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Effects of metal salts on odour, pathogens, dewaterability and orthophosphate during the anaerobic digestion… Abbott, Timothy Lawrence 2014

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EFFECTS OF METAL SALTS ON ODOUR, PATHOGENS, DEWATERABILITY AND ORTHOPHOSPHATE DURING THE ANAEROBIC DIGESTION OF MUNICIPAL WASTE SLUDGE  by  Timothy Lawrence Abbott  B.A.Sc., The University of British Columbia, 2013  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF  Master of Applied Science  in  THE COLLEGE OF GRADUATE STUDIES  (Civil Engineering)   THE UNIVERSITY OF BRITISH COLUMBIA  (Okanagan)  December 2014   © Timothy Lawrence Abbott, 2014 Abstract This research explores the effects of using common metal salts to address some of the barriers to the wider adoption of anaerobic digestion (AD), which include the generation of corrosive and odourous volatile sulfur compounds (VSCs), pathogens remaining in the digestate, along with the re-release nutrients from the digested solids. The effects of chemical addition were studied by adding two different doses of ferric chloride (FC), aluminum sulfate (alum) (AL) and magnesium hydroxide (MG) to the feed of seven lab-scale semi-continuously fed anaerobic digesters. Digesters were operated at mesophilic temperatures (35 ± 2°C) at three different sludge retention times (SRTs) of 20, 12, and 7 days over a one-year period to assess the impacts of dosing on the aforementioned barriers to AD along with the effects of dosing on dewaterability of digested sludge and AD performance and stability. Both doses of FC were very effective in reducing VSCs by up to 87% versus an undosed (control) digester. Both doses of AL increased VSC levels significantly, with the higher dose increasing VSC levels to levels which are extremely dangerous to people and property. Neither level of MG addition had a statistically significant effect on VSCs. The effects of FC on pathogens were mixed with the lower dose showing a modest reduction in pathogens while the higher dose had the opposite effect. Both levels of AL were highly effective in reducing pathogens by up to 86% and may be a valuable tool pathogen reduction targets. Both levels of MG were shown to increase levels of total coliforms and E-coli in digested sludge. FC was far more effective in reducing nutrients in AD effluent versus the control and reduced soluble orthophosphate by nearly 61%. AL was even more effective achieving a near 70% reduction in orthophosphate, while MG had little effect on orthophosphate. None of the compounds appeared to have a significant adverse effect on AD operation at lower organic loading rates, although AL caused digester instability at higher doses and organic loading rates. Neither dose of any of the compounds had a statistically significant effect on sludge dewaterability.    ii Preface  Initial portions of this research were conducted as part of a fourth-year Capstone design project at the University of British Columbia. Cody Bagg, Dustin Creviston, and Joe Krego assisted with some of the preliminary literature review and laboratory work at the beginning of this project.  Portions of this work have been presented at: The 53rd Annual Conference of the Pacific Northwest International Section of the Air & Waste Management Association annual conference in Victoria, British Columbia, Canada (November 5-8, 2013); The University of British Columbia’s School of Engineering Graduate Research Symposium in Kelowna, British Columbia, Canada (May 1, 2014); Water Environment Federation’s (WEF) 2014 Residuals and Biosolids Conference in Austin, Texas, United States of America (May 18-21, 2014); British Columbia Water and Waste Association’s (BCWWA) 7th Annual Canadian Biosolids and Residuals Conference in Vancouver, British Columbia, Canada (June 4-7, 2014). A paper generated from the WEF 2014 Residuals and Biosolids Conference has been published as part of the conference proceedings. A paper summarizing the odour and pathogen results of this work have been submitted to the Journal Water Research and another is being prepared for publication.  This work has been carried out under the sponsorships of the Natural Science and Engineering Council of Canada (NSERC) Strategic Project Grant (#396519-10) and UBC Okanagan Research Grant (F12-01852).   iii Table of Contents  Abstract .................................................................................................................................... ii Preface ..................................................................................................................................... iii Table of Contents ................................................................................................................... iv List of Tables .......................................................................................................................... ix List of Figures .......................................................................................................................... x List of Equations ................................................................................................................... xii List of Abbreviations ........................................................................................................... xiii Acknowledgements .............................................................................................................. xiv Chapter 1 Introduction........................................................................................................... 1 1.1 Background ........................................................................................................................... 1 1.2 Motivation for Research ........................................................................................................ 2 1.2.1 City of Kelowna Wastewater Treatment Plant ................................................................. 3 1.2.2 Anaerobic Digestion ......................................................................................................... 5 1.3 Objectives ............................................................................................................................. 6 1.4 Thesis Outline ....................................................................................................................... 8 Chapter 2 Literature Review ................................................................................................. 9 2.1 Background ........................................................................................................................... 9 2.2 Wastewater Treatment ........................................................................................................ 11 2.2.1 Wastewater Treatment Plants ......................................................................................... 13 2.2.2 Preliminary Treatment .................................................................................................... 14 2.2.3 Primary Treatment .......................................................................................................... 15 2.2.4 Secondary Treatment ...................................................................................................... 15  Activated Sludge .................................................................................................... 15 2.2.4.1 Oxidation Ponds or Lagoons .................................................................................. 16 2.2.4.2 Tricking Filters ....................................................................................................... 16 2.2.4.3 Rotating Biological Contractors ............................................................................. 17 2.2.4.42.2.5 Tertiary Treatment .......................................................................................................... 17  Biological Nutrient Removal ................................................................................. 18 2.2.5.1 iv  Filtration ................................................................................................................. 19 2.2.5.2 Disinfection ............................................................................................................ 20 2.2.5.32.3 Waste Sludge Characteristics .............................................................................................. 21 2.3.1 Sludge Properties ............................................................................................................ 21 2.3.2 Physical and Chemical Characteristics of Sludge ........................................................... 22  Physical Properties of Sludge ................................................................................ 23 2.3.2.1 Sludge Dewaterability ............................................................................................ 23 2.3.2.2 Capillary Suction Time .......................................................................................... 23 2.3.2.3 Zeta Potential ......................................................................................................... 24 2.3.2.4 Sludge Rheology .................................................................................................... 24 2.3.2.52.3.3 Chemical Features of Sludge .......................................................................................... 25 2.3.4 Disposal .......................................................................................................................... 25 2.4 Stabilization ........................................................................................................................ 26 2.4.1 Methods of Stabilization ................................................................................................. 27  Alkaline Stabilization ............................................................................................. 27 2.4.1.1 Aerobic Digestion .................................................................................................. 27 2.4.1.2 Anaerobic Digestion .............................................................................................. 28 2.4.1.3 Influencing Factors in Anaerobic Digestion ..................................................... 30 2.4.1.3.1 Digestion Temperature ...................................................................................... 31 2.4.1.3.2 Digester pH, Volatile Fatty Acids and Alkalinity ............................................. 32 2.4.1.3.3 Sludge and Hydraulic Retention Time .............................................................. 32 2.4.1.3.4 Organic Loading Rate ....................................................................................... 33 2.4.1.3.5 Digester Performance ............................................................................................. 33 2.4.1.4 Problems with Anaerobic Digestion ...................................................................... 35 2.4.1.5 Odour from Wastewater Treatment .................................................................. 35 2.4.1.5.1 Volatile Sulfur Compounds............................................................................... 37 2.4.1.5.22.5 Production of Volatile Sulfur Compounds .......................................................................... 42  Inorganic sulfur ...................................................................................................... 43 2.5.1.1 Organic Sulfur ........................................................................................................ 43 2.5.1.2 Production of Hydrogen Sulfide and Methyl Mercaptan ....................................... 43 2.5.1.3 Creation and Destruction of Other VOSCs ............................................................ 44 2.5.1.4 Destruction of VOSCs ........................................................................................... 44 2.5.1.52.6 Pathogenic Microorganisms ................................................................................................ 45 2.7 Nutrients .............................................................................................................................. 46  v 2.8 Metal Salts to Control Odour .............................................................................................. 47 Chapter 3 Materials and Methods....................................................................................... 49 3.1 Experimental Design ........................................................................................................... 49 3.2 Metal Salt Selection ............................................................................................................ 50 3.2.1 Metal Doses .................................................................................................................... 51 3.2.2 Existing Metal Concentrations ....................................................................................... 51 3.3 Anaerobic Digester Configuration ...................................................................................... 52 3.4 Inoculum and Acclimatization ............................................................................................ 52 3.5 Feed Sludge ......................................................................................................................... 54 3.6 Key Metrics Measured ........................................................................................................ 56 3.6.1 Volume and Composition of Biogas ............................................................................... 56 3.6.2 Biogas Volumes .............................................................................................................. 56 3.6.3 Biogas Composition........................................................................................................ 56 3.6.4 Volatile Sulfur Compounds ............................................................................................ 57 3.6.5 Most Probable Number Analysis for Pathogenic Microorganism Enumeration ............ 58 3.6.6 Dewaterability ................................................................................................................ 59  Capillary Suction Time .......................................................................................... 60 3.6.6.1 Zeta Potential and Conductivity ............................................................................. 60 3.6.6.23.6.7 Nutrient Levels ............................................................................................................... 61 3.6.8 Particle Size .................................................................................................................... 61 3.6.9 Sludge Dynamic Viscosity ............................................................................................. 62 3.7 Digester Performance .......................................................................................................... 63 3.7.1 Chemical Oxygen Demand ............................................................................................. 63 3.7.2 Solids .............................................................................................................................. 64 3.8 Digester Operational Metrics .............................................................................................. 65 3.8.1 pH and Alkalinity ........................................................................................................... 65 3.8.2 Volatile Fatty Acids ........................................................................................................ 66 3.8.3 Protein and Humic Acid ................................................................................................. 66 Chapter 4 Results and Discussion ....................................................................................... 68 4.1 Characterization of Waste Sludge Streams ......................................................................... 68 4.2 Species of VSCs in AD Biogas ........................................................................................... 69 4.3 Effects of Dosing on VSCs ................................................................................................. 71 4.3.1 Ferric Chloride ................................................................................................................ 76  vi 4.3.2 Aluminum Sulfate ........................................................................................................... 76 4.3.3 Magnesium Hydroxide ................................................................................................... 78 4.4 Effect of Dosing on Pathogens ............................................................................................ 78 4.4.1 Ferric Chloride ................................................................................................................ 80 4.4.2 Aluminum Sulfate ........................................................................................................... 80 4.4.3 Magnesium Hydroxide ................................................................................................... 81 4.5 Effects of dosing on dewaterability..................................................................................... 82 4.5.1 Capillary Suction Time ................................................................................................... 82 4.5.2 Zeta Potential .................................................................................................................. 84 4.6 Effects of Dosing on Nutrients ............................................................................................ 86 4.6.1 Ferric Chloride ................................................................................................................ 89 4.6.2 Aluminum Sulfate ........................................................................................................... 89 4.6.3 Magnesium Hydroxide ................................................................................................... 90 4.7 Effects of Dosing on Protein/Humic Acid .......................................................................... 90 4.8 Effects of Dosing on Digester Performance ........................................................................ 93 4.8.1 TS/VS Removals ............................................................................................................ 93  Ferric Chloride ....................................................................................................... 97 4.8.1.1 Aluminum Sulfate .................................................................................................. 97 4.8.1.2 Magnesium Hydroxide ........................................................................................... 98 4.8.1.34.8.2 Chemical Oxygen Demand Removals ............................................................................ 98 4.8.3 Biogas Production ......................................................................................................... 100 4.8.4 Effects of Dosing on Sludge Dynamic Viscosity ......................................................... 103 4.8.5 Effects of Dosing on Sludge Particle Size .................................................................... 103 4.9 Effects of Dosing on Digester Operation and Stability ..................................................... 105 4.9.1 pH ................................................................................................................................. 108 4.9.2 VFA & Alkalinity ......................................................................................................... 108 4.10 Effects of Dosing on Digester Stability ............................................................................ 109 Chapter 5 Conclusion ......................................................................................................... 110 5.1 Summary ........................................................................................................................... 110 5.1.1 Volatile Sulfur Compounds .......................................................................................... 110 5.1.2 Pathogens ...................................................................................................................... 111 5.1.3 Nutrients ....................................................................................................................... 112 5.1.4 Digester Operation and Stability .................................................................................. 112 5.2 Practical Implications of Results ....................................................................................... 112  vii 5.3 Recommendations for Future Work .................................................................................. 113 References ............................................................................................................................ 115 Appendices ........................................................................................................................... 125 Appendix A: British Columbia Organic Matter Recycling Regulations ........................................ 125 Appendix B: Sample Chemical Oxygen Demand Calibration Curve ............................................ 126 Appendix C: Example Protein and Humic Acid Calibration Curves ............................................. 127 Appendix D: Average Total Orthophosphate Levels in mg P/L .................................................... 128 Appendix E: Average Protein and Humic Acid Levels ................................................................. 129 Appendix F: Sludge Dynamic Viscosity ........................................................................................ 130 Appendix G: Effluent Particle Size Distribution............................................................................ 131 Appendix H: Average digester biogas composition (%) ................................................................ 132 Appendix I: Normal Probability Plots of ANOVA Residuals ....................................................... 133  (Appels, Baeyens, Degrève, & Dewil, 2008; Crittenden, Trussell, Hand, Howe, & Tchobanoglous, 2012; Grady, Daigger, & Filipe, 2011; Koh, Shaw, & Tarallo, 2010; Makowska, Spychała, & Mazur, 2013; Metcalf & Eddy, Tchobanoglous, Burton, & Stensel, 2002; Mori, Seyssiecq, & Roche, 2006; Pérez-Elvira, Fdz-Polanco, & Fdz-Polanco, 2011; Pitman, Deacon, & Alexander, 1991; Seyssiecq, Ferrasse, & Roche, 2003; Swanson, Bortman, O’Connor, & Stanford, 2004; Travnicek, Vitez, & Junga, 2013; Wang, Hung, & Shammas, 2007; Wu, Bishop, & Keener, 2005) (Bolzonella, Pavan, Battistoni, & Cecchi, 2005; Hwang, Jang, Hyun, & Kim, 2004; Jefferson, Hurst, Stuetz, & Parsons, 2002; Morgan, Evison, & Forster, 1991; Oropeza, Cabirol, Ortega, Ortiz, & Noyola, 1996; Sten Persson, Ediund, Claesson, & Carlsson, 1990; Salsali & Parker, 2006; Sanin, Clarkson, & Vesilind, 2010; Smith, Lang, Cheung, & Spanoudaki, 2005; Yoshimura, Nakano, Yamashita, Oho, Takahiko, & Koga, 2000)(Cabirol, Barragán, Durán, & Noyola, 2003; Ye Chen, Cheng, & Creamer, 2008; Fytianos, Voudrias, & Raikos, 1998; Wu, Bishop, Keener, Stallard, & Stile, 2001) viii List of Tables  Table 2.1 - Composition of untreated domestic wastewater ................................................... 12 Table 2.2 - Sludge properties .................................................................................................. 22 Table 2.3 - Sludge nutrient levels ........................................................................................... 25 Table 2.4 - Occupational health and safety exposure standards ............................................. 36 Table 2.5 - Odour thresholds of common VSCs ..................................................................... 37 Table 2.6 – Short-term health symptoms and effects of hydrogen sulfide exposure .............. 39 Table 2.7 - Common odourous volatile sulfur compounds..................................................... 41 Table 3.1 - Feed sludge metal concentrations ......................................................................... 51 Table 3.2 - Digester sludge retention times ............................................................................ 54 Table 3.3 - Feed sludge characteristics ................................................................................... 55 Table 4.1 – Average feed sludge characteristics ..................................................................... 68 Table 4.2 – Average steady-state digester volatile sulfur compound concentrations in ppm . 72 Table 4.3 - ANOVA of total volatile sulfur compound results ............................................... 74 Table 4.4 – ANOVA of digestate pathogen concentrations (MPN/gram dry weight) ............ 79 Table 4.5 - ANOVA for capillary suction time (CST) results ................................................ 83 Table 4.6 – ANOVA for digestate zeta potential results ........................................................ 85 Table 4.7 - ANOVA for digestate orthophosphate results ...................................................... 87 Table 4.8 - ANOVA for digestate protein results ................................................................... 90 Table 4.9 - ANOVA for digestate humic acid results ............................................................. 91 Table 4.10 - ANOVA for digester COD results ..................................................................... 99 Table 4.11 –ANOVA for dynamic viscosity of raw and digested sludge ............................ 103 Table 4.12 - Steady-state average digester pH, alkalinity, ammonia and VFA levels ......... 106   ix List of Figures  Figure 1.1 – Simplified treatment flow diagram of Kelowna’s wastewater treatment facility . 4 Figure 2.1 - Common wastewater treatment configuration .................................................... 13 Figure 2.2 - Biological nitrification and denitrification .......................................................... 19 Figure 2.3 – Stages of anaerobic digestion ............................................................................. 29 Figure 2.4 - Dissociation of sulfide between pH 4 & 10 ........................................................ 40 Figure 2.5 - Sulfur cycle ......................................................................................................... 42 Figure 3.1 - Full-factorial experimental design ...................................................................... 50 Figure 3.2 - Laboratory scale anaerobic digesters .................................................................. 52 Figure 3.3 - City of Kelowna wastewater treatment processes ............................................... 55 Figure 4.1 - Average volatile sulfur compound composition ................................................. 70 Figure 4.2 - Volatile sulfur levels after 10 mg/L alum digester failure .................................. 71 Figure 4.3 – Stage One digester headspace total volatile sulfur compound concentrations ... 75 Figure 4.4 - Stage Two digester headspace total volatile sulfur compound concentrations ... 75 Figure 4.5 – Stage One digestate total coliform densities ...................................................... 79 Figure 4.6 - Stage Two digestate total coliform densities ...................................................... 80 Figure 4.7 - Stage One digestate average capillary suction time (CST) results ..................... 83 Figure 4.8 - Stage Two digestate capillary suction time (CST) results .................................. 84 Figure 4.9 - Stage One digestate zeta potential....................................................................... 85 Figure 4.10 - Stage Two digestate zeta potential .................................................................... 86 Figure 4.11 - Stage One digestate total orthophosphate concentrations ................................. 87 Figure 4.12 - Stage Two digestate soluble orthophosphate concentrations ............................ 88 Figure 4.13 - Stage One digestate average humic acid concentrations ................................... 92 Figure 4.14 - Stage Two digestate average humic acid concentrations .................................. 92 Figure 4.15 - Stage One total solids (TS) removals of digesters ............................................ 93 Figure 4.16 - Stage Two total solids (TS) removals of digesters............................................ 94 Figure 4.17 - Stage One volatile solids (VS) removals of digesters ....................................... 94 Figure 4.18 - Stage Two volatile solids (VS) removals of digesters ...................................... 95 Figure 4.19 - Stage One VS/TS ratios of digestates ............................................................... 96 Figure 4.20 - Stage Two VS/TS ratios of digestate ................................................................ 96  x Figure 4.21 - Stage One digester chemical oxygen demand (COD) removal ........................ 99 Figure 4.22 - Stage Two digester chemical oxygen demand (COD) removal ...................... 100 Figure 4.23 - Digester biogas composition ........................................................................... 101 Figure 4.24 - Stage One digester normalized methane volumes .......................................... 102 Figure 4.25 - Stage Two digester normalized methane volumes .......................................... 102 Figure 4.26 – Stage One digester particle size distribution .................................................. 104 Figure 4.27 – Stage Two digester particle size distribution.................................................. 104 Figure A.1 - ANOVA normal probability plot of TVSC residuals....................................... 133 Figure A.2 - ANOVA normal probability plot of MPN residuals ........................................ 133 Figure A.3 - ANOVA normal probability plot of CST residuals ......................................... 134 Figure A.4 - ANOVA normal probability plot of Zeta potential residuals ........................... 134 Figure A.5 - ANOVA normal probability plot of orthophosphate residuals ........................ 135 Figure A.6 - ANOVA normal probability plot of protein residuals ..................................... 135 Figure A.7 - ANOVA normal probability plot of humic acid residuals ............................... 136 Figure A.8 - ANOVA normal probability plot of COD residuals ........................................ 136 Figure A.9 - ANOVA normal probability plot of dynamic viscosity residuals .................... 137     xi List of Equations  Equation 2.1 – Methane production using hydrogen .............................................................. 30 Equation 2.2 – Methane production using acetate .................................................................. 30 Equation 2.3 - Reduction of sulfate to sulfide ........................................................................ 43 Equation 2.4 – Methylation of hydrogen sulfide to form methyl mercaptan.......................... 44 Equation 2.5 - Methylation of methyl mercaptan to form dimethyl sulfide ........................... 44 Equation 2.6 – Oxidation of methyl mercaptan to form dimethyl disulfide ........................... 44 Equation 2.7 - Demethylation of dimethyl disulfide .............................................................. 45 Equation 3.1 – Total solids calculation ................................................................................... 65 Equation 3.2 – Volatile solids calculation .............................................................................. 65 Equation 3.3 – Alkalinity calculation ..................................................................................... 65 Equation 3.4 – Total absorbance ............................................................................................. 67 Equation 3.5 – Blank absorbance ............................................................................................ 67 Equation 3.6 – Protein absorbance .......................................................................................... 67 Equation 3.7 – Humic acid absorbance................................................................................... 67  xii List of Abbreviations ACS - American Chemical Society ANOVA – Analysis of Variance BOD – Biochemical Oxygen Demand BNR – Biological Nutrient Removal  CL – Confidence Limit COD – Chemical Oxygen Demand CST – Capillary Suction Time GC – Gas Chromatograph MLSS - Mixed Liquor Suspended Solids MPN – Most Probable Number SRT – Sludge Retention Time STP – Standard Temperature & Pressure TS – Total Solids VS – Volatile Solids VFA – Volatile Fatty Acids VOSC – Volatile Organic Sulfur Compounds VSC – Volatile Sulfur Compounds TVSC – Total Volatile Sulfur Compounds WWTP – Wastewater Treatment Plant ZP – Zeta Potential  xiii Acknowledgements I would like to express my gratitude to my supervisor, Dr. Cigdem Eskicioglu for her patience, guidance and encouragement throughout this process. She has been an invaluable source of knowledge and I will be forever grateful that she was willing to invest her time in me. I would also like to thank my committee, Dr. Ahmad Rteil and Dr. Sumi Siddiqua for their feedback and suggestions. I would also like to express my gratitude to my fellow students within the UBC Bioreactor Technology Group and the laboratory technicians from the School of Engineering for their advice and assistance.   I would also like to acknowledge funding from the Natural Sciences and Engineering Research Council of Canada (NSERC) Strategic Project Grant (SPG), Canada Foundation for Innovation (CFI) John R. Evans Leaders Fund and the City of Kelowna whose investments into research and research infrastructure make work like this possible.  Finally, I owe my wife an enormous debt for her patience and all of her sacrifices that have allowed me to pursue graduate studies. This would have been impossible without her.    xiv Dedication (MAKE THIS WHITE AFTER TO MAKE IT DISAPPEAR)            To my wife   xv Chapter 1 Introduction   Background 1.1Few technological advances have had such a dramatic impact on human health and the environment as the collection and treatment of domestic and industrial wastewater. For example, the first modern sewers built by Joseph William Bazalgette in the City of London prevented tens of thousands of cholera deaths and helped transform the river Thames from a hazardous and vile open sewer into a safe and pleasant waterbody (Halliday, 2013).   The liquid entering a wastewater treatment plant (WWTP) has a high concentration of suspended and dissolved solids that must be removed before the treated wastewater can be released into the environment. There are a variety of physical and/or chemical unit processes that target specific constituents within the liquid stream. These include screens or bars to remove large objects that may damage pumps and other equipment; grit chambers that target rocks, gravel and other dense material; primary sedimentation where suspended organics are settled. Later processes include secondary treatment that targets dissolved and colloidal organics and often a tertiary treatment process to remove nutrients, residual organic matter, and/or other deleterious substances before the effluent is released to the environment (Metcalf & Eddy et al., 2002).   The organic material that is removed from the primary, secondary and tertiary treatments when collected and thickened is referred to as waste sludge. Waste sludge quickly becomes very odorous, is very putrescible, and can attract disease-carrying vectors therefore it must be dealt with (or stabilized) promptly. Over 7 million dry tonnes of waste sludge are produced in the United States (Viau & Peccia, 2008) while Canadians produce over 660,000 dry metric tons of waste sludge every year (Canadian Council of Ministers of the Environment, 2014).  Waste sludge is challenging and expensive material to deal with and contributes significantly to the environmental burden of treating wastewater. In fact, expenses related to sludge handling and disposal can easily exceed 50% of a typical WWTP operating costs (Taylor & Nowak, 2007) and are a significant component of a WWTP’s carbon footprint (Koh et al.,  1 2010). Unstabilized waste sludge can contain undesirable materials including very high levels of pathogenic microorganisms and trace amounts of heavy metals and micropollutants including a wide range of pharmaceuticals, personal care products, industrial chemicals and flame-retardants.   Waste sludge is a rich organic material that can be transformed into a valuable source of nutrients, energy and other resources if stabilized. Unfortunately, large amounts of waste sludge are currently not utilized effectively, and are often disposed of in landfills or incinerated instead of being utilized in a beneficial manner (Metcalf & Eddy et al., 2002).   There are various methods of stabilizing waste sludge to reduce the volumes of solids, recover energy and/or make it suitable for beneficial re-use in agriculture. These methods include various chemical, thermal and biological processes, each of which has significant benefits and drawbacks. Once stabilized, waste sludge that meets quality criteria are referred to as biosolids, which can be applied to land as a fertilizer or soil amendment for agricultural use or for land reclamation.    Motivation for Research 1.2This research is part of a larger NSERC Strategic Project that is being conducted in collaboration with the City of Kelowna, the British Columbia (BC) Ministry of Environment and Paradigm Environmental Technologies Inc. This project aims to develop a life cycle assessment based, multi-criteria decision making tool, which will allow the City of Kelowna to compare different biosolids disposal options to minimize the environmental burden while maximizing the opportunities for energy and resource recovery while still being cost effective. Various sub-projects have been conducted to explore various treatment options, enhancements to existing sludge pre-treatment methods (including thermal, microwave, high-pressure homogenization and/or sonication), co-digestion of different waste streams, along with exploring various methods of addressing some of the common issues of sludge treatment. Lessons learned from this project are not specific to Kelowna or the region and could likely be applied to other regions and municipalities throughout Canada and beyond.   2 1.2.1 City of Kelowna Wastewater Treatment Plant The City of Kelowna is located in the Okanagan Valley in the southern interior of BC. Kelowna has a population of approximately 117,000 people (City of Kelowna, 2013) and has a WWTP with a treatment capacity of up to 70 million litres per day. The Kelowna WWTP employs a combination of traditional treatment processes as well as a biological nutrient removal (BNR) system which removes compounds containing carbon, nitrogen and phosphorous which could otherwise lead to eutrophication (excess algal growth) to the body of water that receives the treated effluent from the facility.  Currently, the City of Kelowna produces over 17,000 tonnes of dewatered waste sludge annually that is trucked 46 km to a composing facility in Vernon, BC. This sludge is a combination of fermented primary sludge (FPS) and waste activated sludge which is dewatered to approximately 18% solids by using a centrifuge before shipment (City of Kelowna, 2012). The process flow diagram is shown in Figure 1.1.  The composting facility mixes the dewatered mixed sludge with other co-substrates including wood chips and incinerator ash to add bulk and micronutrients to the compost while decreasing the generation of odour during the process (Carpenter & Beecher, 1997). The composing facility uses extended aerated static composting for a period of approximately 55 days where the activity of aerobic bacteria elevates the temperature of the compost pile to a minimum temperature of 55⁰C for at least three days to kill the majority of pathogens that may be present (City of Kelowna, 2012). The resulting material is screened to remove deleterious materials and aged for about 60 days before it is sold as OgoGrow, a soil amendment for agricultural use.  3  Figure 1.1 – Simplified treatment flow diagram of Kelowna’s wastewater treatment facility  Although significant quantities of Ogogrow are sold every year, Ogogrow sales do not cover the cost of operating the facility, which places a large burden on local taxpayers. Additionally, large quantities are given away for free each year to free up room and the facility has received numerous complaints from nearby residents due to odour emanating from the facility (Rolke, 2013). Compounding these issues are ever more stringent government regulations that restrict the levels of metals, remaining pathogens and other potentially harmful material. We are just beginning to understand the impacts of extremely small concentrations of various domestic and industrial chemicals, pharmaceuticals and personal care products on humans and the environment. It is quite likely that many of these compounds of emerging concern will require treatment as their use becomes increasingly common and science increases our understanding of their potential for harm. Population growth and economic diversification will only further exasperate these problems so it is imperative that other solutions are explored.  4  1.2.2 Anaerobic Digestion One of the most effective methods of waste utilization that has been widely implemented is anaerobic digestion (AD). AD is a multi-step biochemical process that utilizes a diverse community of microorganisms to convert organic matter into valuable by-products including a methane-rich biogas and a nutrient-dense digestate (stabilized digester effluent) that is ideal for composing or land application. Through AD and composting, many municipalities have transformed biosolids from municipal and industrial wastewater treatment from a burden to taxpayers and the environment into a valuable source of energy and nutrients. Another benefit of AD is the near 50% reduction in the volume of the biosolids that greatly reduces the environmental impact and costs related to biosolids storage, handling and transportation (Droste, 1996).  The effects of AD are unclear from several viewpoints, particularly odour potential. High levels of corrosive and odourous gasses are generated during AD that can damage biogas handling and utilization equipment and can greatly effect nearby residents. Even without AD, wastewater treatment, sludge handling/transfer/treatment and composting facilities often face significant opposition from nearby residents because they can be quite odourous. Many wastewater treatment facilities go to great lengths to control and treat odour from various unit processes and often locate facilities a distance away from population centres with mixed results. Many odourous compounds have exceptionally low odour thresholds so even the smallest fugitive emissions can have a significant impact on the community. Ever expanding cities and communities often surround once distant facilities making odour issues a key concern for nearly every wastewater treatment facility.   Another barrier to wider adoption of AD is the amount of pathogens that remain in the digested sludge. Land application of the digested sludge is one of the most economical and beneficial uses of this waste material, however there are regulations that limit the amount of pathogenic microorganisms that may be present in the applied material. Longer sludge retention times (SRTs) and elevated operation temperature within the digester results in a decrease in pathogens and a greater removal of solid materials, however it necessitates larger  5 operating volumes and results in higher construction and operating costs. Anaerobic digesters with shorter SRTs are more economical, however due to higher organic loading rates, the AD is more prone to process upset and failure and the resulting biosolids may require further treatment to meet land application regulations (Boe, 2006). Determination of the optimum SRT attempts to balance cost and operational factors while achieving a reasonable amount of solid and pathogen destruction.  Another issue with AD is the solubilisation (release of materials from the solid phase into the liquid phase) of biologically bound nutrients when sludge from a BNR process is digested. Significant amounts of phosphorous (as orthophosphate) can be released during digestion and dewatering of the digested sludge (Pitman et al., 1991). The liquid from the dewatered sludge (centrate) is returned to the beginning of the WWTP for treatment. Excessive amounts of orthophosphate in the centrate can cause a feedback loop with ever-increasing amounts of orthophosphate accumulating within the facility, which can overwhelm the BNR system and may result in effluent orthophosphate levels that exceed discharge limits.   Odour when combined with high levels of pathogens that can remain in the digested sludge, and the release of nutrients from the solid phase to the liquid phase and other practical issues can limit the wider adoption of AD.   Objectives 1.3This research aims to explore and to address some of the barriers to the greater adoption of AD, namely: a) The production of odourous and corrosive volatile sulfur compounds  b) High levels of pathogens remaining after mesophilic AD c) The release of nutrients like orthophosphate from the solid to the liquid phase  While extensive research has been conducted on biosolids and odours, most studies have explored odour from dewatered biosolids by batch testing in small serum bottles, or odour levels in facilities that handle, store or process biosolids. Very little research has been done to identify and /or quantify the odourous compounds present in the biogas of continuously-fed  6 complete mixed AD, while even less has been done on techniques to reduce them. Sludge cake and serum bottle testing conducted by Higgins (2010), Novak (2010), and others, which have explored the impact of iron and aluminum concentrations present and added showed great promise and an opportunity for further study. This and other available literature suggests that by adding certain metal ions into the digester, a significant reduction in the production of one of the most odourous classes of compounds production may be possible.  The author of this thesis proposes a hypothesis that the addition of metal salts to municipal waste sludge before sludge digestion will reduce the production of highly odorous volatile sulfur compounds (VSCs) in AD biogas as existing literature has shown an inverse correlation between iron and aluminum concentrations in AD sludge and the production of VSCs from dewatered sludge cake. Furthermore, magnesium hydroxide has been used for many years to control the production of VSCs in sewers. The second hypothesis proposed is that the addition of metals salts will also improve the dewaterability of digested sludge as metal salts have the ability to change the zeta potential of the sludge flocs.  To validate or invalidate these hypotheses, this study explored the effects of adding two concentrations of three different metal salts to AD feed before digestion.     7  Thesis Outline 1.4This thesis is divided into five chapters. The above introduction chapter briefly explores wastewater treatment, introduces AD and states the motivations for this study. The second chapter explains wastewater treatment in more detail and some of the associated unit processes including sludge treatment, sources and mechanisms of odour/volatile sulfur compound production, issues associated with AD. The third chapter presents the experimental methodology and the materials and methods that were used to conduct this study. The fourth chapter presents the results of the work and compares them to existing literature. The fifth and final chapter of this thesis summarizes the results and provides some suggestions for future work.   8 Chapter 2 Literature Review 2    Chapter: Literature Review  Background 2.1Since the beginning of civilization, untreated human waste has been responsible for countless disease outbreaks and made many cities and urban waterways highly unpleasant. Modern wastewater treatment has saved countless lives and has prevented or even reversed eutrophication in many waterways that receive effluent from human activities.  Management of WWTP residuals has been a challenge since the very first treatment facilities. Early treatment facilities in many coastal cities like London, Manchester, Glasgow, New York, etc. first attempted to dispose of their waste sludge by discharging it untreated into harbours (which soon became very foul and odorous), then later by barges far out to sea; both of which had the potential to expose people to sewage-borne pathogens via swimming or the consumption of shellfish (Hudson, 1995; Swanson et al., 2004). Other cities would often spread cake from filter pressing plants onto soil to provide nutrients for crops (Hudson, 1995), while others would send their waste sludge to landfills. It was not until later that other methods of sludge disposal became more common.   One the most effective ways that was developed to deal with wastewater residuals is AD. AD is a biological process that uses a diverse community of microorganisms to break down organic material within the sludge matrix into a methane rich biogas in the absence of molecular oxygen. A few key benefits of AD include: • a near 50% reduction in solids that require disposal after digestion (Ndon & Dague, 1997); • several orders of magnitude reduction in the amount of pathogenic microorganisms remaining in the solid material after digestion; • production of valuable biogas which can be upgraded to biomethane (97% methane) and sold to outside utilities or utilized onsite for heating and/or electricity generation; • a far more nutrient-dense digestate; • significant cost savings due to reduced volumes of materials to be disposed and the energy recovered from the biogas  9 AD facilities first appeared in Europe in the 1950s and have now become quite common; however, they are not ubiquitous due to some challenges and limitations (Droste, 1996). Some of the key barriers to wider AD adoption include: • limitations to the rate at which microorganisms can break down organic material requiring long retention (holding) times and large digester volumes; • long retention times required (>20 days) to meet government regulations on pathogens in the residuals; • various recalcitrant industrial and domestic chemicals, pharmaceuticals and personal care products that may remain and be harmful in extremely small concentrations; • release of bound nutrients from the solid phase into the liquid phase; • corrosion of reinforced concrete vessels, damage to biogas processing & utilization equipment and odour issues from the odourous compounds in biogas.  There has been a significant amount of research conducted that has attempted to address some of these issues including pre-treatments to increase the rate of hydrolysis, elevated temperatures to kill pathogens, dual-stage digesters to optimize AD, etc. Unfortunately, many treatments that improve one metric often come at the expense of another.  Odour from WWTPs is one of the most challenging aspects of managing many wastewater facilities and typically is the largest sources of complaints from nearby residents. Various odourous compounds are produced throughout the wastewater collection system and treatment processes. Volatile sulfur-based compounds (VSCs) produced from the anaerobic degradation of organics, ammonia, aromatics, etc. all contribute to odour from WWTPs. High concentrations of VSCs produced during anaerobic decomposition combined with their extremely low odour thresholds and very negative hedonic values make VSCs one of the most important classes of compounds to target. One of the members of the VSC family – hydrogen sulfide (H2S) – has immediate and highly toxic properties to humans at concentrations of less than 100 parts per million (ppm), while extended H2S exposures can cause nausea, sleep disturbances and/or airway constriction at concentrations of less than 5 ppm. Therefore, the smallest fugitive emission from any part of an AD or biogas utilization  10 system can have a significant and potentially harmful effect on workers in the facility and the surrounding community.  While some research has been conducted on VSCs produced during anaerobic decomposition of wastewater residuals, most research has focused on odour from the residuals after digestion. Very little is known about the composition and concentrations of VSCs in the biogas from operating ADs and even less is known about means of reducing them. Studies have explored the effects of different metal ions on odour from dewatered cake; however, none has explored the effects of metals on an operating AD, which was the main motivation for this thesis. Furthermore, little work has been conducted exploring the effects of metal salts on pathogens in AD residuals and soluble nutrients so these also deserve further exploration.   Wastewater Treatment 2.2Wastewater typically contains high levels of suspended, dissolved and colloidal organic and inorganic materials, the composition of which vary greatly depending on the source. Domestic, industrial and agricultural wastewaters all tend to contain high levels of organics, which can deplete the available oxygen in the receiving water body as bacteria metabolize the materials. One key wastewater metric that helps to measure the potential impact of wastewater on a receiving water body is the biochemical oxygen demand (BOD), which assesses the amount of oxygen that would be required within the receiving water for aerobic microorganisms to degrade organic material present in the waste. Another key wastewater metric is chemical oxygen demand (COD), which is a test that indicates the mass of oxygen that would be required to mineralize all of the organic materials in the wastewater, including waste that is non-biodegradable. COD is often used to assess and compare wastewater treatment processes.  Some other problematic wastewater constituents include: nutrients which can stimulate the growth of algae in the receiving water which can lead to eutrophication, pathogenic microorganisms which can cause illness and spread disease, and various household and industrial chemicals may be harmful to humans and the environment if not properly treated.  11 The typical composition of untreated domestic wastewater from varying countries is shown in Table 2.1. High strength wastewater is commonly found in developing countries like Bangladesh where their per capita water consumption is quite low, while low strength wastewater is typically produced in developed countries like Canada where per capita water consumption is quite high.  Table 2.1 - Composition of untreated domestic wastewater Contaminant Unit Concentration of Contaminants in Wastewater Low (750 L/capita /day) Medium (460 L/ capita /day) High (240 L/ capita /day)  Total solids (TS) mg/L 390 720 1230 Dissolved solids (DS) mg/L 270 500 860 Suspended solids (SS) mg/L 120 210 400 Biochemical oxygen demand (BOD5) mg/L 110 190 350 Chemical oxygen demand mg/L 250 430 800 Nitrogen (total as N) mg/L 20 40 70 Phosphorus (total as P) mg/L 4 7 12 Organic phosphate (as P) mg/L 1 2 4 Inorganic phosphate (as P) mg/L 3 5 10 Chlorides mg/L 30 50 90 Volatile organic compounds (VOCs) μg/L < 100 100-400 > 400 Oil and grease mg/L 50 90 100 Sulfate mg/L 20 30 50 Total coliform bacteria No./100 ml 106 – 108 107 – 109 107 – 1010 Fecal coliform bacteria No./100 ml 103 – 105 104 – 106 105 – 108 Cryptosporidium oocysts No./100 ml 10-1 – 101 10-1 – 101 10-1 – 102 Giardia lamblia cysts No./100 ml 10-1 – 101 10-1 – 102 10-1 – 103 Adapted from Wastewater Engineering: Treatment and Reuse, 4th edition by Metcalf et al. Copyright 2003 McGraw-Hill Education. Used with permission of the publisher.    12 2.2.1 Wastewater Treatment Plants  WWTPs are designed to settle out solids and to remove conventional pollutants in the liquid stream making it safe for discharge back to the environment. Most wastewater systems employ a combination of treatments that target different wastewater constituents, the exact configuration of which varies greatly depending on the composition of the wastewater.   Domestic WWTPs typically employ a combination of: • preliminary treatments which targets rags, sticks, grit and grease that may cause damage to other treatment operations; • primary treatment which removes a significant portion of the suspended organics; • secondary treatment which targets colloidal and dissolved organics; • tertiary treatment utilizing biological, physical and chemical processes to remove nutrients, particulates, chemical contaminants etc.; • disinfection of the treated effluent as a final step before it is released back to the environment. A common WWTP configuration (which also employs AD as “sludge digester”) is shown in Figure 2.1.   Figure 2.1 - Common wastewater treatment configuration Image from the City of Saskatoon. Copyright 2014 the City of Saskatoon. Used with permission.  13 2.2.2 Preliminary Treatment Wastewater entering a WWTP often contains objects like sticks, branches, trash, rocks, grit etc. that must be removed before treatment. As the wastewater enters the facility, it typically first flows through a bar rack to remove large objects like logs, rags, and trash from the stream as these items could damage and/or cause blockages within the facility. The screens are usually mechanically cleaned and the debris is sent to a sanitary landfill for disposal.   After screening, the influent travels into a grit chamber where the water’s velocity, retention time and other parameters are carefully controlled so that dense materials like sand, silt, broken glass, coffee ground and other dense materials settle out while the lighter organic material remains suspended within the waste stream. It is important that these materials are removed before treatment as rocks and grit can abrade pumps, piping and damage other mechanical devices. Grit can also settle and accumulate in locations within the plant wherever there are quiescent flow conditions, leading to a reduction in hydraulic capacity and blockages (Metcalf & Eddy et al., 2002). The settled material from the grit chamber is collected and is sent to a sanitary landfill for disposal.  The level and concentration of WWTP influent varies greatly throughout the day and often follows a diurnal pattern depending on the community. Peak flow and waste concentrations typically occur earlier in the day as people awake, and then it drops significantly as people go to work. A second peak often occurs in the early evening as people return home and go about their activities. Therefore, many WWTPs employ equalization tanks to provide a near constant level of flow and dilute sudden spikes in waste concentrations entering later unit processes. Not only does this help to significantly improve the performance of the WWTP and prevent process upset from shock loads, it can often provide cost savings by allowing a smaller plant to be built as peak flows can be stored and treated slowly throughout the day and night (United States Environmental Protection Agency, 1979a).    14 2.2.3 Primary Treatment The majority of suspended organic matter remains in the wastewater after preliminary treatment, much of which can be removed by settling under gravity. The flow enters a primary settling basin where quiescent conditions allow the suspended organics to settle to the bottom. The settled material is high in pathogenic microorganisms, is highly putrescible and is comprised of mostly water (95-99%) so it is collected and thickened to become primary sludge (PS). During primary settling, any floating materials (such as oil and grease) are collected by surface skimmers and are removed for later treatment (Droste, 1996).   2.2.4 Secondary Treatment There is still a significant amount of dissolved and colloidal BOD remaining after primary treatment, thus further treatment is required. The primary objective of secondary treatment is to provide an optimum environment where microbial cells can utilize and remove the dissolved organics. Microbial cells accomplish this by using the organics as a carbon source for growth, reproduction and cellular functions by converting the organics into cellular materials and various gasses and other waste products (Metcalf & Eddy et al., 2002).   There are two main types of secondary (biological) treatment systems that include: attached growth where microorganisms grow on a fixed or solid moving surface that is exposed to the wastewater; and suspended growth where dead and alive microbial cells are thoroughly mixed in a solution of wastewater. There are a variety of processes that can accomplish this which include activated sludge, oxidation ponds, trickling filters, rotating biological contactors among others. The microorganisms in each of these processes tend to create clumps of cells and extracellular polymeric substances that have a density slightly greater than water; thus, they can be settled by gravity after biological treatment in a secondary clarifier or removed using processes like dissolved air flotation (DAF).   Activated Sludge 2.2.4.1Activated sludge is a two-part process where microorganisms are “activated” by aeration to form a suspension of wastewater and microbial cells called mixed liquor suspended solids (MLSS) in an aeration basin. As the treated wastewater and MLSS flow out of the aeration  15 basin, the effluent is sent to a clarifier where the microbial cells and other suspended materials are allowed to settle using type III (zone) settling (Droste, 1996). Much of this sludge containing MLSS is recycled back to the aeration basin to maintain constant biomass concentration, while the excess is removed as waste activated sludge (WAS).   Oxidation Ponds or Lagoons 2.2.4.2Oxidation ponds or lagoons are typically shallow pools that use a combination of sunlight, aerobic bacteria and algae in suspended aerobic conditions to remove organics from wastewater. Bacteria metabolize and/or assimilate the organic matter for growth and reproduction and release carbon dioxide and various other inorganic materials. Algae use sunlight, the released carbon dioxide and inorganic materials to perform photosynthesis and release oxygen to the water that is used by the bacteria. Occasionally, oxidation ponds are mechanically aerated to accelerate the organic removal rates allowing for significant reductions in hydraulic retention times and thus, the size of oxidation pond required.   One key benefit of oxidation ponds is lack of a secondary clarifier as most of the organic matter is either oxidized or settles to the bottom of the pond. The effluent from the pond typically has few organic materials other than algae remaining, which can be removed using filtration or a combination of chemical treatments and settling. One disadvantage of oxidation ponds is the gradual loss of hydraulic capacity over time due to an accumulation of sludge at the bottom of the pond. This sludge must be removed by dredging every few years and must be disposed of (Nathanson, 2014).   Tricking Filters 2.2.4.3Trickling filters are a type of an attached growth process where the wastewater is sprayed over a filter media by rotating sprayer arms where it trickles through the filter by gravity and is collected in the bottom. Trickling filters are often placed in series and/or parallel in various combinations for redundancy and to achieve higher levels of organic removals.  The filter media is typically made from crushed rock or plastic with a large amount of surface area where a layer of biologically active slime layer (or biomass) can develop. This layer  16 contains an array of aerobic and facultative organisms including bacteria, fungi, algae, protozoans, worms, insect larvae, snails etc. (Metcalf & Eddy et al., 2002). Since a tricking filter is an aerobic process, air is introduced into the filter by convection currents or is forced through the filter by blowers allowing the microorganisms to metabolize and/or assimilate the substances in wastewater. A portion of the filter effluent (0.5 to 3 times) is recycled to maintain a constant wetting rate, to increase the quality of the effluent after the filter and to dilute sudden spikes in wastewater strengths or toxic materials that may have deleterious effects on the biofilm (Toprak, 2004). As the biofilm grows, excess biofilm is sloughed off into the wastewater stream and wash through the filter. Thus, a secondary clarifier is typically employed after trickling filters to settle out the excess biofilm and remaining organic matter after treatment.    Rotating Biological Contractors 2.2.4.4Another form of attached growth bioreactors are rotating biological contractors that use media made of circular shaped corrugated plastic. A series of media are mounted onto a horizontal shaft that is partially submerged (~40%) in wastewater and is rotated slowly so that the attached biomass gets exposed to wastewater and the atmosphere (Grady et al., 2011). The rotation prevents the biofilm from drying out while providing sufficient contact time with both oxygen in the atmosphere and organic matter in the wastewater. Similar to the trickling filter, excess biomass is sloughed off the media into the effluent and is removed.  2.2.5 Tertiary Treatment After secondary treatment, many municipal WWTPs have added additional steps to further improve or polish the quality of the effluent before it is discharged to the environment. After secondary treatment, there often are large quantities of nutrients that can cause eutrophication if left untreated, along with fine particulate matter and domestic and/or industrial chemicals and their remnants that were not removed during treatment. These materials can be addressed by various unit processes including BNR systems, various forms of filtration, along with adsorption or advanced oxidation that target specific constituents within the effluent.    17  Biological Nutrient Removal 2.2.5.1A key objective of treating wastewater is to remove the extremely high level of nutrients that are present in wastewater and are harmful to the environment. These nutrients typically include organic and inorganic phosphorous as well as nitrogen containing compounds including ammonia (NH3), nitrite (NO2-) and nitrate (NO3-). Nitrogenous compounds are important biological molecules that greatly accelerate the rate of growth of plants and algae. Phosphorous is a limiting nutrient in many freshwater ecosystems. Therefore, controlling nutrients in plant effluent is very important to prevent eutrophication of the receiving water body (Metcalf & Eddy et al., 2002).   Many WWTPs now employ BNRs that use different combinations of aerobic, anoxic and anaerobic zones that enable various bacteria to uptake or utilize different nutrients. Each zone is differentiated by the terminal electron acceptor utilized; in aerobic zones oxygen is the terminal electron acceptor, in anoxic zones nitrate-N is utilized, while in anaerobic zones neither acceptor is present (Grady et al., 2011). These varying conditions cause bacteria known as phosphorous accumulating organisms (PAOs) to assimilate and accumulate the soluble phosphorous from the wastewater into their cells that are later removed as waste sludge. Nitrogenous compounds are removed by different microorganisms that undergo nitrification and denitrification which is a process that ultimately converts ammonia/nitrate/nitrite into nitrogen gas which is then released to the atmosphere as shown in Figure 2.2 (Makowska et al., 2013). BNR is a very effective way of controlling nutrients and have been widely adopted, particularly in newer WWTPs.    18  Figure 2.2 - Biological nitrification and denitrification  (Adapted from Małgorzata et al., 2013)  After treatment, BNR effluent is sent to another clarifier to remove the accumulated biomass from the BNR. The majority of the biomass is returned to the beginning of the BNR to provide an adequate microorganism concentration, while the excess biomass is thickened and is wasted.   Filtration 2.2.5.2Although filtration is nearly ubiquitous in drinking water treatment, it is not nearly as common in wastewater treatment systems however its use is slowly increasing (Metcalf & Eddy et al., 2002). Filtration is becoming a more common a step in some WWTPs as is effective in removing particulate BOD along with remaining suspended solids. This is often accomplished through either depth or surface filters where water pressure (or head) is used as the driving force to push the water through the filter.  Depth filters employ a variety of granular media including sand, crushed anthracite, garnet, and occasionally inert synthetic materials and use a variety of physical and chemical mechanisms to remove particulates as the effluent travels through the filter (Droste, 1996). Physical mechanisms include straining, whereby particles that are larger than the pore space are physically trapped between particles. Sedimentation is another physical mechanism where particles settle onto the filter, which is enhanced by biological growth, as microbial activity within the filter will reduce pore size and volume leading to higher removals.  19 Flocculation is another form where smaller particles agglomerate to a size where they can be removed by a variety of mechanisms (Metcalf & Eddy et al., 2002). Once particles have come into contact with the filter, there are several physical and chemical phenomenon that occur to keep them there which include chemical adsorption (bonding and/or chemical interaction between the particles and the filter media), along with physical adsorption (electrostatic electrokinetic, and/or van der Waals forces) (Metcalf & Eddy et al., 2002). Filter performance decreases as matter accumulates in the filter until a point where the head loss though the filter exceeds a predetermined limit and/or when the quality of the effluent is no longer acceptable. At this point depth filters are regenerated either by backwashing, or by scraping off the top of the filter exposing cleaner material.  Surface filters remove particles by mechanical sieving where the water is pushed through small openings in a thin filtering material. In wastewater treatment, this is often accomplished using cloth-medium surface filters which woven in random 3-dimensional patterns and which result in size openings of 10 to 30 µm or more. Cloth filters typically are attached to a rotating spindle and are mounted vertically in a partially filled filter tank. Unfiltered water is either driven from the centre spindle through the filter into the tank or from the tank into a central collection channel inside the filter. The filter rotates as it operates maximizing the effective area of the filter while allowing for simultaneous filtration and backwashing keeping the filter clean (Metcalf & Eddy et al., 2002).   Disinfection 2.2.5.3Raw wastewater can contain a wide range of pathogenic organisms including various bacteria, viruses, protozoa, and parasitic worms. Although there is a decrease in the overall number of pathogenic microorganisms throughout the wastewater treatment process, treated wastewater typically requires disinfection before it can be discharged. Therefore, most WWTPs employ disinfection as the final step before the treated effluent is released to the environment.  According to Tchobanoglous et al. (2002), three common methods are used to disinfect wastewater, which includes the addition of chlorine or ozone, or exposing the treated water to  20 ultraviolet radiation. Chlorine and ozone are both strong oxidants that damage cell walls, which can lead to cell lysis and death. Chlorine, ozone and their by-products also participate in various chemical reactions that disrupt cellular activity. Ultraviolet light disinfects by causing photochemical reactions that damage the DNA and RNA within the cells. Although the ultraviolet light does not directly kill the organisms, the genetic damage that it inflicts prevents them from reproducing, effectively inactivating the cells (Crittenden et al., 2012; Metcalf & Eddy et al., 2002). Ultraviolet light is particularly effective in treating protozoa like cryptosporidium as the oocysts that transmit them are highly susceptible to UV light, whereas they are mostly resistant to treatment by chlorine and ozone based disinfectants (Crittenden et al., 2012).   Waste Sludge Characteristics 2.3Ever increasing amounts of wastewater sludge from municipal, agricultural and industrial activities are produced each year as human activities and population grow. Waste sludge is produced during primary, secondary and tertiary treatments and varies greatly depending on the origin and type of treatment used. According to Tchobanoglous et al. (2003), waste sludge is a particularly challenging material to deal with due to the fact that it is these solids that give wastewater its offensive nature, the organics in sludge from biological treatments become very offensive rather quickly, while only a small fraction of the sludge is actually solids. The remaining fraction is primarily water which is removed though various dewatering processes including centrifugation, belt presses or thermal treatments. Sludge handling and disposal costs typically exceed 50% of a typical WWTPs operating cost (Taylor & Nowak, 2007) and are comprised of a significant faction of the carbon footprint of wastewater treatment (Koh et al., 2010).  2.3.1 Sludge Properties  Primary sludge (PS) typically is gray and slimy, has a rather offensive odour and contains large amounts of pathogenic microorganisms that have been concentrated during treatment (Metcalf & Eddy et al., 2002). PS contains large amounts of readily degradable organic material and can be concentrated fairly easily within the primary settling tanks (Metcalf & Eddy et al., 2002; Wang et al., 2007). For each litre of wastewater treated, 100-300 mg of PS  21 typically is produced based on a removal of 50-65% of the total suspended solids entering the primary settling tank (Wang et al., 2007). Secondary sludge or WAS generated from biological treatment is comprised mainly of microbial cells, extracellular polymeric substances (high-molecular weight compounds including various proteins, polysaccharides, lipids, etc.) and some inorganic material. Secondary sludge from trickling filters is generally inoffensive when fresh and decomposes more slowly than other sludges (Metcalf & Eddy et al., 2002), while sludge from RBCs tend to be very similar physically and chemically to trickling filter sludge (Shammas & Wang, 2007). Waste sludge from suspended growth BNR processes tend to have similar characteristics as WAS (Metcalf & Eddy et al., 2002).  2.3.2 Physical and Chemical Characteristics of Sludge It is important that the physical and chemical characteristics of wastewater sludge are fully understood before selecting a sludge treatment or disposal method as sludge properties can vary greatly depending upon the originating wastewater and the unit treatment processes used to generate the sludge. Table 2.2 describes physical and chemical properties of some of the common types of sludge.  Table 2.2 - Sludge properties Itemα Untreated Primary Sludge Anaerobically Digested Primary Sludge Activated Sludge pH 5.0-8.0 6.5-7.5 6.5-8.0 Total solids (TS) (% by wt.) 5-9 2-5 0.2-1.2 Volatile solids (VS, as percent of TS) (%) 60-80 30-60 59-88 Specific gravity (-) 1.02 1.03β 1.005 Alkalinity (mg/L as CaCO3) 500-1500 2500-3500 580-1100 Protein (as a percent of TS) (%) 20-30 1.6-3.0 32-41 Adapted from: αMetcalf & Eddy (2003), βEPA (1979)     22  Physical Properties of Sludge 2.3.2.1The physical properties of sludge can have a significant impact on sludge handling, treatment and disposal operations. The difficultly by which sludge can be handled, processed and dewatered can be influenced by sludge rheology, particle size, and surface charges among others.   Sludge Dewaterability 2.3.2.2Handling and disposal of dewatered sludge is very costly, therefore any change in the dewaterability of the digester effluent can result in significant costs or cost savings to the municipality. Since waste sludge is primarily water, removing excess water is particularly important as it greatly reduces the volume of material that needs to be dealt with, makes handling significantly less expensive, and slightly delays the onset of putrefaction (Metcalf & Eddy et al., 2002). Sludge contains several forms of water that can be removed with increasing difficulty. Vesilind and Martel (1989), Vesilind and Hsu (1997), Yin et al. (2004) defined four different types of water within sludge, which include:  • Free or bulk water which is water that is not attached to any solids and can be removed via gravitational settling;  • Interstitial water which is water that is trapped within a sludge floc or within a cell that can be released if the floc is disrupted or if the cell is damaged and can be removed via mechanical means; • Vicinal water is water that is held on the particle surface by the “molecular structure of the water molecules” and cannot be removed by mechanical means;  • Water of hydration is water molecules that are chemically bound to and within the molecules of the particles themselves and can only be removed by thermal and/or chemical means.   Capillary Suction Time 2.3.2.3The capillary suction time (CST) test has been cited extensively in literature as an indirect means of measuring dewaterability; however it does not quantify any particular physical parameter of sludge (Seyssiecq et al., 2003). Kavanagh (1980), Vesilind (1988), Chen et al. (1996), and others have found a correlation between CST and sludge dewaterability; sludges  23 that exhibit higher CSTs generally tend to be more difficult to dewater. Although CST does not directly measure a physical sludge property, the results of CST can be used to indirectly compare the dewaterability of different sludges and treatment methods.   Zeta Potential 2.3.2.4Another indirect means of measuring dewaterability is to measure the surface charges of the sludge flocs as the ability of particles to attract or repel each other is directly related to surface charges of the particles. This surface charge potential is often referred to as zeta potential. Particles with a zeta potential greater than ±30 mV will repel each other highly, while it becomes easier for particles with zeta potentials close to zero to get close to one another (Malvern Instruments Ltd., 2012). If the sludge flocs repel each other, the tendency of the flocs to agglomerate is reduced which will require more energy and effort to get them to come together as the sludge is dewatered. Zeta potential is highly variable depending on temperature, pH, conductivity and concentration of the sample (Malvern Instruments Ltd., 2012).   Sludge Rheology 2.3.2.5Rheology is a tool for the characterisation of the hydrodynamic properties of sludge suspensions and can be used to optimize various sludge processes including sludge transportation, utilization, dewatering, drying, and disposal (Seyssiecq et al., 2003). Rheological properties of sludge can vary greatly between different sludge types and sources (Travnicek et al., 2013), while wastewater sludge suspensions have been shown to behave as non-Newtonian fluids at higher solids concentrations (Mori et al., 2006; Seyssiecq et al., 2003). Metcalf & Eddy (2002) found that wastewater sludge behaves much like a Bingham plastic as the relationship between shear stress and flow is only linear after flow begins and varies greatly from sludge to sludge. Determining rheological properties like dynamic viscosity can be helpful in process design and selection as the behaviour of and fluid motion of sludge has implications on operational parameters throughout sludge handling and treatment (Pérez-Elvira et al., 2011).    24 2.3.3 Chemical Features of Sludge  The chemical properties of sludge vary greatly depending on the constituents of the raw wastewater and the unit processes wasting. Sludge pH, COD, alkalinity and volatile fatty acids (VFA) are often of particular interest for sludge treatment process selection and control. Additional parameters including heavy metals and other inhibitory compounds like toxic organics or sulfide may also affect what treatment process is optimal to meet performance objectives and regulatory requirements.   Another key sludge characteristic is the level of nitrogen and phosphorous contained within the sludge, as both are important nutrients for plant growth. Domestic wastewater sludge contains high levels of nutrients as shown in Table 2.3. Due to these nutrients and high levels of organic matter, wastewater sludge can often be diverted for beneficial use in agriculture after stabilization. Table 2.3 - Sludge nutrient levels Item Untreated Primary Sludge Anaerobically Digested Primary Sludge Untreated Activated Sludge Nitrogen (N, % of TS) 1.5-4 1.6-3 2.4-5.0 Phosphorous (P2O5, % of TS) 0.8-2.8 1.5-4.0 2.8-11.0 Adapted from Metcalf & Eddy (2003)  Although wastewater treatment has greatly improved the quality of many surface waters, wastewater treatment has created a significant solid waste problem as treatment generates staggering quantities of solid waste material. Arguably, the beneficial use and/or disposal of wastewater residuals are one of the most challenging environmental issues facing scientists and engineers today.   2.3.4 Disposal Historically, large quantities of waste sludge were discharged into the ocean, sent to landfills, or were incinerated, however these practices are becoming less and less common (United States Environmental Protection Agency, 1993).  The United States enacted a ban on the ocean disposal of waste sludge in 1991 (United States Environmental Protection Agency, 1993); today few countries use ocean disposal as a means of sludge disposal (International  25 Maritime Organization, 2014). Landfilling of waste sludge has been becoming increasingly more difficult as completion for limited landfill space, poor economics, the potential for groundwater contamination, and increasingly stringent government targets and regulations have discouraged this practice. Incineration is a common disposal method in many countries as it is nearly unaffected by variable sludge characteristics and can achieve significant reductions in the volume of solid material for disposal as nearly all organic matter is mineralized into carbon dioxide and water vapour. Incineration is not a perfect solution however, as it has high capital and operating costs, can generate a hazardous solid waste disposal stream (depending on the levels of pollutants in the remaining ash), as well as it can cause harmful health and environmental effects from particulates and other volatilised materials (Metcalf & Eddy et al., 2002).   Stabilization 2.4There are three primary goals of stabilization which include the reduction of pathogens, the elimination of offensive odours, and the inhibition/reduction/elimination of the potential for putrefaction (Metcalf & Eddy et al., 2002). These objectives are met by controlling the volatile or organic fraction of the sludge so that either few organics remain in the stabilized residuals, or the remainder is unsuitable for microbial activity (Metcalf & Eddy et al., 2002; United States Environmental Protection Agency, 2000). Currently, sludge stabilization is practiced at nearly all WWTPs as it accomplishes the aforementioned objectives, and often provides opportunities for volume reductions, improvements in dewaterability, and the opportunity to recover energy and/or nutrients from this otherwise waste material (Metcalf & Eddy et al., 2002).  Stabilized waste sludge that meets government limits on pathogens, organic removals, heavy metals and other deleterious substances are commonly referred to as biosolids. Biosolids can be land applied for a variety of uses including site remediation, agricultural and/or domestic use as a fertilizer or soil amendment. Various government agencies including the United States Environmental Protection Agency, the BC Ministry of the Environment and others classify biosolids as either Class A or Class B through Code of Federal Regulations Part 503, Standards for the Use and Disposal of Sewage Sludge, or the British Columbia  26 Environmental Management Act and Public Health Act Organic Matter Recycling Regulation (OMRR) respectively. These regulations prescribe Class A and Class B standards, and attach limits on the use of these materials. A table outlining OMRR regulations is included as Appendix A.   2.4.1 Methods of Stabilization Many physical, chemical, and biological methods of sludge stabilization have been established that achieve varying results. Some of the common methods include alkaline stabilization, composting, aerobic digestion and anaerobic digestion.   Alkaline Stabilization 2.4.1.1One of the most common chemical methods of stabilization is through the addition of alkaline chemicals which include hydrated lime (calcium hydroxide), quicklime (calcium oxide), fly ash, lime and cement kiln dust, and carbide lime (United States Environmental Protection Agency, 2000). During alkaline stabilization, sufficient lime is added to the waste sludge either before or after dewatering to raise the sludge pH to 12 or greater. This creates an unfavourable environment for microorganisms and begins various chemical reactions that alter the chemical composition of the sludge. The sludge pH will decrease somewhat after application as lime is consumed by various chemical reactions. Therefore, excess lime is required to keep the pH at elevated levels that will prevent subsequent biological activity. If quicklime or another compound that generates substantial heat during hydration is used, a significant increase in sludge temperature can be achieved resulting in even higher reductions in the levels of pathogenic microorganisms in the residual sludge (Metcalf & Eddy et al., 2002; United States Environmental Protection Agency, 2000).   Aerobic Digestion 2.4.1.2Aerobic sludge digestion (ASD) is a biological process that is an alternative method of sludge stabilization. ASD uses continued aeration of waste sludge without the addition of any new substrates for the biomass to consume. Once any remaining substrates have been depleted, the microbial cells within the biomass enter a phase of endogenous respiration and can only obtain energy to continue cellular activity by consuming their own protoplasm. This  27 converts 75-80% of the cellular matter into carbon dioxide, water, and ammonia greatly reducing the amount of sludge for final disposal (Metcalf & Eddy et al., 2002).   ASD can achieve high levels of organic removals that are similar to AD; however, it has several challenges that have limited its broader use. ASD requires the biomass to digest itself; thus, live biomass is required. Furthermore, ASD can only be used to stabilize PS if PS is mixed with waste activated, tricking filter and/or extended aeration sludges that contain much higher microorganism concentrations than PS. Furthermore, ASD is energy intensive as it requires constant mixing and air input and no energy can be recovered. ASD also often requires long hydraulic/sludge retention times and/or elevated temperatures to achieve sufficient organic and pathogen removals (Metcalf & Eddy et al., 2002).   Anaerobic Digestion 2.4.1.3Anaerobic digestion (AD) is one of the oldest and most well established methods of waste sludge stabilization. AD is a multi-step biological process that utilizes a diverse community of microorganisms to convert organic matter into valuable by-products including a methane-rich biogas (60-70% by volume) and a nutrient-dense digestate that is ideal for composting or land application. AD is an effective method of transforming waste sludge from municipal and industrial wastewater treatment from a burden to taxpayers and the environment into a valuable source of energy and nutrients.  Upwards of 70% of mixed PS and WAS is degradable. During AD, up to 80% of the degradable portion is decomposed reducing the total solids (TS) by about 50% (Ndon & Dague, 1997). This greatly reduces sludge handling and disposal costs and their associated environmental impacts (Taylor & Nowak, 2007). The various steps of AD are shown in Figure 2.3.    28  Figure 2.3 – Stages of anaerobic digestion (adapted from Davis & Cornwell, 2008)  AD begins with the hydrolysis of large suspended organic materials within the sludge matrix into smaller soluble organic materials that can be metabolized. This is accomplished by fermentative microorganisms that secrete extracellular enzymes to breakdown lipids, proteins and nucleic acids, into various sugars, amino acids and long-chain fatty acids (Droste, 1996). These molecules may be transformed further or be transported across cell walls to be used for cellular growth, regeneration and/or metabolism. Hydrolysis is widely considered to be the rate limiting step of AD (Droste, 1996); therefore many AD process improvements have explored using various waste pre-treatments to accelerate this step of the process.  The second step in AD is acetogenesis. Some bacteria have a fermentative metabolic process where they can utilize some of the products of hydrolysis for energy and to produce simpler organics including various VFAs, alcohols, aldehydes, carbon dioxide, hydrogen  and ammonia (van Haandel & Lubbe, 2007).   29 The final step in AD is methanogenesis, which is a process where a group of microorganisms from the domain archaea metabolize the products of acetogenesis. There are two groups of methanogenic bacteria that perform this function via different pathways; hydrogenotrophic methanogens and acetotrophic methanogens (or acetoclastic methanogens). Hydrogenotrophic methanogens use hydrogen as electron donor to reduce carbon dioxide into methane and water as shown in Equation 2.1. Acetotrophic methanogens transform acetate into methane and carbon dioxide by splitting acetate into methane and carbon dioxide as shown in Equation 2.2 (Droste, 1996). Equation 1 – Methane production using hydrogen 𝐶𝐶𝑂𝑂2  +  4𝐻𝐻2  →  𝐶𝐶𝐻𝐻4  +  2𝐻𝐻2𝑂𝑂   Eq. (2.1) Equation 2 – Methane production using acetate 𝐶𝐶𝐻𝐻3𝐶𝐶𝑂𝑂𝑂𝑂−  +  𝐻𝐻+  →  𝐶𝐶𝐻𝐻4  +  𝐶𝐶𝑂𝑂2    Eq. (2.2)  This gas mixture produced by AD is a valuable by-product as it is typically 60~70% methane and 30-40% carbon dioxide by volume and contains trace amounts of other gasses including nitrogen, hydrogen sulfide and water vapour. In many cases, sufficient energy is produced to meet the energy needs of the treatment facility offsetting some of the facility operating costs (Metcalf & Eddy et al., 2002).   Influencing Factors in Anaerobic Digestion 2.4.1.3.1There are a variety of physical and chemical factors that affect the performance and stability of operating ADs. These include digestion temperature, pH, levels of VFA and alkalinity, sludge retention times, organic loading rates, the substrate being digested and the nutrients contained within it, along with the presence of inhibitory compounds. Successful AD requires a delicate balance between acetogenesis and methanogenesis, both of which can be impacted by these factors.    30  Digestion Temperature 2.4.1.3.2Of the AD operating parameters, temperature is one of the most important as it greatly affects the rate of substrate metabolism and the physiochemical properties of substances within the digester that in turn affect the microbial population dynamics within the digester. Elevated temperatures can greatly increase the rate of biological activity and substrate utilization in the reactor by increasing the solubility of hydrolyzed organics and by accelerating chemical and biological reactions (Appels et al., 2008). Increased temperatures also help to reduce levels of pathogenic microorganisms and often produce Class A biosolids.   Although elevated temperatures can accelerate biological activity within the digester, it can cause AD operation issues as levels of inhibitory substances like free ammonia tend to increase at higher temperatures, while increasing temperatures enhance the dissociation of VFA in solution (pKa) increasing the probability of inhibition (Appels et al., 2008). Droste (1996), Appels et al. (2008) and others state that temperature stability is one of the most important AD stability parameters and temperature changes should be limited to no more than 0.6ºC/day.  Anaerobic digesters typically operate with three temperature ranges; thermophilic, mesophilic and psychrophilic. Thermophilic ADs typically operate at 50-60ºC and usually require significant energy inputs to keep the temperature elevated. The key benefit of thermophilic digestion is the enhanced rates of pathogen and solids destruction that allows smaller digesters to be built. Mesophilic ADs operate between 30-40ºC, and are very commonly used as they strike a balance between digester volumes and the associated energy inputs required. Psychrophilic ADs operate at temperatures below 15-20ºC and are uncommon (Droste, 1996). At low temperatures the rates of pathogen destruction, biological activity, and gas production are quite low requiring long holding times and large tank volumes (Droste, 1996).    31  Digester pH, Volatile Fatty Acids and Alkalinity 2.4.1.3.3Another key AD operation parameter is pH, which is directly influenced by VFA and alkalinity generated within the digester. Different groups of microorganisms within the digester have different optimum pH values; however an overall pH value near 7 is optimal (Droste, 1996). Methanogens are extremely sensitive to pH and prefer a pH between 6.5 and 7.2, while fermentative microorganisms are less sensitive to pH and will operate between pH values of 4.0 and 8.5 (Hwang et al., 2004).    Digester pH is primarily determined by the levels of VFA and alkalinity. VFA levels greatly affect the pH of the digester, as the production of VFA tends to decrease the pH. Fermentative microorganisms predominately produce acetic and butyric acids at lower pH values, while at a pH of 8, this shifts to acetic acid and propionic acid (Boe, 2006). VFA production usually is balanced by VFA consumption that produces various alkaline waste products including carbon dioxide, ammonia and bicarbonate. Digester pH is strongly influenced by the carbon dioxide concentration in the gas phase and the various forms of carbonate alkalinity in the liquid phase (Appels et al., 2008) which tends to increase digester pH and act as a buffer against sudden pH changes.  ADs can operate at a wide range of VFA concentrations provided that there is sufficient alkalinity. Droste (1996) reports that VFA levels from less than 100 mg/L to over 5,000 mg/L are acceptable if pH control is practiced, while Appels et al. (2008) suggests that a molar ratio of 1.4:1 of bicarbonate alkalinity to VFA should be maintained for stable operation.    Sludge and Hydraulic Retention Time 2.4.1.3.4The sludge retention time (SRT) is the amount of time that sludge remains within the digester and in a completely mixed reactor with no recycle, it equals to the hydraulic retention time (HRT) of the liquid within the digester. Digester SRT plays a vital role in the overall efficiency of the digester and is a key factor in determining the levels of pathogen destruction and solid removals by the digester. Longer SRTs provide more time for biochemical reactions to take place achieving higher levels of organic removals, however it requires larger  32 reactor volumes and greater energy expenditures to heat and mix volumes of materials. Shorter SRTs will decrease pathogen and solid removal efficiencies, however they allow for smaller reactor volumes that may provide significant savings from construction and operating costs.   When digested sludge is wasted, a portion of the microbial cells are removed from the digester, so it is important that the rate by which organisms are removed does not exceed the rate by which the organisms are able to reproduce. Methanogens are particularly slow growing and are susceptible to being washed-out of the digester at SRTs lower than five days, leading to VFA accumulation and process failure (Appels et al., 2008; Droste, 1996).   Organic Loading Rate 2.4.1.3.5The rate at which organic materials are introduced into the digester is referred to as the organic loading rate (OLR). High OLRs provide more substrate for fermentative microorganisms to convert into intermediate products including VFA. Digester stability will be maintained if VFA production is in balance with the ability of the acetoclastic methanogens to utilize them. The digester will become unstable if VFA begins to accumulate within the digester as it consumes the available alkalinity and creates a positive feedback loop of ever decreasing pH if not corrected. Decreasing pH and alkalinity within the digester results in greater methanogenic inhibition; this causes continued VFA accumulation and greater methanogenic inhibition and eventually leads to process failure.    Digester Performance 2.4.1.4The performance of ADs is determined by many factors that include the characteristics of the feed sludge along with digester temperature, SRT, sludge pretreatments, digester design, etc. Total and volatile solids removals, BOD/COD removal, and methane volumes are some of the common digester performance parameters of particular importance.  Solids in sludge are often measured as total, volatile and fixed solids. Total solids (TS) refer the material that remains after the sludge is completely dried at 105ºC that removes any free and/or trapped water from the sample. Volatile solids (VS) refer to the fraction that is lost  33 when a solids sample is burned at 550ºC for a minimum of one hour. VS encompass all degradable and non-biodegradable organics as well as any inorganic solids that breakdown at high temperatures. VS are generally used to estimate the biodegradability of organic samples (Metcalf & Eddy et al., 2002). Fixed solids refer to the remaining material after combustion and represent the non-organic and mineral fraction of the sample; however some organics will not burn and some inorganic solids volatilize at high temperatures (Metcalf & Eddy et al., 2002).  During digestion, up to 65.5% of VS within the waste sludge will be degraded and converted into biogas resulting in significant solids volume reductions (Water Environment Federation, 2008). Digester performance and “degree of stabilization is often expressed as percent reduction in volatile solids” when compared to untreated sludge (Appels et al., 2008). The ratio of TS vs. VS can be helpful when comparing different digestion and/or treatment scenarios to assess impacts on solid removals.  Two other important and related AD metrics are COD removal and the volume of methane produced by the digester. COD is a common measurement of waste strength and is often used to estimate the BOD of a waste sample. Although no oxygen is consumed during AD, COD is removed by the conversion of organics into methane. Methane is a valuable AD by-product that can be utilized for onsite heating, electrical generation, and/or upgraded and put into natural gas pipelines as a form of green energy. The maximum theoretical methane yield from a waste sample can be estimated by measuring COD as up to 0.35 m3 of methane can be produced per kilogram of COD removed (at standard temperature and pressure or STP of 0ºC and 1 atm) (Droste, 1996). This theoretical methane yield is rarely achieved since there are many factors that affect the conversion of COD to biogas. These include the inclusion of non-biological materials that are oxidized by COD testing as well as recalcitrant materials within the sludge matrix that may cause varying levels of inhibition (Droste, 1996). Another common reason for deviation from this value is due to the potential for gas leakage from the digester and/or gas collection systems.    34  Problems with Anaerobic Digestion 2.4.1.5Anaerobic digestion is an effective method of stabilization that allows for the recovery of energy and nutrients from waste organic materials. However, there are some challenges to AD that prevent the wider adoption of this technology which include the production of highly odourous compounds during digestion, levels of pathogens remaining in the digested materials, along with high levels of nutrients that can be released during digestion.    Odour from Wastewater Treatment 2.4.1.5.1Wastewater treatment facilities are often associated with odour as nearly every step of wastewater collection and treatment has the potential to produce odour. There are a very large number of compounds associated with wastewater treatment that cause odour which include many sulfur, nitrogen, and chlorine based compounds along with various acids, aldehydes, terpenes and others (Water Environment Federation, 2004). Odour detection thresholds of these compounds (the level at which these compounds can be detected by the human sense of smell) often vary by many orders of magnitude. Odour detection thresholds can range from sub-parts per billion (ppb) to hundreds of parts per million (ppm) depending on the compound.  Odours from wastewater treatment, sludge management and sludge handling facilities come from a combination of point and area sources as well as fugitive emissions. Point sources include emissions emanating from a specific area like a stack, vent or duct, etc., while area sources tend to be from the surfaces of liquids or solids like an aeration basin or a biosolids storage bin (Water Environment Federation, 2004). These sources of odour can be monitored and controlled fairly easily by employing foul air collection systems and/or enclosing the offending unit processes. Fugitive emissions are not confined to any one process or area and tend to be from valves, pipe connections, equipment leaks, malfunctions or from human error. Fugitive emissions tend to be much harder to manage as they can be difficult to locate and can vary greatly both spatially and temporally.   In Canada, each province is responsible for regulating odour from WWTPs so guidelines vary from province to province. In B.C., the only odourous compounds that are regulated by  35 B.C. Ambient Air Quality Objectives are total reduced sulfurs, which include gasses like hydrogen sulfide, methyl mercaptan, dimethyl sulfide, etc. as they are often produced in significant quantities during wastewater treatment and have low odour thresholds. B.C. Ambient Air Quality Objectives set a one hour Maximum Acceptable Levels (MAL) of total reduced sulfurs to 28 µg/m3 (20 ppb) and 6 µg/m3 (4 ppb) for one and twenty-four hour exposure periods respectively (B.C. Ministry of Environment, 2013).   Work Safe BC sets occupational health and safety exposure standards through section 5.48 of the B.C. Workers Compensation Act, Occupational Health and Safety Regulation which governs a far wider range of odourous compounds. Some of which are shown in Table 2.4 below.  Table 2.4 - Occupational health and safety exposure standards Compound Time Weighted Average (TWA) Short-term Exposure Limit (STEL) Ammonia 25 ppm 35 ppm Benzene 0.5 ppm 2.5 ppm n-Butanol 15 ppm 35 ppm Carbon disulfide 4 ppm 12 ppm Hydrogen sulfide n/a 10 ppm Methyl mercaptan 0.5 ppm n/a Sulfur dioxide 2 ppm 5 ppm Turpentine and selected monoterpenes 20 ppm n/a Adapted from WorkSafeBC (2014)    36  Volatile Sulfur Compounds  2.4.1.5.2Volatile sulfur compounds (VSCs) have been reported as the most problematic odourous compounds produced during AD due to their high concentrations in AD biogas and extremely low odour thresholds as shown in Table 2.5 (Sanin et al., 2010). Fugitive VSC emissions from WWTPs, even in trace amounts, can have a significant effect on the surrounding community and are often the primary source of complaints from nearby residents (Du & Parker, 2009).   Table 2.5 - Odour thresholds of common VSCs* Compound Formula Odor Detection Threshold (μg/L) Hydrogen Sulfide H2S 0.40 Dimethyl Sulfide (CH3)2S 9.00 Diethyl Disulfide (C2H5)2S 0.25 Dimethyl Disulfide (CH3)S-S(CH3) 1.00 Carbon Disulfide CS2 2.60 Methyl Mercaptan CH3SH 1.10 Ethyl Mercaptan CH3CH2SH 0.19 Propyl Mercaptan CH3CH2CH2SH 0.50 *Adapted from Sludge Engineering: The Treatment and Disposal of Wastewater by Sanin et al., 2010. Copyright 2010, DEStech Publications, Inc. Used with permission of the publisher.  VSCs include H2S and a number of other volatile organic sulfur compounds (VOSCs) as well. Of the components of VSCs, H2S is one of the most problematic VSCs due to its exceptionally low odour threshold and hazard to human health even at low concentrations as shown in Table 2.6. In addition to odour and toxic effects, VSCs can be damaging to plant infrastructure including concrete, piping, process and biogas utilization equipment (Du & Parker, 2009; Parker, 1951).      37 Hydrogen Sulfide Hydrogen sulfide (H2S) is the end product of the sulfur reduction cycle and is a colourless gas that has a strong offensive odour of rotten eggs and is produced by a variety of geological, chemical and biogenic processes. H2S is particularly hazardous as it can quickly anesthetize a victim’s sense of smell removing the perception of danger (Glass, 1990).  The density of H2S is slightly greater than that of air thus it can accumulate easily in low-lying and/or confined areas and can travel great distances along the ground in quiescent conditions (Simonton & Spears, 2007). Average annual ambient air concentrations of H2S in the atmosphere have been estimated to be between 0.15~0.3 μg/m3 (0.1~0.2 ppb) in several European cities, and as high as 55 μg/m3 (37 ppb) in communities downwind from to point sources (World Health Organization Regional Office for Europe, 2000).  Some geological processes include volcanoes and geothermal vents while common biogenic sources include swamps, marshes, sediments, bodies of stagnant water, etc. and are ubiquitous whenever organics undergo microbial degradation in anaerobic environments (Liang, 2008). Natural sources of H2S greatly exceed the amounts produced by human activities, the world health organization estimates that only 10% of the total global emissions of this compound are of anthropogenic origin (World Health Organization Regional Office for Europe, 2000).   38 Table 2.6 – Short-term health symptoms and effects of hydrogen sulfide exposure* Concentration (ppm) Symptom/Effect 0.00011-0.00033 Typical background concentrations. 0.01-1.5 Odor threshold. Above 30 ppm, odor described as sweet or sickeningly sweet. 2-5 Prolonged exposure may cause nausea, tearing of the eyes, headaches or loss of sleep. Airway problems (bronchial constriction) in some asthma patients. 20 Possible fatigue, loss of appetite, headache, irritability, poor memory, dizziness. 50-100 Slight conjunctivitis ("gas eye") and respiratory tract irritation after 1 hour. May cause digestive upset and loss of appetite. 100 Coughing, eye irritation, loss of smell after 2-15 minutes (olfactory fatigue). Altered breathing, drowsiness after 15-30 minutes. Throat irritation after 1 hour. Gradual increase in severity of symptoms over several hours. Death may occur after 48 hours. 100-150 Loss of smell (olfactory fatigue or paralysis). 200-300 Marked conjunctivitis and respiratory tract irritation after 1 hour. Pulmonary edema may occur from prolonged exposure. 500-700 Staggering, collapse in 5 minutes. Serious damage to the eyes in 30 minutes. Death after 30-60 minutes. 700-1000 Rapid unconsciousness, "knockdown" or immediate collapse within 1 to 2 breaths, breathing stops, death within minutes. 1000 Nearly instant death. *Public domain, adapted from the U.S. Department of Labour, Occupational Safety & Health Administration.     39 In solutions, the dissociation of H2S into sulfide ions varies directly with pH as shown in Figure 2.4. At a pH of 9 or greater, nearly all H2S will be contained within the solution as HS- ions, while at a pH of 5 or less, nearly all of sulfide will be in the H2S form. At a neutral pH, both are present in equal concentrations. Volatilization from the aqueous phase depends on Henry’s law and varies with temperature and concentration of both the liquid and gas phases (Droste, 1996).   Figure 2.4 - Dissociation of sulfide between pH 4 & 10  Other Volatile Organic Sulfur Compounds Several other odourous VOSCs can be produced in during wastewater treatment. These compounds vary in their structure and can have varying physical, chemical and hedonic values. The physical properties of several common VSCs are shown in Table 2.7.   40 Table 2.7 - Common odourous volatile sulfur compounds  Hydrogen Sulfide Dimethyl Sulfide Diethyl Disulfide Dimethyl Disulfide Carbon Disulfide Methyl Mercaptan Ethyl Mercaptan Chemical Formula H2S (CH3)2S (C2H5)2S (CH3)S-S(CH3) CS2 CH3SH CH3CH2SH Molecular Weight 34.081 62.13 90.19 94.199 76.14 48.108 62.14 Boiling Point (⁰C) -60.33 37.3 92.1 109.72 46 5.95 35.1 Melting Point (⁰C) -85.49 -98.24 -103.9 -84.67 -112.1 -123 -147.8 Solubility in Water 3,980 mg/L at (20⁰C) 2.2 x 104 mg/L at 25⁰C 3,130 mg/L at 25⁰C Insoluble 2,160 mg/L at 25⁰C 15,400 mg/L at 25⁰C 15,603 mg/L at 25⁰C Gas Colour Colourless Colourless Colourless Colourless/yellow Colourless Colourless Colourless Vapour Pressure 1.56 x 104 mm Hg at 20⁰C 502 mm Hg at 25⁰C 60.2 mm Hg at 25 ⁰C 28.7 mm Hg at 25⁰C 359 mm Hg at 25⁰C 1,510 mm Hg at 25⁰C 529 mm Hg at 25⁰C Adapted from U.S. National Library of Medicine Toxicology Data Network41   Production of Volatile Sulfur Compounds 2.5Municipal wastewater is rich in both inorganic and organic sulfur. Sulfur is an important element for many biological processes and is present in wastewater in varying amounts. Many biological compounds including proteins, amino acids and their metabolites contain sulfur, while inorganic sulfur can come from various sources including personal care products, pharmaceuticals and sometimes freshwater water itself (Du, 2010; United States Environmental Protection Agency, 2012). From the point of entry into a sewer system and throughout the wastewater treatment and sludge handling processes, there are a range of biological processes that work to transform biologically available sulfur into various intermediate compounds as part of the sulfur cycle shown in Figure 2.5.   Figure 2.5 - Sulfur cycle Modified after Water Environment Federation (2007).    42  Inorganic sulfur 2.5.1.1Sulfate (SO42-) is an common precursor to the production of VSCs and plays a significant part of the sulfur cycle. In 1978, Oremland & Taylor demonstrated that sulfate reducing bacteria (SRB) produce H2S from the reduction of various forms of inorganic sulfur like sulfate. Since then, various researchers have identified one particular strict anaerobe SBR desulfovibrio as being responsible for the majority of the reduction of sulfate to sulfide (H2S) as shown in Equation 2.3 (Water Environment Federation, 2007). Equation 3 - Reduction of sulfate to sulfide 𝑆𝑆𝑂𝑂42– + 2𝐶𝐶 + 2𝐻𝐻2𝑂𝑂 → 2𝐻𝐻𝐶𝐶𝑂𝑂3– + 𝐻𝐻2𝑆𝑆   Eq. (2.3)   Organic Sulfur 2.5.1.2Existing literature has shown that many odourous compounds are produced when amino acids contained in bioavailable proteins are degraded (Adams et al., 2003; Y Chen et al., 2005; Higgins et al., 2006). Since biosolids can contain up to 50% protein, organically bound sulfur plays a significant role in the generation of VSCs (Higgins et al., 2006).   Production of Hydrogen Sulfide and Methyl Mercaptan  2.5.1.3Cysteine and methionine are two common amino acids that have been reported to be present within many proteins extracted from both WAS and anaerobically digested sludge (Dignac et al., 1998; Higgins et al., 2004; Higgins & Novak, 1997; Morgan et al., 1991). The production of H2S and methyl mercaptan (MM) in anaerobic environments from the decomposition of cysteine and methionine has been well established (Persson, 1992; Persson et al., 1990; Yoshimura et al., 2000). Higgins et al. (2006) proposed that bioavailable proteins are broken down by the enzyme protease into simpler polypeptides that are further lysed into peptides by peptidase to form free amino acids including cysteine and methionine. The enzyme cysteine lyase catalyzes the formation of H2S from cysteine while methionine lyase catalyzes the formation of MM from methionine (Higgins et al., 2006).     43  Creation and Destruction of Other VOSCs 2.5.1.4Once H2S and MM are produced, various mechanisms can transform H2S and MM into other VOSCs. Several species of anaerobic bacteria isolated from freshwater sediments, soil and water samples and from operating ADs have been shown to add a methyl group to H2S and to MM to form MM and dimethyl sulfide (DMS) respectively ( Higgins et al., 2006; Smet & Van Langenhove, 1998). Bak et al. (1992) and Higgins et al. (2006) proposed that H2S forms DMS through an intermediate MM reaction and “the source of the methyl groups is often methoxylated aromatic compounds” which are indicated in Equation 2.4 and Equation 2.5 as R (Higgins et al. 2006). Various humic acids within biosolids can also be a source of methyl groups for this reaction and may play a significant role in the production of VOSCs from biosolids (Higgins et al., 2006). Equation 4 – Methylation of hydrogen sulfide to form methyl mercaptan 𝑅𝑅–𝑂𝑂 –  𝐶𝐶𝐻𝐻3 +  𝐻𝐻2𝑆𝑆 → 𝑅𝑅–𝑂𝑂𝐻𝐻 + 𝐶𝐶𝐻𝐻3𝑆𝑆𝐻𝐻  Eq. (2.4) Equation 5 - Methylation of methyl mercaptan to form dimethyl sulfide 𝑅𝑅–𝑂𝑂 –  𝐶𝐶𝐻𝐻3𝑆𝑆𝐻𝐻 → 𝑅𝑅 − 𝑂𝑂𝐻𝐻 + 𝐶𝐶𝐻𝐻3𝑆𝑆𝐶𝐶𝐻𝐻3  Eq. (2.5)  One other common VOSC produced during AD of municipal biosolids is dimethyl disulfide (DMDS). Several studies on MM producing cultures in anaerobic conditions were not able to produce DMDS and no biological pathway for DMDS has been reported (Higgins et al., 2006). Persson et al. (1990) suggested that transformation of MM into DMDS was the result of the oxidation of MM, while Chin and Lindsay (1994) observed that the formation of DMDS from MM only occurred in the presence of oxygen. Chin and Lindsay (1994) theorize that the reaction was catalysed by metals and other constituents within the biosolids as they improved the transformation. Equation 2.6 shows the process ( Higgins et al., 2006). Equation 6 – Oxidation of methyl mercaptan to form dimethyl disulfide 𝐶𝐶𝐻𝐻3𝑆𝑆𝐶𝐶𝐻𝐻 +  𝐶𝐶𝐻𝐻3𝑆𝑆𝐻𝐻 + 0.5𝑂𝑂2  →  𝐻𝐻2𝑂𝑂 + 𝐶𝐶𝐻𝐻3𝑆𝑆– 𝑆𝑆𝐶𝐶𝐻𝐻3  Eq. (2.6)   Destruction of VOSCs 2.5.1.5Numerous studies have explored the fate of VOSCs in biosolid storage areas and batch digestion as H2S and VOSC concentrations change rapidly within the first hours of storage/digestion and continue changing for many days afterwards. Methanogenic bacteria  44 have been shown to degrade or demethylate VOSCs into intermediate products and H2S under anaerobic conditions, particularly in low sulfate conditions (Chen et al., 2005; Du & Parker, 2009; Leerdam et al., 2008; Zinder & Brock, 1978). Equation 2.7 shows the stoichiometry of DMS degradation (Lomans et al., 1999). Numerous batch and serum bottle tests by Du & Parker (2009b) and Leerdam et al. (2008) and others have confirmed these findings by demonstrating that MM, DMS and DMDS concentrations do not change if the methanogens are inhibited. These reactions help to explain the relatively low levels of VOSCs found in AD biogas. Equation 7 - Demethylation of dimethyl disulfide 𝐶𝐶𝐻𝐻3𝑆𝑆𝐶𝐶𝐻𝐻3 + 𝐻𝐻2𝑂𝑂 → 0.5𝐶𝐶𝑂𝑂2 + 1.5𝐶𝐶𝐻𝐻4 + 𝐻𝐻2𝑆𝑆  Eq. (2.7)   Pathogenic Microorganisms 2.6Land application or composting is often a cost-effective and environmentally responsible way of disposing of stabilized biosolids. Although there is a decrease in the overall number of pathogenic microorganisms throughout the WWTP, pathogens are concentrated into the sludge and without stabilization they often far exceed government regulations for land application. Pathogens are a major concern to regulatory agencies and residents who live nearby. There is significant concern and some evidence that pathogens present in biosolids may become aerosolized during high-wind events, or may contaminate soils and surface waters from run-off or from extreme precipitation events (Viau & Peccia, 2008).  Pathogens that may be present in biosolids include various bacteria (Salmonella, Shigella, Escherichia coli, etc.), viruses (hepatitis, poliovirus, Coxsackievirus, etc.), protozoa (Cryptosporidium, Giardia, Toxoplasma, etc.) and parasitic worms (Ascaris lumbricoides, Ascaris suum, Necator americanus, etc.) which may cause serious illness and/or mortality (Gerba and Smith, 2005). Most jurisdictions set limits on the maximum number pathogens in biosolids that may be applied to land; The United States Environmental Protection Agency (EPA) sets limits on the levels of pathogens via Rule 503 which outlines the maximum levels of deleterious substances including fecal coliforms that may be present in land applied biosolids. Most Canadian Provinces set similar limits through regulations like B.C. Organic Matter Recycling Regulation (OMRR). Both sets of regulations require biosolids to have less  45 than 1,000 Most Probable Number (MPN) of fecal coliforms per gram of total solids (dry weight basis) to be considered Class A, while Class B biosolids must have less than 2,000,000 MPN of fecal coliforms per gram of total solids (dry weight basis) (EPA, 1993; BC OMRR 2002). Class A biosolids are the most desirable and have the least restrictions as to their use. Class B have increased restrictions as to their placement whereas biosolids that do not meet Class B standards may not be land applied until they are within these limits.  AD is one of the most commonly used methods of stabilization as not only does it greatly reduce the volume of solids requiring disposal and increase nutrient density of the digestate, it significantly reduces the biological activity and pathogens in the biosolids (Chen et al., 2005; Smith et al., 2005). Numerous studies have demonstrated several orders of magnitude reductions in the levels of pathogens in digested sludge. Watanabe et al. (1997) found fecal coliforms in undigested primary sludge ranged from 105 MPN of fecal coliforms per gram of total solids which decreased to 102 ~104 MPN after mesophilic digestion and to 100 (1 MPN) during thermophilic digestion. AD at thermophilic conditions is an effective means of reaching MPN levels suitable for Class A biosolids (Oropeza et al., 1996).    Nutrients 2.7When sludge from a BNR is digested anaerobically, a significant fraction of the phosphorus can be released from the solid phase back into the liquid phase that can cause significant operational issues for WWTPs that combine BNR and AD systems. After digestion, solids and liquids are separated with the solids being sent for composting/disposal while the liquid fraction is returned to the beginning of the WWTP. If an excessive amount of phosphorous is released back into the liquid phase, it will begin to accumulate within the WWTP and can lead to nutrient levels that can overwhelm the BNR leading to effluent discharges that exceed discharge limits. This is one of the key reasons why few WWTPs combine BNR with AD systems (Metcalf & Eddy et al., 2002).     46 Another significant concern with excessive re-solubilisation of phosphorous is the potential for the formation of an insoluble form of phosphorous called struvite (or magnesium ammonium phosphate, NH4MgPO4•6H2O). Struvite is extremely insoluble and forms crystalline deposits inside tanks, pipes, centrifuges and even within the AD leading to significant losses in hydraulic capacity and blockages throughout the sludge and centrate handling systems (Metcalf & Eddy et al., 2002). Struvite is notoriously difficult to remove and is very costly. Baur et al. (2002) found that costs related to the control and removal of struvite in a mid-size WWTP can easily exceed $100,000 per year, a figure likely to be much higher today.  Governments are pushing ever more stringent nutrient control standards; therefore, BNRs are becoming more common along with sludge from chemical control processes. AD of BNR sludge can be challenging, therefore the effects of any AD treatments need to be explored as they may have a significant impact on the WWTP.   Metal Salts to Control Odour 2.8There has been preliminary work that has explored various methods of controlling the production of VSCs from biosolids; however, most of it has been limited to VSC emissions from dewatered sludge cake or from small-bench-scale biochemical methane potential bottles. Some promising work has come from various studies which has shown that applying various metal salts to dewatered biosolids can reduce the generation of VSCs. Adams et al. (2008) demonstrated the link between cation concentration and odours from dewatered AD sludge and biosolids and iron and aluminum ion concentrations. Other studies have indicated that increased trivalent aluminum and iron ion concentrations in anaerobic digester feed sludge decrease the VSCs released from the dewatered biosolids (Adams et al., 2008; Higgins, 2010; Novak & Park, 2010).  Magnesium is another metal that deserves further exploration. Magnesium hydroxide has been used extensively for odour control in sewers by means of increasing sewer pH to 8 or greater retaining sulfides in the HS- form (Jefferson et al., 2002). Elevated magnesium levels have not been found to be detrimental to AD (United States Environmental Protection  47 Agency, 1979b), while magnesium hydroxide improves process stability. Furthermore, increased pH has been shown to negatively affect the SRBs and inhibit SRBs in slime layers (Jefferson et al., 2002) and have also been shown to prevent the degradation of MM into H2S (Leerdam et al., 2008).  One of the most cost-effective ways of elevating metal ion concentrations is through the addition of metal salts as many iron, aluminum, and magnesium salts are already used in many WWTPs for other purposes. Due to promising preliminary research from a review of available literature, the next chapters will focus on reducing VSCs with addition of metal salts.    48 Chapter 3 Materials and Methods 3    Chapter:      Experimental Design 3.1To explore the effects of adding metal salts to the feed of an operating AD, a full-factorial experimental design was conducted in two stages at the Bioreactor Technology Research Laboratory at the University of British Columbia Okanagan campus. This study examined the effects of dosing on a wide range of parameters including the production of odourous VSCs, pathogens remaining in the digestate, quantity of released nutrients, particle size and surface charges, dewaterability, along with a broad range of digester performance and stability metrics as discussed in sections 3.7 and 3.8.  Stage one employed two different chemicals added to two of three ADs (the third one was not dosed, and served as “control”), while stage two employed two different doses of three chemicals added to four of five ADs (the fifth digester was control) as shown in Figure 3.1. At each phase of the experiment, each digester was allowed to reach steady-state and then run for at least three times the SRT in question before switching to the next SRT. Stage one increased the total levels of iron and aluminum to 10 mg/L per gram of TS, while stage two employed a higher dose of 15 mg/L of each compound as ion per gram of TS. Stage two also included two other digesters to which 20 and 60 mg of magnesium hydroxide were added per gram of feed TS. In both stages, “control” digesters represented a conventional AD operation to use for comparison.    49  Figure 3.1 - Full-factorial experimental design (SRT: sludge retention time)   Metal Salt Selection 3.2Ferric chloride and aluminum sulfate were selected as the metal salts due to their low cost and widespread use in existing WWTPs. Magnesium hydroxide was selected as it has been shown to enhance sludge digestion and dewaterability (Wu et al., 2001) and has existing uses for sewer odor control (Jefferson et al., 2002). American Chemical Society (ACS) grade (97 to 102%) ferric chloride hexahydrate was purchased from Fisher Scientific (Thermo Fisher Scientific, Inc., Ottawa); a solution of 49% aluminum sulfate was acquired from Cleartech Chemicals (Saskatoon, Saskatchewan), while reagent grade (95%) magnesium hydroxide was purchased from Sigma-Aldrich (Oakville, Ontario).    50 3.2.1 Metal Doses Doses of ferric chloride and alum were selected based on metal ion concentrations that had been shown in literature to be effective in reducing VSCs from dewatered sludge cake (Higgins, 2010; Novak and Park, 2010). Doses of magnesium hydroxide were selected based on levels that had been shown to improve the dewaterability of digested sludge according to Wu et al. (2005).  3.2.2 Existing Metal Concentrations Existing sludge metal concentrations were initially estimated based on averages from nearly two years of monthly data provided by the City of Kelowna. This data indicated the average monthly values of the metal concentrations in the raw sludge from July 2010 through April 2012. Furthermore, representative samples of feed sludge were sent to an external laboratory to have the concentrations of metals verified (CARO Analytical Services Ltd., Richmond, British Columbia). The samples were prepared by CARO according to EPA 3050B (acid digestion of sediments, sludges, and soils), and analyzed by inductively coupled plasma-mass spectrometry (ICP-MS). The results of this analysis are shown in Table 3.1. The external analysis confirmed that the actual concentration of both aluminum and iron in the feed sludge differed by less than 10% from the estimated values from the historical data.   Table 3.1 - Feed sludge metal concentrations Metal Estimated Concentration in Feed* (mg/L) Dosed Amount (mg/L) Total Estimated Concentration (mg/L) Aluminum 3.585 6.415 10.000 Iron 3.245 6.756 10.000  Actual Concentration in Feed (mg/L) Total Actual Concentration (mg/L) Deviation from Estimated Value (%) Aluminum 4.200 10.615 6.15 Iron 3.700 10.456 4.56 *Estimated values were based on data provided by City of Kelowna    51  Anaerobic Digester Configuration 3.3Eight semi-continuously fed (feed was added and digestate was removed once daily) bench-scale mesophilic (35°C) ADs were run over a one-year period in two overlapping stages. Each AD was constructed out of a 1.5-L borosilicate glass erlenmeyer flask, and was sealed with a two-hole rubber stopper and silicone sealant as shown in Figure 3.2. Hollow Corning® borosilicate glass tubes were inserted through the stoppers for collection of digested sludge and headspace biogas in 2-L Tedlar® bags with polypropylene fittings (Chromatographic Specialties Ltd., Brockville, Ontario) while the digester was fed via the sidearm of the flask. When not being fed, the flasks were incubated at 35 ± 0.1°C in a New Brunswick/Eppendorf Innova 44R temperature controlled shaker (Eppendorf Inc., Enfield, Connecticut) at 80 revolutions per minute to emulate the conditions inside a full-scale complete mix-reactor.   Figure 3.2 - Laboratory scale anaerobic digesters   Inoculum and Acclimatization 3.4Inoculum was obtained from an existing 7 L New Brunswick BioFlo® 115 glass automated fermenter and control centre operating at the University of British Columbia’s Bioreactor Technology Research Laboratory. This digester operates as a continuous-flow, mesophilic  52 (35 ± 2°C) anaerobic digester with a SRT of approximately 20 days. This bench-scale AD has been functioning continuously since January 2012 and had originally been seeded with effluent from a full-scale AD operating at a WWTP in a nearby community (Penticton, BC). The bench-scale AD utilized the same feed sludge used for this study and was well acclimatized to it.   During each stage, the digesters were allowed to reach steady-state before dosing began, and then specific doses of ferric chloride, alum or magnesium hydroxide were added directly to the raw sludge (feed) immediately before being injected into the AD. Steady-state was considered to be the period where methane yield, pH and COD removals varied by 10% or less over a 10-day period. Initially each dosed digester was run at a 20-day SRT, which was then decreased to a 12-day SRT and then finally a 7-day SRT as shown in Table 3.2. During each stage and after changing each SRT, digesters were operated until steady-state was again reached. Data was collected during steady-state operation and each SRT was sustained for more than three times the SRT of interest. The total liquid volume in each digester was 750 ml and was fed 37.5, 62.5 and 107.1 ml of sludge daily at SRTs of 20, 12 and 7 days, respectively. Organic loading rates of the digesters were 1.89, and 3.15 and 5.40 g VS/L/d at corresponding SRTs of 20, 12 and 7 days, respectively. Most of the sludge digesters at the-full scale operate in a range of 0.7-3.0 g VS/L/d (American Chemical Society, 1971; Bolzonella et al., 2005). In this study, in addition to testing conventional organic loading rates, digesters were overloaded to assess the impact of metal addition on the stability of the digesters under high organic loadings. The effect of organic loading rates applied for the desired SRTs is discussed in the following chapter.    53 Table 3.2 - Digester sludge retention times  Stage 1 20 Day SRT Stage 1 12 Day SRT Stage 1 7 Day SRT Stage 2 20 Day SRT Stage 2 12 Day SRT Stage 2 7 Day SRT Minimum Steady-State Operation Recommended (days)* 60 36 21 60 36 21 Number of Days Operated at Steady-State (days) 137 48 56 81 62 44 Organic Loading Rate (g VS/L/Day) 1.89 3.15 5.40 1.89 3.15 5.40 *Ekama, Dold & Marais, (1986). Organic Loading Rate (grams of volatile solids per litre of digester volume per day).   Feed Sludge 3.5Feed was collected weekly from the City of Kelowna’s (population of ~150,000) municipal WWTP that employs a combination of primary, secondary and tertiary treatments and has a capacity of 70 million litres per day. This facility employs a tertiary treatment with a modified Bardenpho unit for nutrient control after primary and secondary treatment. Sludge from the primary setting tanks is thickened by gravity and is fed into a fermenter where it becomes fermented primary sludge (FPS). Waste activated sludge (WAS) from the BNR is thickened by dissolved air floatation (DAF). Excess sludge from these unit processes is mixed in a ratio of 33:67 FPS to WAS (by volume) before polymer is added and the mix is dewatered using a high-speed centrifuge. The sludge is thickened from between 4 to 5% TS by weight to approximately 18% TS by weight where it is collected and shipped to a composting facility in Vernon, B.C. for stabilization. Sludge for this project was collected after the FPS and WAS were mixed right before the point of polymer addition and centrifugation. Therefore, the bench-scale digesters were fed with a mixed sludge (at 4-5% TS) with the same volumetric ratio (33:67 by volume) of fermented primary and thickened secondary sludge. The treatment processes employed at the City of Kelowna WWTP are shown in Figure 3.3.   54  Figure 3.3 - City of Kelowna wastewater treatment processes  Various parameters of the feed sludge were monitored throughout each experimental phase to better understand the impacts of dosing on digester performance. Feed sludge pH, COD, alkalinity, solids (TS and VS), VFA composition and concentrations, ammonia were measured at least weekly and are summarized in Table 3.3 below.  Table 3.3 - Feed sludge characteristics Parameter Measured Value* pH 5.55 ± 0.20 Chemical Oxygen Demand (COD) 51,600 ± 5900 mg/L Alkalinity 1330 ± 860 mg/L Total Solids (TS) 4.2 ± 0.4% Volatile Solids (VS) 3.8 ± 0.4% Total Volatile Fatty Acids (TVFA) 2300 ± 590 mg/L Ammonia 200 ± 5 mg N/L *average values from January 2012 to February 2013    55  Key Metrics Measured 3.6 3.6.1 Volume and Composition of Biogas Biogas volumes collected in each digester’s Tedlar® bag were measured at roughly the same time each day using a manometer. The biogas composition was analyzed using two different gas chromatographs (GCs) in the UBC Bioreactor laboratory.  3.6.2 Biogas Volumes Biogas volumes were measured once per day from each digester’s Tedlar® bag. The collected biogas was pumped out of each Tedlar® bag into a manometer that displaced a known quantity of water. The volume of gas was calculated using a calibration curve created by comparing the displacements of biogas samples in the manometer to the displacements of known volumes of gas and then converted to STP and was corrected for any difference in sampling time.   3.6.3 Biogas Composition Digester biogas is typically composed of methane, carbon dioxide, nitrogen and oxygen with trace amounts of other compounds including moisture, hydrogen and VSCs. The main components of biogas (methane, carbon dioxide, nitrogen and oxygen) present in each digester were tested several times per week using the GC method originally established by van Huyssteen (1967). Biogas was analyzed using an Agilent 7820 GC equipped with a 3 meter Agilent G3591-8003/80002 packed column (2 mm internal diameter (ID), divinylbenzene stationary phase) and thermal conductivity detector (Agilent Technologies, Santa Clara, California). During each analysis, 0.5 ml of biogas was withdrawn from the gas line attached to the Tedlar® bag using an Agilent gas-tight syringe and was injected into the GC inlet septum and the punctured section of tubing was cut-off after sampling (to prevent biogas leakage). Oven, inlet and outlet temperatures were 70, 100 and 150°C, respectively. The GC used helium as the carrier gas (flow rate: 25 mL/min) which was provided by Air Liquide (Kelowna, B.C.).     56 3.6.4 Volatile Sulfur Compounds Each digester’s VSC composition was tested at least once a week during each stage of the experiment after the digesters reached steady-state. VSCs were determined using an Agilent 7890A GC equipped with a 30 meter J&W 123-1035 DB-1 column (0.32 µm ID, dimethylpolysiloxane stationary phase) and a flame photometric detector (Agilent Technologies, Santa Clara, California). A 1 ml sample loop was attached to the GC inlet port and was connected to sample injection and waste lines for sample purging. Both the sample loop and inlet port were treated with a Sulfinert® coating to minimize any reaction between the sulfur compounds in the biogas. Initial oven, inlet and detector temperatures were 50, 200 and 200°C, respectively, while the final temperatures of each aforementioned component was 200°C. The GC used helium as the carrier gas (flow rate: 75 mL/min) which was provided by Air Liquide (Kelowna, B.C.).   The GC was calibrated using three levels of custom gas mixture standards from Air Liquide that contained 1, 10 and 25 ppm each of hydrogen sulfide (H2S), methyl mercaptan, ethyl mercaptan, dimethyl sulfide, carbon disulfide, n−propyl mercaptan, ethyl sulfide with nitrogen comprising the remaining balance. Three additional gas standards also from Air Liquide were used which contained 1, 10 and 25 ppm of dimethyl disulfide with nitrogen comprising the remaining balance. The dimethyl disulfide was stored and calibrated separately due to potential for promoting degradation of the other VSC compounds within the gas standards (ASTM Designation: D5504 – 08). The calibration of the GC was spot checked every 2 to 3 weeks by re-running one or two levels of VSC standards and comparing the results to the known value. Once every 1 to 2 months, each of the gas standards was re-run and a full recalibration was performed to ensure accuracy.  To perform the analysis, biogas was drawn from each digester’s Tedlar® bag into a 60 ml gas-tight syringe that was then attached to the sample inlet line of the GC. Slight pressure was put on the plunger of the syringe to allow the sample to purge the lines, sample loop and inlet port slowly for 30 to 60 seconds. Before beginning the analysis, pressure on the syringe was released for 3 to 4 seconds to allow any excess pressure to dissipate.   57 Biogas from each bioreactor contained levels of VSCs that would often saturate the detector, so each sample was diluted before it entered the GC column. Two analysis methods were developed that used either a 1:10 or a 1:20 biogas to nitrogen split within the GC inlet and were calibrated separately using the aforementioned standards. Levels of methyl mercaptan, ethyl mercaptan, dimethyl sulfide, carbon disulfide, n−propyl mercaptan, ethyl sulfide typically were sufficiently low enough to be quantified using the 1:10 inlet split, however H2S was present in much higher quantities thus was measured using a 1:20 split. In instances where H2S levels still saturated the detector using the 1:20 split, an external dilution was prepared using medical grade pure nitrogen (99.0%) from Praxiar Ltd. (Kelowna, B.C.) in an evacuated 2-L Tedlar® bag to bring the sample’s H2S concentration within the detection range of the GC before injection. Each week the VSCs in each digester were tested twice to accurately measure H2S along with the seven other VSCs.   3.6.5 Most Probable Number Analysis for Pathogenic Microorganism Enumeration The most probable number (MPN) of pathogenic microorganisms was quantified by using indicator organisms (total coliforms and Escherichia coli) which are commonly used to indirectly measure pathogens in water and waste samples. Measuring total coliforms (versus measuring fecal coliforms only) provides a conservative estimate of the MPN of fecal coliforms as the EPA and BC OMRR only regulate fecal coliforms. If the measured total coliforms are within EPA or OMRR class A or B limits, then fecal coliforms will be even further within the regulated limits.  Levels of total coliforms and E. coli were determined using IDEXX Colilert media in IDEXX Quanti-Tray/2000 media trays (IDEXX Laboratories, Inc., Westbrook, ME). This method can count up to 2419 colony forming units (CFU)/mL without dilution and provides confidence limits comparable to membrane filtration analysis methods (IDEXX Laboratories, 2012).  Colilert uses a proprietary Defined Substrate Technology® media to detect total coliforms and E. coli simultaneously. This media uses a nutrient-rich media along with two different nutrient-indicators, ortho-nitrophenol (ONPG) and 4-Methylumbelliferyl β-D-Glucuronide  58 (MUG) which can be metabolized by the coliform enzyme β-galactosidase and the E. coli enzyme β-glucuronidase, respectively (IDEXX Laboratories Inc., 2014). The media also contains a variety of other salts, nutrients and other agents to avoid interference or false-positives from non-target organisms with similar metabolic pathways (IDEXX Laboratories, 2012). When ONPG is metabolized by lactose reducing organisms, it cleaves a yellow indicating dye which causes a positive sample well to change colour from clear to yellow. When MUG is metabolized by E. coli, it releases a bluish indicator dye which fluoresces under long wave ultraviolet light (UV) (366 nm) (Cheeptham & Lal, 2013). A sample well must be both yellow (from the ONPG) and fluoresce (from MUG) under 366 nm light in order to be considered positive for E. coli.  Sludge digestate was collected into sterilized containers and was analyzed within one hour of sampling. Serial dilutions were performed using sterilized glassware and Type 1 water to bring the MPN of the digestate within the quantification range of the method, after which the manufacturer’s procedure was followed. Type 1 water is ultra-pure water that had nearly all impurities and ions been removed through a distillation (or comparable) process, which is then followed by an ion exchange resin and 0.2 µm filter to produce exceptionally pure water (ASTM International, 2011). The trays were incubated for 24 hours at 35 ± 0.1°C in a Thermotron S-1.5C benchtop environmental chamber (Thermotron Industries, Holland, Michigan). A fraction of each digestate sample used for MPN analysis was used to determine the TS of sample so that the results could be normalized and reported as per EPA and B.C. OMRR regulations.  3.6.6 Dewaterability Two tests that are commonly performed to estimate the dewaterability of sludge is the capillary suction test (CST) and specific resistance to filtration. Zeta potential is a measurement that can be used to measure the strength of the repulsion between particles and can be used to estimate the impacts of a treatment on the resistance of particles coming together. Both CST and zeta potential measurements were performed in this study on the various digester effluents to gauge what (if any) impact that the dosed metals had on the dewaterability of the digester effluents.   59  Capillary Suction Time 3.6.6.1The CST test involves the placement of 5 ml of sludge into a metal cylinder which rests on a 50 by 50 mm section of Type 17 chromatography paper. The liquid portion from the sludge is drawn into the filter paper by capillary suction forces and radiates outward from the cylinder while the solids in the sludge are retained above the filter. Electrical contacts on a plastic top plate placed at two different radii are connected to a timer which measures the amount of time it takes for the filtrate to soak outward from the first set of electrodes to the outer set. The required time was recorded and was normalized by the percentage TS in each sample so that the results can be compared.   CST testing was performed on each digester towards the end of each SRT at least three times per digester and in triplicate. Digested sludge was wasted from each digester and was split into two separate beakers that were promptly sealed with Parafilm wrap. One beaker was used to determine the total solids in each sample, while the other was gently stirred before withdrawing each 5 ml CST sample for analysis. Temperatures of the sludge samples were also recorded, as CST is a temperature dependent parameter.    Zeta Potential and Conductivity 3.6.6.2The sludge zeta potentials were tested using a Zetasizer Nano ZS (Malvern Instruments Ltd., Westborough, MA) which also measures the electrical conductivity of the sample. The Zetasizer Nano ZS uses laser Doppler Micro-electrophoresis to measure zeta potential. This works by applying an electric field to the sample solution, which causes the particles to move with a velocity related to their zeta potential. This velocity is then measured using a laser interferometric technique called phase analysis light scattering. From this, the Zetasizer can calculate the electrophoretic mobility, zeta potential, and zeta potential distribution of the particles within the sample matrix (Malvern Instruments Ltd., 2012).   Zeta potential measurements were made from room temperature, uncentrifuged sludge (as to not disturb the floc structure) which had been withdrawn from the AD less than 2 hours before measurement. Sludge samples were filtered through a 47 mm diameter, 1 μm Whatman™ membrane filter to remove any suspended particles while retaining only the  60 particles within the colloidal and dissolved ranges (Thermo Fisher Scientific, Inc., Ottawa). The filtrate was collected into a 1.5 ml plastic cuvette and was injected into a disposable capillary cell (Malvern model DTS1070) for measurement. The cell was held in the machine for a minimum of 60 seconds to allow the temperature of the sample to equilibrate to 25°C, and each sample was measured a minimum of 10 times with 15-30 sub-runs per analysis to ensure data reliability.  3.6.7 Nutrient Levels Centrate from digester feed and effluents were analyzed for ammonia and orthophosphate by first centrifuging raw and digested sludge samples in 55 ml centrifuge tubes with a Sorval LEGEND XT centrifuge for 15 minutes at 8,000 rpm (Thermo Fisher Scientific, Inc., Ottawa). After centrifugation, each sample was filtered through a disposable 13mm Millex™ 0.45 μm nylon syringe filter into clean plastic test tubes that were then diluted with Type 1 water until the concentrations were within the test ranges. Samples were capped and frozen at approximately -25°C in a laboratory freezer until sufficient quantities of samples had accumulated to process a batch.  Batches of samples were unfrozen and analyzed using an Astoria-Pacific Method A2 Autoanalyser with an Astoria-Pacific 311 automated sampler to determine the concentrations of ammonia and orthophosphate in the centrate and were multiplied by the dilution ratio of the original sample.  3.6.8 Particle Size The size distribution of the particles within the sludge matrix was measured several times during each SRT using a Malven Mastersizer 3000 equipped with a Hydro LV wet dispersion module (WDM) (Malvern Instruments Ltd., Westborough, MA.) The LV WDM holds 600 ml of sample and dispersant and is attached to a measurement cell that sits inside the Mastersizer 3000. Sonication and/or a variable-speed impeller inside the module can be used to help to keep particles dispersed while the internal pump circulates the dispersed sample through a closed loop between the WDM and the measurement cell.  61 The Mastersizer 3000 uses laser diffraction to measure the distribution and size of particles by measuring scattered light as the dispersed particles flow through the laser light path. A series of detectors both in front and behind the sample measure the angle and intensity of the scattered light which is used to calculate the size and number of particles that created the scattering pattern (Malvern Instruments Ltd., 2013). Samples are analyzed over a wide range of angles using both red (632.8 nm) and blue (470 nm) light wavelengths.   The feed and effluent from each digester was tested multiple times during each SRT. Samples of fresh sludge were withdrawn from each digester were analyzed within 1 hour of sampling. Before analysis the WDM was filled with 600 ml of reverse osmosis (RO) treated water and a baseline was established without any sample. Once the machine was ready to read, the sample was gently mixed and approximately 3-5 ml of sludge was placed into the WDM using a transfer pipette until the amount of light dispersed was within the target range (6-13%). Each sample was analysed by the machine by using a minimum of 10 runs under both wavelengths of light.   The Mastersizer 3000 would output both the particle size distribution of the sample along with D10, D50 and D90 which represents the maximum particle diameter below which 10%, 50% and 90% of the sample volume exists respectively (Malvern Instruments Ltd., 2013). Before each sample and between samples, the WDM was repeatedly flushed with RO water and cleaned with high-power mixing and sonication to prevent any carry-over between samples.   3.6.9 Sludge Dynamic Viscosity The dynamic viscosity of raw and digested sludge was measured 20 times over a period of four months to better understand the physical properties of the raw and the digested sludges, as viscosity is an important operational parameter. Both doses of each compound were measured in triplicate for each measurement. Measurements were made using a Brookfield DV-E Viscometer with an LV-5 spindle. Samples were equilibrated until they measured 25 ± 0.5°C before reading. Before each run, a sample of Type 1 water was measured to ensure the machine was operating correctly.    62 Samples of both raw and digested sludge were sampled and tested at approximately the same time each day (± 30 minutes). Representative samples of raw and digested sludge from each digester were placed into 60 ml centrifuge tubes, capped to prevent evaporation, and were allowed to equilibrate to 25 ± 0.5°C prior to measurement. Since viscosity is highly dependent on temperature, temperatures were verified using a temperature probe before each measurement. Each centrifuge tube contained 45 ml of sample and was recapped and gently inverted 2-3 times to ensure the sample was well-mixed. The spindle was then lowered into centre of the sample (to avoid interference from the walls of the vessel) until the liquid level reached the measurement mark on the spindle. The measurement would begin after the machine reached a stable value and at least 5 complete rotations or until 30 seconds had elapsed, and would be re-capped and re-inverted between each measurement. Between each sample, the spindle would be cleaned using RO water and the spindle would be dried using a tissue.   Digester Performance 3.7Various digester parameters were measured to assess the impacts of metal dosing on digester operation, performance and/or stability. Parameters measuring digester performance included removals of COD and solids (TS and VS) across the digesters. Key digester operational metrics included pH, alkalinity, and VFA levels in the digesters. Other metrics including protein and humic acid concentrations along with particle size were also measured on both raw sludge and effluent samples from digesters.  3.7.1 Chemical Oxygen Demand COD of the feed and digested sludges were measured using the closed reflux colorimetric method according to Standard Methods for the Examination of Water and Wastewater procedure 5220D. This method utilizes a change in the oxidation state of dichromate ions that causes a colour change in the sample as COD in the sample is oxidized. The change in absorbance is directly proportional to the amount of chromium ions changing from a hexavalent to trivalent state that can be measured using a spectrophotometer at a 600 nm wavelength.    63 Feed sludge and AD effluent both contain levels of COD that far exceed the maximum range of the method thus samples were diluted using RO water and were thoroughly mixed for 2 minutes at 8,000 RPM using a Kinematica™ Polytron™ PT 10-35 GT Benchtop Homogenizer. 10 ml of the diluted sludge samples were transferred into triplicate Kimax™ borosilicate glass vials to which COD digestion and catalyst solutions were added. 6 ml of a COD digestion solution containing mercuric sulfate, potassium dichromate, sulfuric acid and water were added to each vial along with 14 ml of a COD catalyst solution containing 98% sulfuric acid and silver sulfate. The tubes were then capped with a polytetrafluoroethylene (PTFE) faced rubber-lined screw cap, mixed using a vortex mixer and digested for 3 hours at 150 ± 0.1°C in a Thermotron S-1.5C benchtop environmental chamber (Thermotron Industries, Holland, Michigan). The COD tubes were measured a 600 nm wavelength using a Spectronic 20D+ (Thermo-Electron Corporation) spectrophotometer for most of stage one and two of the digester studies. Towards the end of stage two the Spectronic 20D+ was replaced with a Thermo Scientific™ GENESYS™ 10S UV-Vis Spectrophotometer (Thermo Fisher Scientific, Inc., Ottawa). COD results were calculated by comparing the absorbance to known quantities of compared to a calibration curve generated from the digestion of potassium hydrogen phthalate solution (Sigma BioXtra >99.95% pure). A sample calibration curve is shown in Appendix B.  3.7.2 Solids Solids (TS and VS) entering and exiting an AD is an important metric that can be directly impacted by metal dosing. TS and VS in digester feed and effluent were measured at least weekly using Standard Methods 2540 B and 2540 E respectively.   CoorsTek™ porcelain evaporating dishes were prepared for use by soaking them in a 20% sulfuric acid solution, rinsed with water and dried in a furnace at 550°C before being cooled and stored in a desiccator. Well-mixed, representative samples were added to the dry dishes and weighed on a Mettler-Toledo Excellence XA-105 analytical balance (Thermo Fisher Scientific, Inc., Ottawa) before being placed in a gravity oven set at 98 ± 2°C. Once the visible free water had evaporated, the temperature was increased to 105 ± 2°C and the dishes were left to dry overnight to a constant weight. After cooling, the dishes were weighed again  64 and then fired in a furnace at 550°C and allowed to cool before the final weighing. The difference between the wet and dry weights represents the TS fraction (Equation 3.1), while the difference between the dry weight and the ash represents the VS fraction (Equation 3.2). Equation 8 – Total solids calculation % 𝑇𝑇𝑇𝑇𝑇𝑇𝑇𝑇𝑇𝑇 𝑆𝑆𝑇𝑇𝑇𝑇𝑆𝑆𝑆𝑆𝑆𝑆 (𝑇𝑇𝑆𝑆) =   𝑚𝑚𝑚𝑚𝑚𝑚𝑚𝑚 𝑜𝑜𝑜𝑜 𝑑𝑑𝑑𝑑𝑑𝑑 𝑚𝑚𝑚𝑚𝑚𝑚𝑠𝑠𝑠𝑠𝑠𝑠𝑚𝑚𝑚𝑚𝑚𝑚𝑚𝑚 𝑜𝑜𝑜𝑜 𝑤𝑤𝑠𝑠𝑤𝑤 𝑚𝑚𝑚𝑚𝑚𝑚𝑠𝑠𝑠𝑠𝑠𝑠 𝑥𝑥 100%   Eq. (3.1) Equation 9 – Volatile solids calculation % 𝑉𝑉𝑇𝑇𝑇𝑇𝑇𝑇𝑇𝑇𝑆𝑆𝑇𝑇𝑉𝑉 𝑆𝑆𝑇𝑇𝑇𝑇𝑆𝑆𝑆𝑆𝑆𝑆 (𝑉𝑉𝑆𝑆) =   𝑚𝑚𝑚𝑚𝑚𝑚𝑚𝑚 𝑜𝑜𝑜𝑜 𝑑𝑑𝑑𝑑𝑑𝑑 𝑚𝑚𝑚𝑚𝑚𝑚𝑠𝑠𝑠𝑠𝑠𝑠 − 𝑚𝑚𝑚𝑚𝑚𝑚𝑚𝑚 𝑜𝑜𝑜𝑜 𝑚𝑚𝑚𝑚ℎ𝑚𝑚𝑚𝑚𝑚𝑚𝑚𝑚 𝑜𝑜𝑜𝑜 𝑤𝑤𝑠𝑠𝑤𝑤 𝑚𝑚𝑚𝑚𝑚𝑚𝑠𝑠𝑠𝑠𝑠𝑠 𝑥𝑥 100% Eq. (3.2)   Digester Operational Metrics 3.8 3.8.1 pH and Alkalinity pH and alkalinity are extremely important AD operational parameters and were measured in digester feed and effluents regularly. pH measurements were conducted using a combination pH electrode attached to an Accumet™ Excel XL25 pH/mV/Temperature/ISE meter (Thermo Fisher Scientific, Inc., Ottawa), while alkalinity was measured according to Standard Methods for the Examination of Water and Wastewater Method 2320B. This method uses a dilute sulfuric acid solution to determine the concentrations of various carbonate, bicarbonate and hydroxide ions in an aqueous sample using Equation 3.3 below. Equation 10 – Alkalinity calculation 𝐴𝐴𝑇𝑇𝐴𝐴𝑇𝑇𝑇𝑇𝑆𝑆𝐴𝐴𝑆𝑆𝑇𝑇𝐴𝐴 �𝑚𝑚𝑚𝑚𝐿𝐿𝑇𝑇𝑆𝑆 𝐶𝐶𝑇𝑇𝐶𝐶𝑂𝑂3� =   𝐴𝐴 𝑥𝑥 𝑁𝑁 𝑥𝑥 50,000𝑚𝑚𝑠𝑠 𝑜𝑜𝑜𝑜 𝑚𝑚𝑚𝑚𝑚𝑚𝑠𝑠𝑠𝑠𝑠𝑠     Eq. (3.3) Where A = ml of standard acid added and N = normality of standard acid    65 Approximately 30 ml of sludge was placed into a glass erlenmeyer flask and the exact volume of the sample was recorded. A magnetic stir bar was added to the flask and both were placed on a magnetic stir plate and were titrated down to a pH value of 4.6 using a 50 ml burette filled with 0.1 N sulfuric acid.   3.8.2 Volatile Fatty Acids Centrate from centrifuged feed sludge and each digester’s effluent was sampled weekly and each sample was filtered through a 0.22 μm membrane filter to remove any suspended matter that may damage or clog the GC column. Two vials containing 0.5 ml of the filtrate from each digester were sealed in a sample vial, capped and stored in a laboratory freezer at approximately -25°C until a sufficient number of samples were collected. Before analysis, each sample was thawed and 0.5 ml of an internal standard containing sulfuric acid, formic acid and 2% isobutyric acid are added to each vial as an internal standard before GC analysis.   VFA present in the samples (including acetic, propionic, and butyric acid) were measured using an Agilent 7890A GC equipped with a 25-meter Agilent 19091F-112 column (0.32 mm ID, nitroterephthalic acid modified polyethylene glycol stationary phase), a flame ionization detector (FID) as well as an auto-sampler (Agilent Technologies, Santa Clara, CA). Initial oven, inlet and detector temperatures were 70, 220 and 300ºC, respectively, while the final temperatures each aforementioned component was 200, 220 and 300ºC, respectively. The GC used helium as the carrier gas (flow rate: 40 mL/min) which was provided by Air Liquide (Kelowna, B.C.). A mixed standard containing 2000 mg/L of each of the three VFAs were analysed before each run and after every 20 samples to ensure VFA recovery was within acceptable ranges (±50 mg/L).   3.8.3 Protein and Humic Acid Changes in total protein and humic acid (HA) within the sludge samples were ascertained by using a protocol based on Lowry et al. (1951) but modified according to Frølund et al. (1995). The modified protocol uses a combination of chemicals that include aqueous sodium carbonate, sodium hydroxide, copper (II) sulfate, potassium sodium tartrate and a Folin-Ciocalteu phenol reagent to cause a colour change that can be detected by a  66 spectrophotometer. In alkaline conditions, the copper (II) ions in the protein reagent react with a peptide found in protein and are reduced by the Folin reagent to produce a blue colour. Humic acids contain phenolic functional groups which develop a blue colour in the presence of Folin-Ciocalteu phenol reagent (Sharma & Krishnan, 1966). Both of these colour changes can be detected and quantified by a spectrophotometer at a wavelength of 750 nm.  Protein and HA in raw and digested sludges were determined several times during each SRT. Sludge samples were diluted using RO water to bring the concentrations of protein and HA within the detection range. Vials containing standards were made to create calibration curves using either bovine serum albumin for protein analysis (Thermo Fisher Scientific, Inc., Ottawa) or a humic acid standard solution (Sigma-Aldrich, Ltd.). These standards were created and measured each time the test was performed to help ensure accurate results. To analyze the samples, 0.5 ml of diluted sludge was pipetted into 13 × 100 mm disposable borosilicate glass culture tubes along with 2.5 ml of either the protein or HA reagent solution. The tubes were covered with parafilm and placed in the dark for 10 minutes, after which 0.25 ml of the Folin-Ciocalteu phenol reagent (Thermo Fisher Scientific, Inc., Ottawa) was added and the tubes were again covered and returned to the dark. After an additional 30 minutes, three samples of 200 μl from each vial was pipetted into wells on a Fisher Scientific™ 96 well microplate and read using BioTek Synergy HT microplate reader (BioTek Instruments, Inc., Winooski, VT) set at 750 nm wavelength. Sample readings and calibration curves are shown in Appendix C. Absorbance of the protein and HA solutions were calculated according to the equations proposed by Frølund et al. (1995) as shown in Equation 3.4 through Equation 3.7 below: Equation 11 – Total absorbance 𝐴𝐴𝑇𝑇𝑜𝑜𝑤𝑤𝑚𝑚𝑠𝑠 =  𝐴𝐴𝑃𝑃𝑑𝑑𝑜𝑜𝑤𝑤𝑠𝑠𝑃𝑃𝑃𝑃 + 𝐴𝐴𝐻𝐻𝐻𝐻𝑚𝑚𝑃𝑃𝐻𝐻 𝐴𝐴𝐻𝐻𝑃𝑃𝑑𝑑    Eq. (3.4) Equation 12 – Blank absorbance 𝐴𝐴𝐵𝐵𝑠𝑠𝑚𝑚𝑃𝑃𝐵𝐵 =  0.2𝐴𝐴𝑃𝑃𝑑𝑑𝑜𝑜𝑤𝑤𝑠𝑠𝑃𝑃𝑃𝑃 + 𝐴𝐴𝐻𝐻𝐻𝐻𝑚𝑚𝑃𝑃𝐻𝐻 𝐴𝐴𝐻𝐻𝑃𝑃𝑑𝑑   Eq. (3.5) Equation 13 – Protein absorbance 𝐴𝐴𝑃𝑃𝑑𝑑𝑜𝑜𝑤𝑤𝑠𝑠𝑃𝑃𝑃𝑃 =  1.25(𝐴𝐴𝑇𝑇𝑜𝑜𝑤𝑤𝑚𝑚𝑠𝑠 −  𝐴𝐴𝐵𝐵𝑠𝑠𝑚𝑚𝑃𝑃𝐵𝐵)    Eq. (3.6) Equation 14 – Humic acid absorbance 𝐴𝐴𝐻𝐻𝐻𝐻𝑚𝑚𝑃𝑃𝐻𝐻 𝐴𝐴𝐻𝐻𝑃𝑃𝑑𝑑 =  𝐴𝐴𝐵𝐵𝑠𝑠𝑚𝑚𝑃𝑃𝐵𝐵 − 0.2𝐴𝐴𝑃𝑃𝑑𝑑𝑜𝑜𝑤𝑤𝑠𝑠𝑃𝑃𝑃𝑃    Eq. (3.7)   67 Chapter 4 Results and Discussion 4    Chapter:   Characterization of Waste Sludge Streams 4.1Key chemical and physical characteristics of the undigested mixed sludge from the City of Kelowna’s WWTP are presented in Table 4.1. The pH of the feed sludge is somewhat acidic (5.55) as it contains roughly two-thirds FPS which contains high levels of VFA. Levels of ammonia in the feed were well below levels reported to be inhibitory to AD (Appels et al., 2008; Ye Chen et al., 2008). Furthermore, low pH values of the feed and mesophilic AD operating temperatures helped to further reduce the risk of ammonia inhibition during digestion.   Feed sludge parameters were relatively consistent throughout the experiment, however there were several instances of process upset at the Kelowna WWTP that were noticeable in the feed. There were several wide fluctuations in VFA concentrations including VFA spikes up to 2400 mg/L, and drops to below 700 mg/L. These fluctuations often coincided with sudden pH changes leading to a maximum pH of 6.13 and a minimum pH of 5.06. TS and VS concentrations were relatively consistent around 4.2% and 3.8% respectively, however there were TS and VS increases up to 5.0 and 4.0% and drops down to 3.2 and 2.9% respectively. Many of these fluctuations in the digester feed sludge characteristics had discernable effects on digester performance metrics including levels of biogas production and feed and digester COD levels. Table 4.1 – Average feed sludge characteristics Parameter Value Standard Deviation Number of Replicates pH 5.55 0.20 207 COD (mg/L) 51,600 5900 68* TS (% w/w)  4.19 0.40 66* VS (% w/w) 3.80 0.4 66* VS/TS 0.908 0.01 66* Ammonia (mg N/L) 200 - - Alkalinity (mg/L as CaCO3) 1120 300 23 VFA (mg/L)** 1500 590 62** *Samples were analyzed each time in triplicate. **Samples were analyzed each time in duplicate.  68  Species of VSCs in AD Biogas 4.2Steady-state headspace VSCs concentrations in each digester fluctuated from week to week due to varying feed characteristics and conditions within the digesters. Total VSCs (TVSCs) in stage one and stage two control digesters ranged from 2070 ppm to 550 ppm, averaging 1460 ppm across all SRTs.   Figure 4.1 shows the average VSC composition of the 15 mg/L ferric chloride and alum dosed digesters along with the 60 mg/L magnesium hydroxide dosed digesters during the 20-day SRT. Similar results were observed in each of the lower dosed digesters and across each 20, 12 and 7 day SRTs. H2S was the dominant VSC in the headspace of each digester which confirms the results of existing literature that has examined the constituents of AD biogas (Appels et al., 2008; Higgins et al., 2006; Metcalf & Eddy et al., 2002). As suggested by Higgins et al. (2006), methyl mercaptan (MM) was the second most prevalent VSC species present followed by dimethyl sulfide (DMS). The metabolic pathways for the creation and destruction of VOSCs have been well established.    69  Figure 4.1 - Average volatile sulfur compound composition (Stage Two digesters, 20-day sludge retention time, n=5)  An interesting observation was made when the 10 mg/L alum dosed digester begun to exhibit signs of instability shortly after reaching steady-state during the final 7-day SRT. Digester pH, alkalinity and methane volumes decreased sharply as VFA levels exceeded 2000 mg/L, strongly suggesting methanogenic inhibition. While this occurred, VOSC levels increased sharply as shown in Figure 4.2. MM spiked up to 89.6 ppm, nearly an 8-fold increase while DMS increased to 6.4 ppm, nearly double the control digester. Higgins et al. (2006), Du and Parker (2009b) and others have observed that methanogenic inhibition prevented the degradation of VOSCs within serum bottles; these results suggest that this observation extends to AD biogas when municipal sludge is digested. There are several reasons for this increase that are discussed in section 4.3.2.   70  Figure 4.2 - Volatile sulfur levels after 10 mg/L alum digester failure (n=1)   Effects of Dosing on VSCs 4.3Ferric chloride and alum had significant effects on VSC production; however, magnesium hydroxide had little effect as shown in Table 4.2. The highest doses of both ferric chloride and aluminum sulfate had the greatest impact on the production of VSCs across all SRTs; however, the results for the magnesium hydroxide varied between doses and SRTs (Table 4.2).   71 Table 4.2 – Average steady-state digester volatile sulfur compound concentrations in ppm   Hydrogen Sulfide Methyl Mercaptan Ethyl Mercaptan Dimethyl Sulfide Carbon Disulfide N−Propyl Mercaptan Ethyl Sulfide Dimethyl Disulfide Total VSC Removal 20-Day SRT 10 mg/L Control 1204.87 16.38 3.79 5.92 4.69 3.58 1.64 3.14 1244.02   10 mg/L Ferric Chloride 264.53 11.47 2.68 5.54 3.66 3.12 1.49 5.98 282.71 77% 10 mg/L Alum 4960.98 20.51 5.07 5.87 5.27 3.49 2.61 5.83 4941.98 -297% 15 mg/L & Mg(OH)2 Control 998.31 8.26 4.57 5.29 5.12 4.77 4.81 5.24 1036.36   15 mg/L Ferric Chloride 109.75 5.69 3.51 4.64 4.92 4.73 4.26 4.56 142.06 86% 15 mg/L Alum 9791.04 18.39 7.54 7.63 5.69 4.83 4.11 5.08 9844.31 -850% 20 mg/L Magnesium Hydroxide 371.80 8.60 4.48 5.06 5.09 4.68 4.73 5.89 407.69 61% 60 mg/L Magnesium Hydroxide 668.87 6.14 4.56 4.72 5.35 4.77 4.82 5.24 701.57 32% 12-Day SRT 10 mg/L Control 1318.75 8.21 4.99 5.52 5.31 5.28 4.78 5.08 1446.40   10 mg/L Ferric Chloride 380.53 7.09 4.99 5.12 5.31 4.53 4.68 5.81 418.06 71% 10 mg/L Alum 5532.41 22.12 6.39 8.70 5.44 5.54 5.47 6.78 5592.84 -287% 15 mg/L & Mg(OH)2 Control 1459.12 371.92 4.19 4.47 5.40 3.49 2.81 4.54 1538.32  15 mg/L Ferric Chloride 331.07 8.43 3.23 4.27 5.32 2.51 2.72 4.88 362.44 76% 15 mg/L Alum      15605.36 54.56 7.37 8.38 5.85 4.51 2.89 4.01 15692.94 -920%  72  Hydrogen Sulfide Methyl Mercaptan Ethyl Mercaptan Dimethyl Sulfide Carbon Disulfide N−Propyl Mercaptan Ethyl Sulfide Dimethyl Disulfide Total VSC Removal 20 mg/L Magnesium Hydroxide 1108.09 9.85 3.20 4.74 5.00 3.07 2.67 3.15 1139.77 26% 60 mg/L Magnesium Hydroxide 1630.70 7.78 2.69 4.03 4.31 2.54 2.22 3.64 1984.06 -29% 7-Day SRT 10 mg/L Control 1270.56 9.17 4.20 6.07 4.96 4.34 4.36 5.21 1308.87  10 mg/L Ferric Chloride 523.85 6.71 4.31 4.96 5.38 4.47 4.43 4.31 558.41 57% 10 mg/L Alum 10997.89 38.26 6.57 10.01 5.59 5.24 4.82 6.09 11074.46 -746% 15 mg/L & Mg(OH)2 Control 2152.50 13.93 3.23 4.14 5.10 2.90 2.50 6.20 2190.50  15 mg/L Ferric Chloride 249.73 6.84 3.17 4.05 5.10 2.85 2.50 6.20 280.44 87% 15 mg/L Alum* n/a n/a n/a n/a n/a n/a n/a n/a n/a n/a 20 mg/L Magnesium Hydroxide 2554.74 17.27 3.44 5.75 5.12 2.92 2.50 6.20 2597.93 -19% 60 mg/L Magnesium Hydroxide 1776.84 14.88 3.32 5.58 5.12 2.91 2.50 6.20 1817.35 17% *15 mg/L Alum dosed digester failed before final sludge retention time (SRT).    73 A two-way analysis of variance (ANOVA) was performed to determine if the VSC results are statistically significant. ANOVA is an important analysis tool for analyzing the differences between group means and examining variation among and between groups. It allows one to accept or reject the hypothesis that there is (or is not) a statistically significant difference between groups within a confidence interval. The level of significance is determined by the P-value, the lower of which indicates the higher level of certainty that there is a statistically significant difference. At a 95% confidence limit (CL), a P-value ≤ 0.05 means that there is a significant difference between the means. A P-value ≥ 0.05 would allow one accept the hypothesis that there is not difference between the means with a 95% CL (Montgomery, 2013). ANOVA also generates an F-ratio that allows one to determine the level of significance between the factors. The factor with the highest F-ratio has the highest level of importance (Montgomery, 2013).   For this ANOVA, TVOSC was used as the response variable, while the chemical dose, digester SRT and any interactions between the two variables were used as factors. TVSC ANOVA results are presented in Table 4.3. Normal probability plots for residuals are presented in Appendix H.   Table 4.3 - ANOVA of total volatile sulfur compound results Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical Dose 9 1012865025 112540558 32.96 0.000 Digester SRT 2 65026968 32513484 9.52 0.000 Interaction 2 14890427 7445214 2.18 0.117 Error 138 471211100 3414573   Total 151 1545226082     The results of the ANOVA indicated that one could say with 95% confidence that both chemical dose and digester SRT have statistically significant differences between their means. However, the interaction between the two factors was not statistically significant at the 95% confidence limits. Of the two factors, chemical dose has the larger impact on TVSCs with higher F-ratio. Figure 4.3 and Figure 4.4 compare the VSC concentrations between the digesters.  74  Figure 4.3 – Stage One digester headspace total volatile sulfur compound concentrations (n=8, 6, 5)  Figure 4.4 - Stage Two digester headspace total volatile sulfur compound concentrations (n=11, 5, 4)  75 4.3.1 Ferric Chloride Previous work by Higgins (2010), Novak and Park (2010) and others suggesting that elevated levels of iron in sludge can have a beneficial impact on VSC production was confirmed by this study. Using ferric chloride to elevate divalent iron ion concentrations to 10 mg/L greatly decreased TVSC levels while ion concentrations of 15 mg/L were even more effective.  The 10 mg/L ferric chloride dose reduced TVSC concentrations by an average of 77.3 and 77.1% during both the 20-day and 7-day SRTs versus the control respectively, while a 15 mg/L ferric chloride dose reduced TVSC concentrations by an average of 86% during the 20-day SRT and 87% during the 7-day SRT versus the control as shown in Table 4.2. TVSC measurements of the 15 mg/L ferric chloride digester were as low as 36 ppm, which is exceptionally low for untreated AD biogas.  Higgins et al. (2006) determined that H2S was chemically precipitated as iron(II) sulfide (FeS) when sufficient iron was present in most types of wastewater sludge. Furthermore, Dentel and Gossett (1982) have shown that the addition of iron and aluminum can reduce the bioavailability of protein and other molecules through complexation; however, it is likely that chemical precipitation of sulfide as FeS plays a larger role. Although ferric chloride is a moderately strong Lewis acid and was shown to slightly decrease the pH of the AD, any additional volatilization of HS- ions due to the decrease in digester pH was more than offset by the precipitation and complexation of sulfur ions within the digester. Adams et al. (2008) observed that ironbound biopolymers are often released during high-shear centrifugation leading to increased VSC generation in dewatered biosolids. Therefore, the addition of ferric chloride may lead to increased odour emissions from sludge storage, composting or land-application facilities.  4.3.2 Aluminum Sulfate Alum had the complete opposite effect as both doses immensely increased VSC levels across all SRTs to extremely high levels. The 10 mg/L alum dose increased VSC concentrations versus the control by an average of 298% during the 20-day SRT and 746% during the 7-day SRT versus the control while the 15 mg/L alum dose increased VSC concentrations versus  76 the control by an average of 850% during the 20-day SRT and 920% during the 12-day SRT as shown in Table 4.2. Although the digester failed before the final 7-day SRT was reached, one could infer that had the digester continued to operate, the VSC levels would have been even higher due to high protein and sulfate loading at such a short SRT.  H2S in the 15 mg/L alum dosed digester registered nearly 20,000 ppm (or 2% of the total gas volume) shortly before failure. H2S concentrations at this level are extremely dangerous to both people and equipment. Any leak at this level would likely harm facility workers and/or the public as H2S is slightly heavier than air and during calm atmospheric conditions it can travel great distances and will collect in low-lying areas (Tarver & Dasgupta, 1997). H2S in moist environments causes corrosion in steel and in reinforced concrete, thus high H2S levels will increase the probability of leaks in tanks, piping as well as in biogas handling and utilization equipment due to accelerated corrosion. Furthermore, the smallest fugitive emission at such high concentrations could result in significant odour issues over a wide area. For example, just 1 m3 of leaked biogas containing 10,000 ppm of H2S could cause over 21 million m3 of ambient air to exceed the odour threshold of H2S (0.00047 ppm).  Despite previous work by Higgins (2010) and others that had shown that adding alum before polymer addition and dewatering decreased VSC levels from stored dewatered sludge cake, using alum to achieve trivalent aluminum ion concentrations of 10 and 15 mg/L in AD feed dramatically increased VSC levels in the biogas. The most likely explanation for this result is due to the chemical composition of alum versus the aluminum contained within it. The chemical formula for alum is Al2(SO4)3;  therefore each mole of alum added will also add three moles of sulfate to the digester. The conversion of inorganic sulfate into H2S by SBRs like desulfovibrio is well known, thus adding such large amounts of additional sulfate likely encouraged the proliferation of various SBR species within the alum-dosed digester. A metagenomic comparison of dosed versus an undosed AD would likely confirm this hypothesis, which was outside the scope of this work.    77 4.3.3 Magnesium Hydroxide The effects of magnesium hydroxide dosing were somewhat inconclusive and fluctuated significantly from day to day. Initial results during the 20-day SRT indicated that the lower magnesium hydroxide dose had a greater effect on VSCs than the higher dose, decreasing them by 61% and 32% respectively versus the control. These results did not hold when the SRT was decreased to 12 days as the higher dose increased VSCs by 29%, while the lower dose decreased VSCs by 26% relative to the control. The results changed again at the 7-day SRT, when the lower dose decreased VSCs by 19% while the higher dose increased them by 17% versus the control.   Despite its use in sewer odor control, magnesium hydroxide had inconsistent effects on VSCs. There was no discernable pattern to the magnesium hydroxide results or apparent correlation to digester pH, VFA, or ammonia levels suggesting some other factor or random fluctuations were involved. A second 2-way ANOVA comparing the stage two control versus the magnesium dosed digesters was performed. Similar to the first ANOVA, it compared VSCs, SRT and any interaction between them. The chemical dose had a P-value of 0.084 and the interaction between dosing and SRT had a P-value of 0.434, neither of which met the minimum P-value for a 95% CL to be significant. Similar to the first ANOVA, SRT had a significant effect and had a P-value of 0.000 indicating it was very significant.   Effect of Dosing on Pathogens 4.4There were significant fluctuations in digester effluent samples (digestates) in terms of pathogens (MPN) from week to week, and over the course of the experiment. An ANOVA was performed to interpret the results that used MPN as the response variable and the chemical dose, digester SRT and any interaction between the two as factors and is shown in Table 4.4. At a 95% CL, the ANOVA generated very low P-values for both the chemical dose and SRT, suggesting that both factors did have a statistically significant effect on pathogens. SRT had the most effect as it had the highest F-ratio.    78 Table 4.4 – ANOVA of digestate pathogen concentrations (MPN/gram dry weight)* Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical Dose 7 3.02E+13 4.31E+12 5.35 0.000 Digester SRT 2 1.89E+13 9.44E+12 11.72 0.000 Error 69 5.56E+13 8.06E+11   Total 78 1.10E+14    *Interactions between chemical dose and digester SRT were tested but were not significant.  Figure 4.5 and Figure 4.6 show the average normalized MPN of total coliforms per dry weight in stage one and two digesters. Similar to the VSC results, there was a significant amount of variation from day to day; however, the difference between the controls vs. dosed digesters and between SRTs was readily apparent. As expected from literature (Han & Dague, 1997), longer SRTs resulted in smaller numbers of pathogens with two exceptions in stage two.   Figure 4.5 – Stage One digestate total coliform densities (n=3, 4, 3) EPA/BC OMRR Class B Limit for Fecal Coliforms      79  Figure 4.6 - Stage Two digestate total coliform densities (n=3, 4, 3)  4.4.1 Ferric Chloride The 10 mg/L ferric chloride dose was effective in reducing total coliforms and E-coli versus the control digesters, decreasing the MPN of total coliform by approximately 24, 24 and 35 percent at 20, 12 and 7-day SRTs respectively. E-coli counts were reduced by proportional, but slightly smaller amounts. The trend was not as clear for the 15 mg/L ferric chloride dosed digester, as the 4% decrease in MPN of total coliform was well within the margin of error, while the 24% and 48% increase during the 12 and 7-day SRTs respectively contradict the initial findings.   4.4.2 Aluminum Sulfate Both doses of alum were very effective in reducing coliforms and E-coli during each SRT. The 10 mg/L alum dose decreased the MPN of total coliforms by approximately 55, 38 and 68 percent versus the control at 20, 12 and 7-day SRTs respectively. The 15 mg/L alum dose decreased the MPN of total coliforms by approximately 55 and 86 percent versus the control at 20 and 12-day SRTs respectively. The 15 mg/L digester failed before it reached the final EPA/BC OMRR Class B Limit for Fecal Coliforms    80 7-day SRT, therefore data was not collected for coliforms or E-coli. Both alum doses at each SRT showed a significant reduction in total coliforms and E. coli. Pathogen destruction appeared to increase as VFA levels increased, which may have been a result of increased VFA levels or more acidic conditions within the digester.   An extensive review of the available literature provided little insight as to the method of action for the reduction of pathogens; however, pH, VFA accumulation, and/or agglomeration may play some role. McFeters & Stuart (1972) reported that varying pH values had effects on pathogen survival, thus lower pH values in the alum digesters may have been a factor. Salsali & Parker (2006) observed similar pathogen destruction results at lower pH values and at elevated VFA levels. Since alum is used as a coagulant, it was surmised that alum dosing might cause pathogens to become bound within the flocs themselves rather than increasing the rate of pathogen destruction or inactivation. This theory was tested on two different occasions where a sample of diluted effluent from the stage two alum dosed digester was split immediately after dilution. Half of the sample was used as a control for the MPN analysis, while the other half underwent gentle shear mixing for one minute at a low RPM using a Polytron stand mixer with a sterilized shear blade. On both occasions, the MPN of the disturbed sample was nearly identical to the undisturbed sample suggesting that entrapment within the particles during agglomeration is not a significant factor.   4.4.3 Magnesium Hydroxide MPN results from the magnesium hydroxide dosed digesters were inconsistent, similar to the VSC results. In several instances magnesium hydroxide dosing actually increased the levels of both total coliforms and E-Coli. The 60 mg/L dose decreased the MPN of total coliforms by approximately 9% and 6% during the 20 and 7-day SRTs respectively while it increased total coliforms by approximately 70% during the 12-day SRT. The 20 mg/L dose increased total coliforms by 11, 68, and 48% during the 20, 12 and 7-day SRTs respectively.     81 Magnesium hydroxide appeared to have a stimulating effect on the MPN of both total coliforms and E. Coli; this may be due to magnesium’s role as an essential nutrient for many bacterial enzymes and cell walls (Madigan et al, 2010). Magnesium cations have been shown to be stimulatory at a concentration range of 75-150 mg/L and may help to reduce the toxicity of other cations present at inhibitory levels (Wu, 2002).   Effects of dosing on dewaterability 4.5 4.5.1 Capillary Suction Time Capillary suction time (CST) of sludge samples varied significantly from week to week, even when normalized by TS of digestate samples as recommended in Standard Methods in method 2710G. The results from the final 7-day SRT in stage two were not as expected as the shortest SRT had higher CST values than either the 20 or 12-day SRTs in each stage two digester. This is abnormal as shorter SRTs typically reduce the amount of hydrolysis that can occur within the digester thus improving dewaterability and decreasing CST. One possible explanation for this was the laboratory switched to a new supplier of CST chromatography paper in-between the 12 and 7-day SRT during stage two which may have affected CST results. The City of Kelowna WWTP experienced some disturbances in their treatment processes during the end of stage one and for most of stage two of this experiment, which may have caused differences in the PS to WAS ratios within the digester feed sludge. This could have also had an influence on CST results.  An ANOVA with a 95% CL indicated that chemical dosing did not have a statistically significantly impact on sludge CST as shown in Table 4.5. Thus, one could interpret that dosing did not affect the dewaterability of the sludge. The only factor that was significant was the SRT of the digester with the sludge from digesters with the shortest SRT being easiest to dewater during stage one, and during the 20 and 12-day SRTs of stage two. The CST results are shown in Figure 4.7 and Figure 4.8 for stage one and stage two digesters, respectively.     82 Table 4.5 - ANOVA for capillary suction time (CST) results* Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical dose 7 1236324 176618 1.95 0.071 Digester SRT 2 653775 326887 3.61 0.031 Error 84 7601283 90491   Total 93 9.624E+10    *Interactions between chemical dose and digester SRT were tested but were not significant.    Figure 4.7 - Stage One digestate average capillary suction time (CST) results (n=4, 4, 5)     83  Figure 4.8 - Stage Two digestate capillary suction time (CST) results (n=4, 4, 3)  A review of existing literature has shown that both iron and alum salts help to enhance filtration of undigested WAS by bridging negatively charged surface sites between sludge particles (Forster, 2001), however this effect was not apparent during either stage. The two doses of magnesium hydroxide were selected based on levels that had been shown to enhance dewaterability of batch digested sludge (Wu et al., 2001; Wu et al., 2005; Wu & Bishop, 2004), however these levels of magnesium hydroxide did not appear to generate a statistically significant results in the semi-continuously fed reactors used in this study. This may be due to differences between batch, complete-mix, and semi-continuously fed ADs.  4.5.2 Zeta Potential Similar dewaterability results were observed when comparing the zeta potential of the sludges. Although alum is often used as a coagulant and is known to act by modifying the floc surface charges, no significant zeta potential differences were observed in any of the iron, alum or magnesium hydroxide dosed digesters. An ANOVA with 95% CL indicated that chemical dosing did not have a statistically significantly impact on sludge zeta potentials shown in Table 4.6. Zeta potential results are shown in Figure 4.9 and Figure 4.10.  84  Table 4.6 – ANOVA for digestate zeta potential results* Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical Dose 7 97.49 13.927 1.39 0.223 Digester SRT 2 43.64 21.82 2.17 0.121 Error 80 803.68 10.046   Total 89 944.06    *Interactions between chemical dose and digester SRT were tested but were not significant.   Figure 4.9 - Stage One digestate zeta potential (n=12, 5, 4)   85  Figure 4.10 - Stage Two digestate zeta potential (n=7, 5, 4)   Effects of Dosing on Nutrients 4.6Nutrients (i.e. N and P) released during AD of BNR sludge are a well-known barrier to the more widespread use of AD. Discharge of nutrients into surface waters (particularly phosphorous), can lead to excess algae growth and eutrophication as phosphorous is a limiting nutrient in most freshwater systems. A two-way ANOVA with a 95% CL showed that both chemical dosing and the SRT resulted in statistically significant differences in the levels of orthophosphate solubilised during AD as shown in Table 4.7. Digestate soluble (<0.45 µm) orthophosphate concentrations are plotted in Figures 4.11 and 4.12, while total orthophosphate concentrations are tabulated in Appendix D. The SRT did have an impact on orthophosphate levels during both stages; however, the results conflicted between them as orthophosphate levels in the first stage were highest during the 20 and 7-day SRTs, while in stage two nutrient levels increased significantly as the SRT decreased across all digesters. The conflicting results between stage one and stage two are likely due to process fluctuations and an upset at the City of Kelowna’s WWTP that occurred during 12 and 7-day SRT of stage one. Orthophosphate in the feed sludge decreased from an average of 560 mg P/L during the 20-day SRT to less than 310 mg P/L during the 20-day SRT. Feed orthophosphate levels begun to increase during the 7-day SRT approaching near normal values (520 mg P/L).   86 Table 4.7 - ANOVA for digestate orthophosphate results* Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical Dose 8 2415204 301900 10.77 0.000 Digester SRT 2 992601 496301 17.71 0.000 Error 172 4821222 28030   Total 182 7588696    *Interactions between chemical dose and digester SRT were tested but were not significant.    Figure 4.11 - Stage One digestate total orthophosphate concentrations (n=7, 6, 5)  87  Figure 4.12 - Stage Two digestate soluble orthophosphate concentrations (n=6, 5, 4)   88 4.6.1 Ferric Chloride Both levels of ferric chloride were effective in reducing the concentration of orthophosphate in AD effluent achieving a maximum orthophosphate removal of 60.8% with the 10 mg/L dose during the 12-day SRT when compared to the AD feed. The higher dose achieved a maximum orthophosphate removal of nearly 51% during the initial 20-day SRT, however this decreased to 24 and 19% during the 12 and 7 day SRTs for the 15 mg/L ferric chloride digester.  Ferric chloride is one of the most commonly used chemical for struvite control in wastewater treatment facilities (Water Environment Federation, 2008). Trivalent iron and phosphate ions form a variety of precipitates including iron(III) phosphate (FePO4(s)), and other iron phosphate species like Fe2.5PO4(OH)4.5(s) depending on the molar ratios of iron and phosphate (Fytianos et al., 1998). Excess iron and an acidic pH are required to achieve optimum removal of phosphate (Fytianos et al., 1998).  4.6.2 Aluminum Sulfate Alum was also effective in controlling orthophosphate at both levels of dosing. The largest effect observed was a 66.9% reduction in orthophosphate versus the feed containing the 10 mg/L dose during the 12-day SRT. The 15 mg/L dose during the 20-day SRT decreased average orthophosphate levels to less than 240 ± 54.8 mg P/L, and continued to be effective during the final 12-day SRT before and after the digester failure.  Alum is another commonly used chemical for controlling phosphorus in wastewater. Trivalent aluminum in alum combines in equimolar ratios to form insoluble aluminium phosphate (AlPO4(s)) (Wu et al., 2001) and various other aluminum-phosphorous compounds (Yang et al., 2006). Therefore, orthophosphate released during the digestion of PAOs contained BNR sludge are likely controlled by alum via chemical precipitation.     89 4.6.3 Magnesium Hydroxide Both doses of magnesium hydroxide tested were slightly effective in controlling the release of orthophosphate, as dosing resulted in lower levels of soluble orthophosphate than the control digesters with the 60 mg/L dose having the greatest impact. This effect was noted across each SRT for both doses with the exception of the 20 mg/L dose during the 20-day SRT that experienced a 13.3% increase versus the control.   Magnesium hydroxide is often used in AD side-stream treatment processes to remove phosphorous, as magnesium is an essential component of struvite (NH₄MgPO₄·6H₂O) which can be precipitated out of the sludge supernatant solution (Wu & Bishop, 2004). Struvite is normally a nuisance material within WWTPs, however since it contains equimolar ratios of magnesium, ammonia and phosphorous, it makes an excellent fertilizer. Wu et al. (2004) used the addition of magnesium hydroxide to strip phosphorous in reaction vessels and found that struvite nucleation was enhanced by the addition of magnesium hydroxide.   Effects of Dosing on Protein/Humic Acid 4.7Although protein degradation supplies a considerable amount of the sulfur-bearing amino acids that are VSC precursors (Higgins et al., 2006), any correlation between VSC and protein levels was not discernable in this study. Chemical dosing did not have a statistically significant impact on the extent of protein degradation according to an ANOVA with 95% CL. The results of the ANOVA are shown in Table 4.8.   Table 4.8 - ANOVA for digestate protein results* Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical Dose 7 5403408 771915      0.49 0.837 Digester SRT 2 3439734 1719867      1.08 0.349 Error 34 53898101   1585238   Total 43 63152692    *Interactions between chemical dose and digester SRT were tested but were not significant.     90 Although digestate protein levels did not show a significant difference, there was a statistically significant impact of chemical dosing on humic acid concentrations according to an ANOVA with 95% CI. The results of the ANOVA are shown in Table 4.9, while average protein and humic acid levels within each digester are presented in Figures 4.13 and 4.14 as well as in Appendix E.  Digester SRT had a larger effect on humic acid concentration than chemical dosing as it had the largest F-ratio. Several humic acids produced during the biodegradation of organic matter can be difficult to degrade during digestion, thus higher organic loading rates at shorter SRTs may lead to higher levels. Both alum dosed digesters had the lowest levels of humic acid across each SRT (with the exception of the 20-day SRT of the stage two alum digester). This may be due to the increased VOSC levels produced by the alum dosed digesters as Higgins et al. (2006) found that humic acids were an important contributor of methyl groups for the formation of VOSCs.  Table 4.9 - ANOVA for digestate humic acid results* Factors Degrees of freedom Sum of squares Mean square F-ratio P-value Chemical Dose 7 57650472 8235782     3.76     0.004 Digester SRT 2 45541055   22770527     10.40 0.000 Error 34 74448258    2189655   Total 43 176943366    *Interactions between chemical dose and digester SRT were tested but were not significant.    91  Figure 4.13 - Stage One digestate average humic acid concentrations (n=2, 2, 2)  Figure 4.14 - Stage Two digestate average humic acid concentrations (n=2, 2, 2)    92  Effects of Dosing on Digester Performance 4.8 4.8.1 TS/VS Removals There is a substantial body of literature suggesting that concentrations of various metals in sludge can have an effect on solid (TS and VS) removals during anaerobic sludge digestion (Park (2002), Higgins (2010), Novak and Park (2010), Wu (2002), etc.). This was confirmed in this study as ferric chloride, alum and magnesium hydroxide all had an adverse effect on the removal of both TS and VS in the digested sludge. Figures 4.15 through 4.20 show the effects of metal additions on both TS and VS removals and VS/TS ratios.     Figure 4.15 - Stage One total solids (TS) removals of digesters (n=19, 8, 8)  93  Figure 4.16 - Stage Two total solids (TS) removals of digesters (n=15, 8, 6)  Figure 4.17 - Stage One volatile solids (VS) removals of digesters (n=19, 8, 8)   94  Figure 4.18 - Stage Two volatile solids (VS) removals of digesters (n=15, 8, 6)  95  Figure 4.19 - Stage One VS/TS ratios of digestates (n=19, 8, 8)   Figure 4.20 - Stage Two VS/TS ratios of digestate (n=15, 8, 6)    96  Ferric Chloride 4.8.1.1Although Adams et al. (2008), Park et al. (2006) and Higgins et al. (2010) found that increasing iron concentrations in raw sludge increased VS removals, Higgins et al. (2010) found that iron added within the plant for phosphorous control had an adverse effect on VS removals. Iron contained within wastewater behaves very differently than iron added at the WWTP as influent iron is predominately bound within organic matter, while external iron additions within the facility are predominately inorganic and quickly reduce both organic and inorganic matter (Adams et al., 2008). The results from the iron dosed digesters support this finding.  Both doses of ferric chloride had an adverse effect on TS and VS removals in each digester during each SRT. These effects were up to a 6.1% decrease in the average TS removal and a 6.8% decrease in average VS removal. This effect was consistent across each SRT with the higher dose having the most adverse effect. This is likely due to the precipitation of sulfur compounds as iron (II) sulfide, a dense substance with a melting point far above the 550°C used for VS testing (Royal Society of Chemistry, 2014). A significant portion of the various sulfur species are being shifted from the gas/liquid phase into the sludge solids; this is further supported by the increase in the VS/TS ratios (Figures 4.19 and 4.20) which suggest an increase in mineralization within the digested solids.   Aluminum Sulfate 4.8.1.2Aluminum concentrations within the digested sludge have been shown to have an effect on the levels of VS removals; moderate levels of aluminum have been shown to decrease VS removals slightly (Gossett et al. 1978; Higgins, 2010). Gossett et al. (1978) and Park et al. (2006) and Higgins (2010) suggested that this was due to the reduced bioavailability of “aluminum-associated organics, particularly proteins” (Higgins, 2010).   Alum dosing had similar effects to that of ferric chloride; both alum doses had a negative impact on both TS and VS removals in each digester and across all SRTs with the exception of the initial 20-day SRT of the 15 mg/L digester. TS removals were impacted by up to 4.0%, while VS was impacted up to 6.2% during the shortest SRT of the 10 mg/L digester. Since  97 the 15 mg/L digester failed before it reached the final 7-day SRT, it is very likely that TS and VS removals would have been affected even further. Alum dosing appear to affect VS/TS ratios suggesting a change in sludge mineralization, however it was far more variable as it increased in some digesters and decreased in others at different SRTs. Exploration of the effects on VS and TS removal by using an alternative aluminum bearing compound like polyaluminum chloride would likely yield more definitive results.    Magnesium Hydroxide 4.8.1.3Little literature exists examining the effects of magnesium hydroxide on TS/VS removals. Wu et al. (2001) found that a magnesium hydroxide dose of 300 mg/L was stimulatory to AD and improved overall digestion and total suspended solid (TSS) removals from anaerobically digested sludges; however, they did not report the effects of digestion on total or volatile solids. Wu et al. did not report the TS of the undigested sludge, nor did they normalize the results based on solids so a direct comparison is difficult.  Magnesium hydroxide had a mixed, but slightly negative overall impact on both TS and VS removals at both doses across each SRTs challenging Wu et al.’s claim that magnesium hydroxide improved digestion. The greatest impact of a 3.9% decrease was observed in the 20 mg/L dosed digester during the 12-day SRT.  4.8.2 Chemical Oxygen Demand Removals One key digester performance metric is the removal of COD from AD sludge as COD removals are directly related to the production of biogas. A multifactor ANOVA examining the effects of dosing, SRT and any interaction between them was performed using a 95% CL. The ANOVA indicated that there was not any significant interaction between dosing and the SRT and showed that dosing did not have a statistically significant impact on effluent COD levels; however, the SRT did. It is expected that longer SRTs would achieve greater COD reductions due to an increased amount of time for the organics to be solubilized and utilized. The results of the ANOVA are shown in Table 4.10 and graphs of COD removals are shown in Figure 4.21 and Figure 4.22.     98 Table 4.10 - ANOVA for digester COD results* Factors DF Sum of squares Mean square F-ratio P-value Chemical Dose 7 96069830    13724261      1.64 0.125 Digester SRT 2 1425221352   712610676     85.37 0.000 Error 197 8347081    Total 206     *Interactions between chemical dose and digester SRT were tested but were not significant.   Figure 4.21 - Stage One digester chemical oxygen demand (COD) removal (n=18, 7, 5)   99  Figure 4.22 - Stage Two digester chemical oxygen demand (COD) removal (n=12, 10, 5)  4.8.3 Biogas Production The composition (volume ratio of gases present) of the biogas produced varied little between each digester and each SRT, therefore both metal dosing and SRT had little or no effect on the gas composition. Figure 4.23 and Appendix H show the average biogas composition of each digester throughout both stages. Although there were some fluctuations in the normalized average daily volume of methane produced per gram of COD removed, there was not a statistically significant difference (P-value > 0.05) in methane production between each digester as shown in Figure 4.24 and Figure 4.25.  The average daily volume of methane produced from each of the digesters was somewhat lower than the theoretical maximum yield of 0.350 litres of methane per gram of COD removed and ranged from 0.245 to 3.330 litres of methane per gram of COD removed. This is most likely due to small leaks from the AD and/or the Tedlar® gas collection bag, measurement error and/or recalcitrant compounds within the sludge matrix.    100  Figure 4.23 - Digester biogas composition (n=74,74,65,64,64,52,64,64)  101  Figure 4.24 - Stage One digester normalized methane volumes  Figure 4.25 - Stage Two digester normalized methane volumes  102 4.8.4 Effects of Dosing on Sludge Dynamic Viscosity The dynamic viscosity of the raw and digested sludge varied significantly from day to day. One could not conclude that dosing had any significant impact on the dynamic viscosity of the sludge as an ANOVA with a 95% CL exceeded the 5% P-value. Sludge dynamic viscosity ranged from 300 to 813 centipoise (cP) averaging 543 cP. Zhou (2013) found the dynamic viscosity of aerobically digested WAS varied between 80 to 260 cP; however, it is difficult to compare these results to literature as viscosity is highly variable with temperature, pH and solid concentrations (Zhou, 2003). Few researchers have explored the dynamic viscosity of AD sludge and instead have focused on the viscosity of centrate or sludge after polymer addition (Erden, 2004; Ormeci, 2007). The results of the ANOVA and average sludge dynamic viscosities are shown in Table 4.11 and Appendix F respectively.   Table 4.11 – ANOVA for dynamic viscosity of raw and digested sludge Factors Degrees of Freedom Sum of squares Mean square F-ratio P-value Chemical Dose 7 264135    37734      1.70 0.115 Error 117 2592334    22157   Total 124 2856468     4.8.5 Effects of Dosing on Sludge Particle Size There was little evidence that either chemical dosing or SRT had a significant effect on the particle size distribution of the sludge flocs within the digesters. Results were mixed throughout stage one and two, as one compound would appear to have an effect on the particle size distribution during one SRT only to have the opposite effect at a different SRT or chemical dose. One would expect there to be some difference as alum is often used as a flocculating agent, and longer SRTs should increase the amount and extent of organic hydrolysis leading to an increased quantity of fine materials. The lack of an effect may be due to the chemical precipitation of the added metal ions with the various sulfur species present within the digesters. Sample particle size distribution plots are shown in Figures 4.26 and 4.27 while complete particle size results are presented in Appendix G.    103  Figure 4.26 – Stage One digester particle size distribution   Figure 4.27 – Stage Two digester particle size distribution    104  Effects of Dosing on Digester Operation and Stability 4.9During each 20-day SRT, all digesters at both dose levels reached steady-state quickly and did not exhibit any indication of instability or inhibition. Common indicators of digester stability including alkalinity, ammonia, VFA and pH levels are shown in Table 4.12 and were within established expected ranges. Within each 12-day SRT, all digesters performed well with the exception of the 15 mg/L alum digester that began to exhibit signs of instability (decreasing pH and alkalinity, increasing VFA) shortly after reaching steady-state.   105 Table 4.12 - Steady-state average digester pH, alkalinity, ammonia and VFA levels Digester SRT pH Alkalinity (mg/L as CaCO3) Ammonia (mg N/L) Total VFA (mg/L)   Control (Stage 1) 20 7.50 5350 760 58 n, σ 150, 0.17 12, 390 18, 90 14, 24 12 7.32 4850 990 119 n, σ 38, 0.15 5, 210 6, 100 6, 140 7 7.21 3800 1030 446 n, σ 36, 0.14 4, 334 4, 140 8, 28 10 mg/L Ferric Chloride 20 7.44 5050 780 24 n, σ 150, 0.16 12, 540 18, 50 14, 14 12 7.32 4600 1010 64 n, σ 38, 0.16 5, 80 6, 110  6, 66 7 7.17 3900 860 200 n, σ 34, 0.15 4, 340 4, 120 5, 4 10 mg/L Alum 20 7.52 5400 920 17 n, σ 150, 0.59 12, 591 18, 120 13, 11 12 7.35 5100 970 30 n, σ 37. 0.16 5, 230 6, 30 6, 12 7 6.30 2600 1100 3176 n, σ 15, 1.01 3, 1561 3, 110 9, 1331 Control (Stage 2) 20 7.41 5100 1250 35 n, σ 76, 0.15 10, 260 8, 180 16, 28 12 7.33 4400 1210 78 n, σ 32, 0.07 4, 1360 7, 140 6,66 7 7.28 5650 1140 13 n, σ 49, 0.07 3, 170 4, 200 5, 4  106 15 mg/L Ferric Chloride 20 7.38 4900 1150 16 n, σ 75, 0.17 10, 300 8, 200 15, 16 12 7.23 3900 1060 41 n, σ 32, 0.10 4, 200 7, 110 7, 42 7 7.11 4450 1100 33 n, σ 49, 0.11 3, 340 4, 280 9, 44 15 mg/L Alum 20 7.35 5200 1090 15 n, σ 75, 0.15 10, 330 8, 100 18, 8 12 6.72 2800 1060 1967 n, σ 18, 0.78 3, 1300 3, 220 4, 702 7 - - - - n, σ - - - - 20 mg/L Magnesium Hydroxide 20 7.34 5050 1150 98 n, σ 76, 0.15 10, 340 8, 200 18, 144 12 7.39 4650 920 78 n, σ 31, 0.09 4, 230 7, 140 6, 69 7 7.29 5200 1240 15 n, σ 48, 0.09 3, 220 4, 240 4, 3 60 mg/L Magnesium Hydroxide 20 7.35 5200 1060 83 n, σ 76, 0.14 9, 280 8, 190 18, 121 12 7.35 4400 1120 531 n, σ 31, 0.11 4, 400 7, 125 6, 718 7 7.39 5300 990 56 n, σ 48, 0.13 3, 305 4, 100 5, 14 n= number of measurements, σ= standard deviation 107 4.9.1 pH Steady-state pH values were relatively stable and were well within expected values from literature (Appels et al., 2008). The pH of the control digesters typically ranged between 7.2 and 7.5, while the ferric chloride digesters were somewhat lower (likely due to the acidic nature of ferric chloride) and ranged between 7.1 and 7.4. Alum digesters had slightly higher operating pH values ranging between 7.3 and 7.7; however, pH values in both digesters had strong downward trends as organic loading rates increased. Magnesium hydroxide dosing also increased average pH values within the digesters that typically ranged between 7.2 and 7.6.  4.9.2 VFA & Alkalinity Both VFA and alkalinity levels are key AD stability parameters and often foreshadow pending issues long before other symptoms like decreased biogas volumes begin to manifest. VFA levels in the digester feed ranged between 725 and 2450 mg/L, while digesters ranged between 10 and 200 mg/L during normal operation and rose with each shorter SRT. Both the 10 mg/L and 15 mg/L alum dosed digesters experienced ever increasing levels of VFA as organic loading rates increased until VFA levels reached a point where they became inhibitory and biogas production ceased.   Alkalinity in the AD feed was typically quite low and ranged between 570 and 1900 mg/L (as CaCO3), while digester alkalinity ranged between 3500 and 7100 mg/L (as CaCO3) during normal operation. During the first 20-day SRTs and part way into the 12-day SRTs, both alum dosed digesters averaged between 5 to 20% more alkalinity than either of the control or the iron dosed digesters. This may be due to the mechanism by which sulfate is reduced by SBRs to generate H2S (Equation 2.3) where each mole of inorganic sulfate is combined with two moles of carbon and water to generate two moles of bicarbonate alkalinity and one mole of H2S.      108  Effects of Dosing on Digester Stability 4.10None of the chemicals dosed appeared to adversely affect the operation or stability of the ADs during long SRTs and at low organic loading rates. Both doses of ferric chloride and magnesium hydroxide did not appear to have any adverse effects on digester operation; however, alum did have a significant negative impact as the experiments progressed.  Initially alum did not appear to adversely affect digester operations; however, it was evident that the 15 mg/L dose was very close to the level of inhibition as it took very little time to push the digester to the point of failure. The 10 mg/L dose was better tolerated, although it too failed at higher organic loading rates whereas the other digesters continued to operate without issue.  Jackson-Moss and Duncan (1991) reported that an acclimatized AD could tolerate aluminum ion concentrations up to 2500 mg/L, while Cabirol et al. (2003) observed inhibitory effects of aluminum on methanogenic and acetogenic bacteria at concentrations near 1000 mg/L. Therefore, it is unlikely that the aluminum added feed was the sole source of inhibition. The instability is likely due to the alum itself as each mole of alum contains three moles of sulfate, so it is very probable that sulfate ions increased to the point where they became inhibitory to the methanogens but not the acetoclastic bacteria causing ever-increasing levels of VFA, which further inhibited the methanogens. The combination of aluminium and sulfides in digester feed have been shown to have synergistic inhibitory effects on methanogens (Cabirol et al., 2003)  Park and Novak (2007) have suggested that organic materials associated with aluminum are not readily degradable anaerobically, while Cabirol et al. (2002) noted difficulties digesting alum sludge and low conversion rates of organic matter to methane that likely helps to account for the slightly decreased VS removals in the alum dosed digesters. Cabirol et al. (2003) also noted VFA accumulations and decreases in buffering capacity when anaerobically digesting alum sludge at elevated organic loading rates and toxicity of sulfide to the anaerobic culture.     109 Chapter 5 Conclusion 5    Chapter: Conclusion  Summary 5.1This full-factorial experimental study conducted over two stages explored the effects of adding two different levels of three common metal salts on some of the most common barriers to the wider adoption of AD. These include the production of corrosive and odourous VSCs, pathogens remaining in the digestate; along with the re-release nutrients (specifically orthophosphate) from the digested solids into the digester supernatant as the released nutrients can overwhelm biological nutrient removal systems and cause excessive algae growth in receiving waters.  Conditions within a full-scale WWTP are never static as the quantity and composition of the waste stream continually changes along with the performance of the biological treatment systems. Therefore, AD sludge characteristics are never exactly the same from week to week. Although this study was conducted in a laboratory under semi-controlled conditions, there were significant variations in the feed sludge entering the digesters that added significant amounts of variability to the results. Despite the variability, this research has demonstrated that both ferric chloride and aluminum sulfate in particular may have a role in overcoming some of the barriers to more widespread implementation of AD.   5.1.1 Volatile Sulfur Compounds The results of this study validate the hypothesis that the addition of metal salts has an effect on the production of VSCs in AD biogas. Ferric chloride is a valuable chemical if odour control, infrastructure protection, or the control of hydrogen sulfide in biogas is a priority. A dose of 10 mg/L as ion per gram of TS is very effective in controlling VSCs achieving up to a 77% removal in TVSCs, while a dose of 15 mg/L as ion per gram of TS is even more effective and can reduce TVSCs by up to 87% versus an undosed (control) digester. Dosing with ferric chloride may be an option instead of adding additional unit processes to remove H2S depending on the intended use of the biogas.    110 Alum is contraindicated for controlling odour and VSC production in biogas. Furthermore, digestion of sludge that contains alum added for coagulation and /or for phosphorous control is very likely to generate high levels of VSCs when digested. Alum added to increase aluminum ion concentrations to 15 mg/L as ion per gram of TS increased TVSC levels by up to 920% to levels exceeding 16,000 ppm, levels which are extremely dangerous to people and property. At this concentration, the smallest fugitive emission from an AD or from biogas utilization or handling equipment could create significant odour problems for a large area, and/or could endanger workers if the emissions occurred within an enclosed area or under quiescent atmospheric conditions. These levels may greatly increase the likelihood of corrosion damage from the acids formed from the combination of moist surfaces and high VSC levels.  Results from both levels of magnesium hydroxide dosing were inconclusive, during some SRTs VSC levels decreased while they increased in others. This suggests that the dose was either insufficient, the magnesium had little or no effect, or some interaction occurred that was not apparent.  5.1.2 Pathogens  The effects of ferric chloride on pathogens were mixed; the 10 mg/L dose during stage one resulted in a modest reduction in both total coliforms and E-coli; however, the complete opposite was demonstrated by the 15 mg/L dosage during stage two. Further work would need to be completed to make a more definitive determination on the effectiveness of ferric chloride and pathogens.   Both levels of alum were extremely effective in reducing the levels of pathogens in anaerobically digested sludge. 15 mg/L of alum was able to reduce total coliforms by up to 86% and may be a valuable tool for meeting Class B pathogen targets for the land application of biosolids.    111 Magnesium hydroxide is contraindicated for pathogen control and was shown to increase levels of total coliforms and E-coli in digested sludge. The most likely reason for this is due to magnesium’s role as an essential micronutrient for many microbial cells.  5.1.3 Nutrients Ferric chloride was far more effective in reducing nutrients in AD effluent versus the control and reduced soluble orthophosphate concentrations by nearly 61%. This is most likely due to the chemical precipitation of phosphorous as iron (III) phosphate. Alum was even more effective achieving a near 70% reduction in orthophosphate, likely by the production of various aluminum-phosphorous compounds inducing insoluble aluminium phosphate. Magnesium hydroxide had little effect on orthophosphate and reduced orthophosphate by up to 13.3%, likely due to the production of struvite within the wasted sludge or within the digester itself, potentially creating problems in the future.  5.1.4 Digester Operation and Stability Dosing with magnesium hydroxide did not have a statistically significant effect on the dewaterability of digested sludges invalidating the second hypothesis. None of the compounds appeared to have a significant adverse effect on AD operation or stability at longer SRTs and/or lower organic loading rates, although each compound did have slight effects on TS and VS removals. Alum caused digester instability at higher doses and organic loading rates suggesting that the maximum level that can be tolerated is near the 15 mg/L dose and an organic loading rate between 3.15 and 5.40 g VS/L/d.    Practical Implications of Results 5.2Wastewater varies greatly, as does wastewater treatment methods and objectives. This work has demonstrated that both ferric chloride and alum both have practical applications in AD depending on the end objective. If minimization of VSC production to prevent odours, corrosion, and/or the elimination of additional unit processes to remove VSCs are the desired outcome, then ferric chloride would be an ideal candidate. If the minimizations of soluble orthophosphate within AD centrate and/or pathogens in the digestate are priorities, then either ferric chloride or alum could be considered. Magnesium hydroxide did not appear to  112 have any adverse AD operational parameters, however, it did not show any significant beneficial effects on VSC, or dewaterability, and only had modest effects on soluble orthophosphate at the doses tested.  Although some of these results are promising, due to the variability of wastewater, waste sludge, and biological treatments, pilot-scale testing would be required before implementation.   Recommendations for Future Work 5.3The following future work is recommended to help expand upon the findings of and to answer questions raised by this work:  i. Exploration of other trivalent and divalent iron and aluminum compounds like ferrous chloride and polyaluminum chloride. ii. Further examining the potential of alum to decrease pathogens in digested sludge and determination of a dosage and organic loading rate that would not cause inhibition. iii. Metagenomic analysis of dosed versus undosed digesters with alum to see what, if any effects dosing has on the SBRs and methanogen species present. iv. Determination of the effects of thermophilic versus mesophilic anaerobic digestion of VSC production. v. Various studies have demonstrated that the addition of alum will decrease VSC production from dewatered biosolids, while high shear centrifugation increases odour from iron-dosed sludges. Since the opposite holds true for VSCs during digestion, a study monitoring VSCs during and after digestion and dewatering would likely yield some interesting results. vi. Both alum and ferric chloride can chemically bind to, can precipitate out, and/or alter surface charges and interactions of a variety of compounds with varying levels of effectiveness. Therefore, an examination of the effects of these compounds on various pharmaceuticals, personal care products, synthetic hormones, flame-retardants, and other compounds of emerging concern is warranted.  113 vii. The effects of a dual-stage AD on VSCs could produce interesting results. Low pH values due to high VFA concentrations in the first phase could generate favourable conditions for SBRs and would encourage the removal of sulfur due to the volatilization of H2S. Second stage biogas would like contain far smaller concentrations of VSCs and may require much lower ferric chloride additions for VSC control. viii. Exploration of the effects of higher and lower doses of ferric chloride on VSCs and determination of the dosage required to remove nearly all VSCs. ix. Completion of a mass-balance for sulfur entering and exiting the digester and quantification of the sulfur species present in the liquid and solid phases. x. 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Applied and Environmental Microbiology, 35(2), 344–352.   124 Appendices Appendix A: British Columbia Organic Matter Recycling Regulations  OMRR Criteria Class A Biosolids Class B Biosolids Process Criteria Pathogen reduction √ √ Vector attraction reduction √ √ Fecal coliform (MPN/g dry weight) <1,000 <2,000,000 Quality Criteria (µg/g dry weight) Arsenic 75 75 Cadmium 20 20 Chromium Not required 1,060 Cobalt 150 150 Copper Not required 2,200 Lead 500 500 Mercury 5 15 Molybdenum 20 20 Nickel 180 180 Selenium 14 14 Zinc 1,850 1,850 Foreign matter ≤1%* ≤1%* Other Criteria Dry weight (%), organic matter (% dry weight) and C:N ratio  Not required Not required Sampling plan Systematic random, simple random, stratified random Systematic random, simple random, stratified random Type of sample Composite Composite, seven discrete samples for fecal coliform Number of samples (minimum) 3 (each composed of seven subsamples) 3 (each composed of seven subsamples) *Further requirements of no sharp foreign matter that can cause injury.    125 Appendix B: Sample Chemical Oxygen Demand Calibration Curve      126 Appendix C: Example Protein and Humic Acid Calibration Curves      127 Appendix D: Average Total Orthophosphate Levels in mg P/L  SRT Digester Average Change from Feed 20-Day SRT Control (Stage 1) 499 -15.6% 10 mg/L Ferric Chloride 496 -16.0% 10 mg/L Alum 454 -23.2% Control (Stage 2) 362 -38.8% 15 mg/L Ferric Chloride 290 -50.9% 15 mg/L Alum 239 -59.5% 20 mg/L Magnesium Hydroxide 381 -35.5% 60 mg/L Magnesium Hydroxide 313 -46.9% Feed 591 - 12-Day SRT Control (Stage 1) 295 -59.3% 10 mg/L Ferric Chloride 284 -60.8% 10 mg/L Alum 240 -66.9% Control (Stage 2) 572 -21.1% 15 mg/L Ferric Chloride 550 -24.2% 15 mg/L Alum 348 -52.0% 20 mg/L Magnesium Hydroxide 561 -22.6% 60 mg/L Magnesium Hydroxide 401 -44.7% Feed 725 - 7-Day SRT Control (Stage 1) 371 -51.7% 10 mg/L Ferric Chloride 333 -56.7% 10 mg/L Alum 480 -37.7% Control (Stage 2) 830 7.8% 15 mg/L Ferric Chloride 620 -19.4% 15 mg/L Alum  - 20 mg/L Magnesium Hydroxide 816 6.0% 60 mg/L Magnesium Hydroxide 691 -10.3% Feed 770 -     128 Appendix E: Average Protein and Humic Acid Levels  Digester Protein (mg/L) Standard Deviation n Humic Acid (mg/L) Standard Deviation n Stage One Control 4612 2070 6 2488 1580 6 10 mg/L Ferric Chloride 4824 1787 6 2442 1467 6 10 mg/L Alum 4579 1426 6 2260 1418 6 Stage Two Control 5078 682 6 4483 1988 6 15 mg/L Ferric Chloride 5361 621 7 4816 1932 7 15 mg/L Alum 5928 984 7 4234 1894 7 20 mg/L Magnesium Hydroxide 5198 527 7 4665 2122 7 60 mg/L Magnesium Hydroxide 5111 740 7 5019 2156 7 Feed 4936 1788 9 3765 2535 9     129 Appendix F: Sludge Dynamic Viscosity   Digester  Average Dynamic Viscosity (cP)* % Shear Average pH Average % Total Solids (% TS) Stage One Control value 550 2.7 7.480 2.44 n, σ 19, 156 19, n/a 15, 0.204  10, 0.22 10 mg/L Ferric Chloride value 620 3.1 7.319 2.47 n, σ 19, 197 19, n/a 15, 0.274 10, 0.30 10 mg/L Alum value 520 2.6 7.434 2.48 n, σ 19, 125 19, n/a 15, 0.104 10, 0.27 Stage Two Control value 500 2.5 7.428 2.25 n, σ 15, 90 15, n/a 12, 0.301 9, 0.24 15 mg/L Ferric Chloride value 540 2.7 7.431 2.37 n, σ 15, 112 15, n/a 12, 0.247 9, 0.17 15 mg/L Alum value 580 2.9 7.423 2.32 n, σ 15, 219 15, n/a 12, 0.105 9, 0.25 20 mg/L Magnesium Hydroxide value 500 2.5 7.445 2.38 n, σ 14, 120 14, n/a 12, 0.150 8, 0.10 60 mg/L Magnesium Hydroxide value 510 2.6 7.386 2.39 n, σ 13, 112 13, n/a 12, 0.190 8, 0.15 *20-day SRT sludge measured at 25 ± 0.5°C using an LV-5 spindle.    130 Appendix G: Effluent Particle Size Distribution   Digester Average Particle Size Standard Deviation n Dx 10 (μm) Dx 50 (μm) Dx 90 (μm) Dx 10 (μm) Dx 50 (μm) Dx 90 (μm) 20-Day SRT Control (Stage 1) 25.3 67.3 163.5 4.9 13.1 31.6 6 10 mg/L Ferric Chloride 25.4 68.3 186.6 4.1 11.9 47.1 6 10 mg/L Alum 25.0 67.8 209.0 4.1 11.9 77.6 6 Control (Stage 2) 22.4 59.5 148.9 3.4 10.0 23.2 3 15 mg/L Ferric Chloride 20.6 54.7 144.5 1.5 5.6 13.1 3 15 mg/L Alum 21.2 56.7 151.1 3.3 8.9 10.1 3 20 mg/L Magnesium Hydroxide 21.7 57.3 149.9 1.9 6.3 14.0 3 60 mg/L Magnesium Hydroxide 23.7 63.9 174.8 3.8 11.7 37.9 3 12-Day SRT Control (Stage 1) 22.9 60.1 146.6 1.3 2.8 8.5 3 10 mg/L Ferric Chloride 22.2 59.2 313.8 3.7 9.9 248.4 3 10 mg/L Alum 22.0 58.2 146.9 2.3 7.2 19.0 3 Control (Stage 2) 24.2 68.7 239.5 1.6 5.6 84.1 2 15 mg/L Ferric Chloride 24.3 69.3 194.4 1.0 4.1 24.9 2 15 mg/L Alum 21.7 67.7 239.2 - - - 1 20 mg/L Magnesium Hydroxide 23.6 67.4 189.1 0.7 1.8 0.4 2 60 mg/L Magnesium Hydroxide 24.2 68.6 187.9 1.4 4.5 20.6 2 7-Day SRT Control (Stage 1) 23.8 66.4 217.5 4.3 13.1 114.8 3 10 mg/L Ferric Chloride 22.6 62.3 169.0 2.6 7.8 18.3 3 10 mg/L Alum 20.7 61.4 197.7 3.4 11.3 35.1 2 Control (Stage 2) 24.1 64.7 154.9 6.1 15.3 22.1 2 15 mg/L Ferric Chloride 24.1 64.7 158.7 1.0 2.7 9.3 2 15 mg/L Alum - - - - - - - 20 mg/L Magnesium Hydroxide 25.9 68.8 160.0 3.0 6.5 9.8 2 60 mg/L Magnesium Hydroxide 23.5 63.0 151.0 3.5 9.5 19.4 2     131 Appendix H: Average digester biogas composition (%)  Digester CH4 CO2 N2 O2 Control (Stage 1) (n=74) 65.34 ±1.74 29.39 ±2.01 4.18 ±1.79 1.42 ±3.11 10 mg/L Ferric Chloride (n=74) 64.65 3.10 29.14 ±3.10 4.75 ±2.72 1.27 ±4.57 10 mg/L Alum (n=65) 64.99 ±2.95 28.97 ±2.33 4.83 ±0.60 1.25 ±4.60 Control (Stage 2) (n=64) 65.18 ±0.99 29.25 ±1.51 4.46 ±0.32 1.21 ±1.52 15 mg/L Ferric Chloride (n=64) 66.32 ±1.15 29.81 ±2.41 3.07 ±0.61 0.97 ±1.34 15 mg/L Alum (n=52) 66.45 ±1.51 29.34 ±1.60 3.18 ±0.30 1.02 ±0.95 20 mg/L Magnesium Hydroxide (n=64) 66.81 ±1.72 28.79 ±2.15 3.54 ±0.53 0.99 ±1.86 60 mg/L Magnesium Hydroxide (n=64) 67.10 ±1.41 29.78 ±1.45 2.38 ±0.31 0.76 ±0.89     132 Appendix I: Normal Probability Plots of ANOVA Residuals  Figure A.1 - ANOVA normal probability plot of TVSC residuals  Figure A.2 - ANOVA normal probability plot of MPN residuals  1000050000-500099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is TVSCs)3000000200000010000000-1000000-2000000-300000099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is Average MPN) 133  Figure A.3 - ANOVA normal probability plot of CST residuals   Figure A.4 - ANOVA normal probability plot of Zeta potential residuals  10005000-500-100099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is CST)1050-5-1099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is Zeta) 134  Figure A.5 - ANOVA normal probability plot of orthophosphate residuals   Figure A.6 - ANOVA normal probability plot of protein residuals 5002500-250-50099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is Orthophosphate)40003000200010000-1000-2000-3000999590807060504030201051ResidualPercentNormal Probability Plot(response is Protein) 135  Figure A.7 - ANOVA normal probability plot of humic acid residuals   Figure A.8 - ANOVA normal probability plot of COD residuals  3000200010000-1000-2000-3000999590807060504030201051ResidualPercentNormal Probability Plot(response is HA)150001000050000-5000-1000099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is COD) 136  Figure A.9 - ANOVA normal probability plot of dynamic viscosity residuals  (Jackson-Moss & Duncan, 1991) (Du & Parker, 2009) (Vesilind, 1988)(Jackson-Moss & Duncan, 1990) (Yin, Han, Lu, & Wang, 2004) (Ekama, Dold, & Marais, 1986) (Wu, Bishop, & Keener, n.d.; Wu et al., 2001)(United States Environmental Protection Agency, 1996) (Vesilind & Martel, Freezing of Water and Wastewater Sludges, 1990) (Vesilind & Hsu, Limits of sludge dewaterability, 1997) (Mcfeters & Stuart, 1972) (Chin & Lindsay, 1994) (Baur, Benisch, Clark, & Sprick, 2002; Zhou, 2003) (van Huyssteen, 1967) (ASTM International, 2008)  5002500-250-50099.99995908070605040302010510.1ResidualPercentNormal Probability Plot(response is Dynamic Viscosity) 137 

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