2 principal components

PC1 – 44.5%

PC5 – 1.16%

PC2 – 27.4%

PC3 – 5.2%