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Recalcitrant nutrient removal using heterogeneous struvite precipitation in anaerobic digestion dewatering… Abel-Denee, Marco Mathew 2017

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   RECALCITRANT NUTRIENT REMOVAL USING HETEROGENEOUS STRUVITE PRECIPITATION IN ANAEROBIC DIGESTION DEWATERING CENTRATE by Marco Mathew Abel-Denee B.A.Sc., University of British Columbia, 2015  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)    UNIVERSITY OF BRITISH COLUMBIA (Okanagan)   December 2017  © Marco Abel-Denee, 2017 ii  The following individuals certify that they have read, and recommend to the College of Graduate Studies for acceptance, a thesis/dissertation entitled: RECALCITRANT NUTRIENT REMOVAL USING HETEROGENEOUS STRUVITE PRECIPITATION IN ANAEROBIC DIGESTION DEWATERING CENTRATE  submitted by   Marco Abel-Denee    in partial fulfillment of the requirements of the degree of   Master of Applied Science .   Dr. Cigdem Eskicioglu, School of Engineering  Supervisor Dr. Sumi Siddiqua, School of Engineering  Supervisory Committee Member Dr. Gholamreza Naser, School of Engineering  Supervisory Committee Member Dr. Jianbing Li, University of Northern British Columbia  University Examiner     iii  Abstract    The primary objective of this research was to remove recalcitrant nutrients from anaerobically digested sludge dewatering centrate. Recalcitrant nutrients are defined as small (<0.45 µm) particles or molecules containing bound phosphorus and/or nitrogen and are commonly referred to as non-reactive dissolved phosphorus and dissolved organic nitrogen. Recalcitrant nutrients will persist through wastewater treatment processes and contaminate treated wastewater. The situation is complicated by anaerobic digestion of solid residuals, which load a wastewater treatment facility with more recalcitrant nutrients through internal sidestream recirculation loops. Recalcitrant nutrient treatment may be required for wastewater treatment facilities looking to use anaerobic digestion for energy (biogas) recovery and meet stringent nutrient discharge requirement.     A struvite precipitation methodology is proposed where salt crystals are encouraged to ballast colloidal particles through heterogeneous nucleation. Four biologically unique dewatering centrates were used to test the precipitation methodology on the variety of anaerobic digester configurations that can be expected from municipal wastewater treatment plant. The effect of digestion sludge retention time (2 day, 20 day) and digestion temperature (35°C, 55°C) on the removal of dissolved unreactive phosphorus and nitrogen was monitored. Averaged across all four centrates, the precipitation methodology resulted in dissolved unreactive phosphorus and nitrogen removal of 82.4% and 66.6%, respectively. Antimicrobial contaminants (triclosan, triclocarban) were observed in the precipitates at minute concentrations (<18 ng/g-dry solids). Therefore, heterogeneous struvite nucleation can provide a means of recalcitrant nutrient treatment and reactive nutrient recovery without the micropollutant burden of biosolids land application.    iv  Preface    The research program, including design of experiments, operation of the reactors,  data characterization, and data analysis was done by myself. Cigdem Eskicioglu provided assistance with the design of experiment and data analysis.    Sections of Chapter 3 and 4 have been presented at the 2017 Water Environment Federation (WEF) Nutrient Symposium and published in the conference proceedings. Dr. Cigdem Eskicioglu assisted with data analysis and the production of the proceeding manuscript. Dr. Giampiero Galvagno provided the preliminary phosphorus analysis methods required for phosphorus nutrient spices characterization. Dr. Galvagno also assisted in producing the proceeding manuscript.     Sections of Chapter 3 and 4 have been submitted to the Water Research journal and have been conditionally accepted. Sections of Chapter 3 and 4 have been submitted to the 2018 WEF Residuals and Biosolids conference and have been accepted for podium presentation.    v  Table of Contents Abstract............................................................. ............................................................................................ iii Preface............................................................................................................................ .............................. iv Table of Contents....................................................................................................................... v List of Table................................................................... ......................................................................... ix List of Figures............................................................................................................................................... x List of Abbreviations .................................................................................................................xii Chapter 1: Introduction.......................................................................................................... 1 1.1 Motivation of Research ................................................................................................. 1 1.2 Objective .................................................................................................................... 3 1.3 Novelty of Research ..................................................................................................... 3 1.4 Thesis Organization ...................................................................................................... 3 Chapter 2: Literature Review.................................................................................................. 5 2.1 Wastewater ................................................................................................................. 5 2.1.1 Wastewater Characterization ................................................................................... 5 2.1.1.1 Nutrients........................................................................................................... 7 2.1.1.2 Individual Organic Constituents ........................................................................... 8 2.1.2 Wastewater Treatment ........................................................................................... 9 2.1.2.1 Primary Treatment ........................................................................................... 11 2.1.2.2 Secondary Treatment ........................................................................................ 12 2.1.1.1.1 Nitrification and Denitrification .................................................................... 12 2.1.2.2.1 Biological Phosphorus Removal.................................................................... 13 2.1.2.3 Tertiary Treatment ........................................................................................... 14 2.2 Wastewater Treatment Sludge ...................................................................................... 15 2.2.1 Sludge Characteristics .......................................................................................... 15 2.2.2 End Use for Biosolids .......................................................................................... 16 2.2.3 Anaerobic Digestion ............................................................................................ 17 2.2.3.1 Anaerobic Digestion Control Parameters ............................................................. 19 2.2.3.1.1 Alkalinity, VFAs, and pH............................................................................. 19 2.2.3.1.2 Solid and Hydraulic Residence Time ............................................................. 19 vi  2.2.3.1.3 Organic Loading Rate.................................................................................. 20 2.2.3.1.4 Temperature............................................................................................... 21 2.2.3.2 Anaerobic Digestion Performance Parameters ...................................................... 21 2.2.3.3 Advanced Anaerobic Digestion .......................................................................... 23 2.2.3.4 Biologically Phased Anaerobic Digestion (Acid/Gas-Phase AD) ............................. 23 2.2.3.5 Scaling ........................................................................................................... 25 2.2.3.5.1 Iron Salt Addition ....................................................................................... 26 2.2.3.5.2 WASSTRIP™ Process ................................................................................ 27 2.3 Struvite..................................................................................................................... 29 2.3.1 Struvite Solubility ............................................................................................... 29 2.3.2 Precipitation Mechanisms ..................................................................................... 31 2.3.3 The Phosphorus Cycle ......................................................................................... 33 2.3.4 Suitability of Struvite as Fertilizer .......................................................................... 35 2.3.4.1 Heavy Metals  .................................................................................................. 36 2.3.4.2 Trace Organic Constituents................................................................................ 38 2.4 Summary .................................................................................................................. 40 Chapter 3: Materials and Methods ......................................................................................... 41 3.1 Materials................................................................................................................... 41 3.1.1 Sludge Substrate ................................................................................................. 42 3.1.2 Anaerobic Inoculum ............................................................................................ 42 3.2 Equipment................................................................................................................. 43 3.2.1 AD Vessels ........................................................................................................ 43 3.3 Experiment Layout ..................................................................................................... 44 3.3.1 Centrate Production ............................................................................................. 45 3.4 Experiment Design ..................................................................................................... 46 3.4.1 Anaerobic Digestion Experiment ........................................................................... 46 3.4.2 Struvite Precipitation Experiment  .......................................................................... 49 3.5 Analytical Methods for Sample Characterization ............................................................. 52 3.5.1 Phosphorus......................................................................................................... 52 3.5.2 Nitrogen ............................................................................................................ 53 3.5.3 Colloidal Solids and Zeta Potential......................................................................... 54 vii  3.5.4 Struvite and Centrate Trace Organic Constituents ..................................................... 54 3.5.5 Struvite Precipitate Heavy Metal Content................................................................ 57 3.5.6 Total solids and Volatile Solids ............................................................................. 57 3.5.7 Chemical Oxygen Demand (COD) ......................................................................... 57 3.5.8 Capillary Suction Time ........................................................................................ 58 3.5.9 Alkalinity........................................................................................................... 59 3.5.10 Volatile Fatty Acids............................................................................................. 59 3.5.11 Biogas Volume ................................................................................................... 59 3.5.12 Biogas Composition ............................................................................................ 60 3.5.13 Statistical Analysis of Data ................................................................................... 60 Chapter 4: Results & Discussion ........................................................................................... 61      4.1 Using Heterogeneous Struvite Nucleation to Remove Recalcitrant Nutrients from Anaerobic  Digestion Dewatering Centrate ..................................................................................... 61 4.1.1 Centrate Characterization ..................................................................................... 61 4.1.2 Solids Analysis ................................................................................................... 64 4.1.3 Colloidal Nutrient Removal .................................................................................. 67 4.1.4 Trace Organic Constituents ................................................................................... 69 4.1.5 Precipitate Heavy Metal Contamination .................................................................. 72 4.2 Incorporating Struvite Ballasted Coagulation into Acid/Gas-Phase Anaerobic Digestion ....... 74 4.2.1 Total Solids Removal........................................................................................... 76 4.2.2 Volatile Solids Removal ....................................................................................... 76 4.2.3 Methane Production............................................................................................. 77 4.2.4 Anaerobic Digestion Effluent Dewaterability ........................................................... 78 4.2.5 Anaerobic Digestion Effluent Nutrient Concentrations .............................................. 79 4.2.5.1 Orthophosphate ............................................................................................... 80 4.2.5.2 Anaerobic Digestion Non-reactive Dissolved Phosphorus Concentrations ................. 80 4.2.5.3 Anaerobic Digestion Total Phosphorus Concentrations .......................................... 81 4.2.5.4 Anaerobic Digestion Ammonia Concentrations..................................................... 82 4.2.5.5 Anaerobic Digestion Dissolved Organic Nitrogen Concentrations............................ 83 4.2.5.6 Anaerobic Digestion Total Nitrogen Concentrations .............................................. 84 Chapter 5: Conclusion ......................................................................................................... 85 5.1 Contribution and Significance ...................................................................................... 86 viii  5.2 Future Work .............................................................................................................. 87 5.3 Limitations ................................................................................................................ 87 References................................................................................................................................................... 88 Appendices.................................................................................................................................................. 98 Appendix A - Preliminary Fermentation Experimental Results .................................................... 98 Appendix B - Preliminary Polymer Dose Optimization Results.................................................... 99 Appendix C - Sample Calibration Curves ................................................................................100 Appendix D - Residual Plots for Methane Production................................................................103    ix  List of Tables Table 2.1 - Typical Wastewater Constituent Concentration for Domestic Municipal Wastewater ............ 6 Table 2.2 - Characteristics of Wastewater Sludge .......................................................................... 16 Table 2.3 - Reported Heavy Metal Contamination in Struvite.......................................................... 37 Table 2.4 - Chemical Properties of Pharmaceuticals ...................................................................... 38 Table 3.1 - List of Chemical Materials......................................................................................... 41 Table 3.2 - Sludge Substrate Characterization ............................................................................... 42 Table 3.3 - List of Equipment..................................................................................................... 43 Table 3.4 - Summary of Polymer Dose Used for Dewatering .......................................................... 46 Table 3.5 - Anaerobic Digestion Experimental Design ................................................................... 47 Table 3.6 - Anaerobic Digestion Experiment Data Output .............................................................. 49 Table 3.7 - Struvite Precipitation Experiment Design..................................................................... 50 Table 3.8 - Struvite Experiment Data Output ................................................................................ 52 Table 3.9 - Triclocarban and Triclosan Recovery through Extraction and Analysis  ............................. 56 Table 3.10 - List of Trace Organic Compounds Quantified in Struvite .............................................. 56 Table 4.1 - Centrate Characterization .......................................................................................... 63 Table 4.2 - Precipitate Sample Concentration ............................................................................... 72 Table 4.3 - Heavy Metal Concentrations in Struvite Precipitates ...................................................... 73 Table 4.4 - Effluent Anaerobic Digestion Characterization ............................................................. 75   x  List of Figures Figure 2.1 - Wastewater Nutrient Species (Not to Scale) .................................................................. 8 Figure 2.2 - Wastewater Treatment Facility Process Flow Diagram.................................................. 10 Figure 2.3 - Trend in the Extent of Wastewater Treatment in the U.S. .............................................. 11 Figure 2.4 - Intermediate Steps in the Anaerobic Digestion Process ................................................. 18 Figure 2.5 - Scale Build-up in Piping .......................................................................................... 26 Figure 2.6 - WASSTRIP™ Process Flow Diagram ........................................................................ 27 Figure 2.7 - Struvite Conditional Solubility Product in Anaerobic Digestion Dewatering Centrate ........ 30 Figure 2.8 - Thermodynamic States of Crystal Precipitation............................................................ 32 Figure 2.9 - Struvite Precipitation Mechanism .............................................................................. 33 Figure 2.10 - Phosphorus Cycle Considering Human Impacts ......................................................... 34 Figure 2.11 - Modeled Global Phosphate Ore Extraction Using Known Phosphate Ore Reserves.......... 35 Figure 3.1 - Image of a Bench-Scale Anaerobic Digestion Vessel .................................................... 44 Figure 3.2 - AD Vessel Substrate Flow ........................................................................................ 45 Figure 3.3 - Anaerobic Digestion Experiment Sampling Points ....................................................... 48 Figure 3.4 - Struvite Precipitation Experimental Flow Diagram for Mesophilic and Thermophilic                                   Anaerobic Digestion Operating Temperatures ............................................................. 50 Figure 3.5 - Jar Tester Used for Struvite Precipitation .................................................................... 51 Figure 3.6 - Struvite Experiment Sampling Points ......................................................................... 51 Figure 4.1 - Thermophilic Acid-Phase Anaerobic Digestion Centrate Before and After Struvite                                                                                                                                                                         Precipitation........................................................................................................... 65 Figure 4.2 - TSS Removal ......................................................................................................... 66 Figure 4.3 - Zeta Potential of Untreated and Treated Samples ......................................................... 67 Figure 4.4 - Non-Reactive Dissolved Phosphorus Removal............................................................. 68 Figure 4.5 - Dissolved Organic Nitrogen Removal ........................................................................ 69 Figure 4.6 - Presence of Triclosan and Triclocarban in Precipitate Samples and Centrate Samples ........ 70 Figure 4.7 - Precipitate Sample TOC Concentration Prior to Drying................................................. 71 Figure 4.8 - Heavy Metal Capacity Use of Resultant Struvite Precipitates ......................................... 73 Figure 4.9 - Anaerobic Digestion Total Solids Removal ................................................................. 76 Figure 4.10 - Anaerobic Digestion Volatile Solids Removal ........................................................... 77 Figure 4.11 - Anaerobic Digestion Specific Daily Methane Production............................................. 78 xi  Figure 4.12 - Anaerobic Digestion Effluent Dewaterability............................................................. 79 Figure 4.13 - Orthophosphate Concentration in Anaerobic Digestion Effluents .................................. 80 Figure 4.14 - Non-reactive Dissolved Phosphorus Concentration in Anaerobic Digestion Effluent ........ 81 Figure 4.15 - Total Phosphorus Concentration in Anaerobic Digestion Effluents ................................ 82 Figure 4.16 - Ammonia Concentration in Anaerobic Digestion Effluents .......................................... 83 Figure 4.17 - Dissolved Organic Nitrogen Concentration in Anaerobic Digestion Effluents ................. 83 Figure 4.18 - Total Nitrogen Concentration in Anaerobic Digestion Effluents.................................... 84 xii  List of Abbreviations AD: Anaerobic Digestion BOD: Biochemical Oxygen Demand BNR: Biological Nutrient Removal COD: Chemical Oxygen Demand CST: Capillary Suction Time DON: Dissolved Organic Nitrogen KWWTF: Kelowna Wastewater Treatment Facility NRDP: Non-Reactive Dissolved Phosphorus PAO: Phosphorus Accumulating Organisms PS: Primary Sludge rbCOD: Readily Degradable Chemical Oxygen Demand sCOD: Soluble Carbonaceous Oxygen Demand SSR: Supersaturation Ratio tCOD: Total Carbonaceous Oxygen Demand TCC: Triclocarban TCS: Triclosan TOC: Trace Organic Constituents  TS: Total Solids VFA: Volatile Fatty Acid VS: Volatile Solids WWTF: Wastewater Treatment Facility 1  Chapter 1: Introduction    Human society relies on the consumption of limited natural resources. In particular, the consumption of fossil fuels and nutrients are causing imbalances in the natural systems necessary for life on earth. Atmospheric nitrogen is depleting as nitrogen gas is fixed to ammonia for fertilizer production (Bernhard, 2010). Phosphorus reserves are depleting as demand for fertilize and detergent increases (Elser & Bennett, 2011). Oceans are acidifying as a result of rising atmospheric carbon dioxide concentrations (Orr et al., 2005), and there is a scientific consensus agreeing with the theory that anthropogenic greenhouse emissions are causing global climate change (Cook et al., 2013). As demand on limited natural resources increases wastewater is beginning to be viewed as a resource rather than waste.     At a global scale, wastewater contains a wealth of untapped carbon and nutrient resources. Wastewater treatment facilities (WWTFs) are beginning to develop and incorporate processes that extract carbon and nutrients from wastewater to produce useful products that are competitive on the open market. This research integrates two well-established resource recovery processes, anaerobic digestion (AD) and struvite precipitation.  1.1 Motivation of Research    Anaerobic digestion (AD) of wastewater sludge has the benefit of: reducing waste solids; producing methane-rich biogas; destroying infectious organisms; and reducing the putrescible nature of wastewater solids (Appels et al., 2008). Despite the benefits, a drawback noted by Appeals et al. (2008) is the deterioration of quality in the liquid centrate recycle flow produced during the dewatering process. A centrate produced by AD will recirculate nutrients back into the main processes of a WWTF. This will impact the capability of a WWTF to produce compliant treated wastewater (Murakami et al., 1989). Incorporating AD in to a nutrient-removal WWTF can be met with hesitation.  2     The gap of knowledge in nutrient-removal facilities has been shrouded by the production of dissolved recalcitrant nutrients during AD. Recalcitrant nutrients are defined as small (<0.45 µm) particles or molecules containing bound phosphorus and/or nitrogen. Recalcitrant nutrients tend to persist through treatment processes and contaminate treated wastewater (Neethling & Stensel, 2013; Galvagno et al., 2016). Through AD of solid residuals, the overall recalcitrant nutrient load to the WWTF increases (Galvagno et al., 2014). Recalcitrant nutrient treatment may be required to meet stringent nutrient discharge requirements.     Galvagno et al. (2016) demonstrated biodegradability of dissolved organic nitrogen (DON) through a bioreactor treating AD centrate, but the dissolved non-reactive phosphorus (NRDP) was left unchanged through the biological sidestream treatment process. Arnaldos & Pagilla (2010) have shown that NRDP and DON are treatable using chemical dosing and ultrafiltration, but the technology is operationally costly due to chemical consumption and waste production (Smith et al., 2008; Gu et al., 2013). For AD to be incorporated into a WWTF with nutrient-removal, an operationally feasible method of treating recalcitrant nutrients is desirable.     Recalcitrant nutrient treatment is pursued by applying a struvite precipitation process. Struvite precipitation occurs when ammonia, magnesium, and orthophosphate are abundant in an alkaline solution. The conditions in AD centrate are often favorable for struvite precipitation with magnesium addition (Münch & Barr, 2001). By manipulating the magnesium dosing and other mechanical inputs to a struvite precipitation process, it may be possible to treat recalcitrant nutrients using struvite precipitation.     During struvite precipitation, crystals are initially produced by nucleation. A nucleation mechanism known as heterogeneous nucleation is envisioned as the mechanism for recalcitrant nutrient treatment. Briefly, the struvite crystal is encouraged to nucleate onto suspended particles such as colloidal material containing recalcitrant nutrients. Then, the struvite crystal is grown into a sufficient mass to be settled from solution. The crystal may then be harvested and used as a phosphorus-rich fertilizer. Through the production of a recovered resource and removal of recalcitrant nutrients, the integration of AD and 3  struvite precipitation may be seen as a beneficial system that opens up an opportunity for AD to be incorporated into a nutrient-removal WWTF without the risk of compromising treated wastewater quality.    1.2 Objective     The primary objective of this research is to quantify the removal of recalcitrant nutrients using a heterogeneous nucleation struvite precipitation process. The ancillary objectives are to:  1. assess the quality of struvite produced by monitoring the trance organic constituent and heavy metal concentrations present in the resultant precipitates,  2. study the fate of recalcitrant nutrients through polymer assisted centrifugal dewatering, and 3. monitor the performance and effluent quality of an AD system with an intermediate heterogeneous struvite precipitation step. Parameters of interest include solids removal, gas production, dewaterability, and nutrient concentration.  1.3 Novelty of Research    The novelty of this research is introduced by applying the heterogeneous precipitation mechanism for the purpose of coagulation. Struvite precipitation is a well-established process for the treatment and recovery of reactive nutrients, but the literature has not explored the potential to use struvite precipitation to coagulate and remove colloidal materials such as recalcitrant nutrients. Successful demonstration of this practice will help industry and academic professionals evaluate new methods of precipitating struvite. With this precipitation mechanism, incorporation of AD into a nutrient-removal WWTF may become more feasible and resource-friendly. 1.4 Thesis Organization    The thesis is structured to introduce the main topic of research focus in Chapter 1: Introduction. Chapter 2: Literature Review provides general context to the field of wastewater treatment and describes the theory and parameters governing AD and struvite precipitation in further detail. Chapter 3: Materials and 4  Methodology outlines the testing procedures used to characterize the experiments and provides description of the experiments performed. Chapter 4: Results and Discussion presents the results of the experiments and discusses the results within the context of the literature and the relevance to the practical field of application. Finally, Chapter 5: Conclusion outlines the major findings of the research.          5  Chapter 2: Literature Review 2.1 Wastewater    Wastewater is defined as a domestic water supply after it has been used for its intended purpose. Water use adds constituents and causes the water to become unfit for further use. Constituents found within wastewater may be physical, chemical, and/or biological. For example: organic materials can ferment into nuisance conditions, creating odorous and harmful gasses; nutrients in wastewater facilitate plant growth within aquatic environments, impacting ecosystems; and trace chemicals may be carcinogenic and accumulate into food and water sources. The harmful potential of wastewater constituents justifies the necessity for wastewater treatment.  2.1.1 Wastewater Characterization    When developing wastewater treatment strategies, it is helpful to understand the nature and physical properties of wastewater constituents. Wastewater constituents vary considerably based on raw water quality, cultural practices, industry presence, and the age and type of collection system. For example, a municipality with a large wine industry will have more organic constituents in its wastewater than a municipality with a battery manufacturing industry. Further consideration and discussion of wastewater and wastewater constituents within this thesis will consider typical domestic municipal wastewater. However, large variation exists within the context of domestic wastewater. In watersheds that experience water scarcity, municipalities will often mandate water conservation and per capita water consumption rate decreases. Therefore, the dilution of wastewater constituents will decrease and the produced domestic wastewater will be considered “high strength”. High, medium, and low strength wastewater is defined by populations using approximately 190, 380, and 570 L/capita/day (Metcalf & Eddy, 2014). A characterization of typical wastewater constituent concentrations for various wastewater strength categories is presented in Table 2.1.  6  Table 2.1 - Typical Wastewater Constituent Concentration for Domestic Municipal Wastewater at Low, Medium, and High Strength (Adapted From: Metcalf & Eddy, 2014)         Wastewater Concentration Constituent1   Unit   Low strength  Medium Strength  High Strength Total Solids   mg/L   537  806  1612    Total Dissolved Solids   mg/L   374  560  1121       Fixed Dissolved Solids   mg/L   224  336  672       Volatile Dissolved Solids   mg/L   150  225  449    Total Suspended Solids   mg/L   130  195  389       Fixed Suspended Solids   mg/L   29  43  86       Volatile Suspended Solids   mg/L   101  152  304 Biochemical Oxygen Demand2   mg/L   133  200  400 Chemical Oxygen Demand   mg/L   339  508  1016 Total Nitrogen   mg/L   23  35  69     Organic Nitrogen   mg/L   10  14  29    Ammonia   mg/L   14  20  41    Nitrite   mg/L   0  0  0    Nitrate   mg/L   0  0  0 Total Phosphorus   mg/L   3.7  5.6  11    Organic Phosphorus   mg/L   2.1  3.2  6.3    Inorganic Phosphorus   mg/L   1.6  2.4  4.7 1: Bolded parameters are regulated under the British Columbia Municipal Wastewater Regulation (Municipal Wastewater       Regulation (B.C.), 2012) 2: Values represent 5 day BOD at 25°C     Wastewater constituents are classified as aggregate constituents and individual constituents. Aggregate constituents are a collective number of compounds grouped into one category. For example, “organic compounds” is an example of an aggregate category and may include proteins, humic acids, lipids, fats, and carbohydrates. Two common classifications for organic compounds include: biochemical oxygen demand (BOD) and chemical oxygen demand (COD). The BOD constituent is described as the amount of oxygen needed to reduce the groups of organic constituents into byproducts of biological respiration. The extent of biological degradation will depend on temperature and time, so the BOD value is often associated with a time and temperature value. On the other hand, COD represents the amount of oxygen needed to chemically oxidize the organic material into byproducts of chemical reactions. In raw wastewater, BOD is a fraction of the COD value between 0.3-0.8 because chemical reactions can utilize 7  organic material that is unavailable for biological oxidation, such as hair tissue or finger nails (Metcalf & Eddy, 2014).  2.1.1.1 Nutrients    Nutrients are aggregate constituents and have become a constituent of developing concern. The earliest documentation observing anthropogenic nutrient pollution causing adverse effects on receiving water quality dates to 1890 in a lake near Zürichsee (Switzerland) where quantitative measurements were used to observe the duration and intensity of algal blooms (Schindler & Vallentyne, 2008). Algal blooms deplete oxygen which causes a cascade of consequence in the affected water, including fish kills, reduced water quality, and consequences for drinking water treatment. The cause of eutrophication in freshwater lakes and rivers is predominantly linked to a surplus of phosphorus (Jarvie et al., 2006; Schindler et al., 2016). However, nitrogen has been found to be a limiting nutrient in costal estuaries and marine environments (Howarth & Marino, 2006). To preserve water quality in receiving waters, point sources of nutrient pollution such as wastewater treatment facility (WWTF) treated wastewater have been a target of stringent nitrogen and phosphorus regulations (Bott et al., 2012).    To satisfy the stringent regulations, WWTFs are beginning to employ complex treatment processes that target the different nutrient species present in wastewater. Phosphorus may be categorized into dissolved reactive phosphorus, NRDP and particulate phosphorus; and nitrogen may be categorized into inorganic nitrogen, DON, and particulate nitrogen (Figure 2.1) (Urgun-Demirtas et al., 2008; Arnaldos & Pagilla, 2010; Neethling & Stensel, 2013; Qin et al., 2015; Galvangno et al., 2016). Neethling & Stensel (2013) observed nutrient removal through multiple WWTFs considering the different nutrient species. The conclusion was that inorganic ions (ammonia, nitrate, nitrite, and orthophosphate) and large particulate nitrogen and phosphorus can be efficiently removed by biological, chemical, and physical treatment processes. However, NRDP and DON persist through the WWTF and end up in the treated wastewater (Bott et al., 2012; Neethling & Stensel, 2013; Galvagno et al.; 2016).  8   Figure 2.1 - Wastewater Nutrient Species (Not to Scale)    Although NRDP and DON are nutrient constituents that persist through the biological processes in a WTTF, studies have shown that NRDP and DON contribute to eutrophication in receiving waters. For example, algal bioassays done using treated wastewater have found that NRDP can become 73.7-80% bioavailable (Qin et al., 2015; Li & Brett, 2015). Similarly, studies have demonstrated DON bioavailability from BNR treated wastewater was between 19-68% (Urgun-Demirtas et al., 2008; Sattayatewa et al., 2009). As watershed management practices develop and nutrient discharge regulations become more stringent, it will become necessary to develop treatment strategies that prevent NRDP and DON from being discharged to receiving waters. 2.1.1.2 Individual Organic Constituents     Individual constituents are described by the presence of one specific constituent, like toluene or ibuprofen. Generally, treated wastewater is regulated by aggregate constituents such as BOD. However, emerging concerns towards chemical compounds such as antibiotics, pharmaceuticals, personal care products, flame retardants (commonly referred to as trace organic constituents (TOCs)) are starting to be expressed.     Recent technological advances have allowed analytical laboratories to observe TOCs at concentrations found in natural environments (Ternes & Joss, 2007). A survey of WWTFs across the United States (U.S.) has found a 70% correlation between TOCs found in human and wastewater samples (Venkatesan & Halden, 2014), suggesting that human excretion is a major source of TOC contamination found in a 9  WWTF. Unfortunately, as little as 12% of TOC contamination can be removed during conventional wastewater treatment (Luo et al., 2014). Downstream of the WWTF, residual TOCs may end up in water reservoirs, food sources or in the fertilizers which creates a feedback loop for TOCs to accumulate into human tissues (Venkatesan & Halden, 2014).    The European Union has developed legislation to monitor ten individual constituents or groups of similar individual constituents over the course of four years (European Union, 2013). The list of priority constituents has recently been proposed and include the following substances (Carvalho et al., 2015): Diclofenac, 17-Beta-estradiol (E2), Estrone (E1), 17-Alpha-ethinylestradiol (EE2), Oxadiazon, Methiocarb, 2,6-ditert-butyl-4-methylphenol, Tri-allate, Imidacloprid, Thiacloprid, Thiamethoxam, Clothianidin, Acetamiprid, Erythromycin, Clarithromycin, Azithromycin, 2-Ethylhexyl 4-methoxycinnamateincludes. The priority pollutants consist of pharmaceuticals, antimicrobials, and pesticides. Despite the proposed list of individual constituents, a monitoring stage will be performed before regulations on individual organic constituents may be expected in the European Union. However, this is the first step in implementing rules that force WWTF to employ treatment processes that specifically target TOC treatment.  2.1.2 Wastewater Treatment    The objective of wastewater treatment is to remove harmful constituents from wastewater. The required removal is governed by regional, national and international regulations. In 1972, the first U.S. law regulating the discharge of wastewater was rectified, termed the Federal Water Pollution Control Act (Smith, 1972). Moreover, wastewater discharge may be regulated by international agreements such as the Great Lakes Water Quality Agreement of 1972 between Canada and The United States, which successfully regulated phosphorus discharge into Lake Erie (Dolan, 1993). Wastewater treatment regulations vary regionally, but the underlying objective of constituent removal with the goal of protecting public and environmental health remains consistent. 10     Physical processes, chemical processes, and biological processes are used to remove wastewater constituents. Satisfactory constituent removal is difficult or impossible to achieve using single treatment processes, so a combination of treatment processes are used. A system of combined treatment processes is collectively known as a WWTF. Process flow diagrams are used to illustrate the system of treatment processes within a WWTF, as shown in Figure 2.2.  Figure 2.2 - Wastewater Treatment Facility Process Flow Diagram    Primary treatment, secondary treatment, and tertiary treatment is terminology used to describe the extent of treatment provided by a WWTF. The arbitrary terminology is useful when estimating the quality of treated wastewater being produced by a WWTF. In general, an increase in treatment level results in an increase in treated wastewater quality. As population increases, water bodies that receive treated wastewater experience larger constituent loads. To sustain environmental health, more stringent regulations are being legislated, requiring higher levels of treatment from WWTFs. The implementation of stringent regulations may be interpreted from survey data of WWTFs in the U.S. The data show the number of WWTFs practicing only primary treatment is decreasing and the number practicing tertiary treatment is increasing (U.S. EPA, 1997; U.S. EPA 2008), summarized in Figure 2.3. As an example, the Okanagan Water Basin (British Columbia (B.C.), Canada) has a low total phosphorus (0.25 mg-Phosphorus/L) and total nitrogen (6.0 mg-Nitrogen/L) discharge requirement (Municipal Wastewater 11  Regulation (B.C.), 2012). The objective of obtaining greater treated wastewater quality will continue to be relevant as population growth continues.             *Plant treated wastewater is not discharged into an aquatic environment Figure 2.3 - Trend in the Extent of Wastewater Treatment in the U.S. (Data From: EPA (1997); EPA (2008))  2.1.2.1 Primary Treatment    Primary treatment consists of preliminary screening of bulk materials, sedimentation of heavy inorganic grit, and gravity clarification. Conventional primary treatment is considered a physical treatment process. However, some full-scale applications use chemically enhanced primary treatment in which metal salt, such as ferric chloride, is used to flocculate solids into larger masses that are more responsive to gravity settling. Further, chemically enhanced primary treatment has been shown to remove trace constituents. For instance, lead removal during primary treatment increased from 20% to 95% when dosed 40 mg/L ferric chloride and 0.5 mg/L polymer (Johnson et al., 2008). On the other hand, downstream treatment processes may require readily degradable carbon, so primary treatment is modified to include biological fermentation for volatile fatty acid (VFA) production (Barnard, 1984; Chanona et al., 2006). A schematic showing primary clarification with a fermentation step was shown in Figure 2.2. Depending on the 0%10%20%30%40%50%60%70%80%90%100%Less than Secondary Secondary Greater than Secondary Non-Discharge*12  WWTF needs, primary treatment will consist of physical treatment processes and may also include chemical and/or biological processes. Primary sludge (PS) is a consequential by-product of primary treatment. Significant attention is dedicated to the management of the putrid PS waste stream.   2.1.2.2 Secondary Treatment    Secondary treatment uses oxygen transfer to provide microorganisms with an electron acceptor for biological oxidation. The combined process is considered a chemical and biological treatment. The underlying objective is to ballast dissolved and suspended constituents into a microbial floc that has a larger mass and a greater response to physical separation processes. After substrate consumption and adsorption, biomass from the wastewater slurry is partitioned using physical separation. Physical, chemical, and biological treatment processes are used during secondary treatment.     Methods used to employ biological growth vary considerably. The most common distinction is made based on the growth pattern of microorganisms. Systems maintaining microorganisms in liquid suspension within a slurry of wastewater, microorganisms, and air are termed suspended growth processes. On the other hand, systems using microorganisms growing on a solid medium with intermittent exposure to wastewater and air are called attached growth processes. Historic use of suspended growth processes date back to 1914 (Alleman et al., 1983) where Ardern and Lockett (1914) applied the activated sludge process to a 45 m3/d continuous flow treatment plant in Salford, England. Regardless of suspended or fixed growth orientation, secondary treatment produces secondary sludge with a unique set of characteristics related to its biological origin. The treatment and disposal of secondary sludge is of great concern to a WWTF.  2.1.2.2.1 Nitrification and Denitrification    Biological treatment processes may be used to remove inorganic nitrogen species (Figure 2.1) from wastewater. Nitrification is a two-step process in which inorganic ammonia is biologically oxidized by chemoautotrophic bacteria to nitrite, and nitrite is oxidized further to nitrate by a second type of 13  chemoautotrophic bacteria. When concentrations of ammonia are lower, bacteria in the genus Nitrosospira and Nitrospira dominate the nitrification process. On the other hand, when ammonia concentration is higher, the bacteria in the genus Nitrosomonas Europea and Nitrobacter will grow faster and become the dominant nitrification species (Metcalf and Eddy, 2014). Nitrification may occur in suspended growth and fixed film processes.     Denitrification is a biological process in which nitrate and nitrite are reduced to nitrogen gas. Nitrate is reduced to nitrogen gas by heterotrophic microbes that use nitrate as an electron acceptor during metabolism of organics. To encourage the denitrification process, an anoxic setting is used to limit dissolved oxygen. Anoxic conditions are pursued in order to limit the advantage heterotrophic microorganisms using oxygen have over nitrate reducing heterotrophic microorganisms. Oxygen yields more energy than nitrate during metabolism, allowing oxygen utilizing microorganisms to out-compete denitrifying microorganisms. Inorganic nitrogen in the form of ammonia, nitrite, and nitrate is removed from the wastewater by biological conversion to dissolved nitrogen gas and subsequent vaporization. Several species of bacteria have been identified as denitrifiers, including: Achromobacter, Acinetobacter, Agrobacterium, Alcaligenes, Arthrobacter, Bacillus, Chromobacterium, Corynebacterium, Flavobacertium, Halobacterium, Hypomicrobium, Methanomonas, Moraxella, Neisseria, Paracoccus, Propionibacterium, Pseudomonas, Rhizobium, Rhodopseudomonas, Spirillum, and Vibro (Metcalf and Eddy, 2014). The denitrification process may be considered secondary or tertiary treatment depending on the interpretation of the terminology. In general, denitrification cannot be efficiently incorporated into fixed growth processes and is limited to suspended growth processes.  2.1.2.2.2 Biological Phosphorus Removal    Biological treatment processes may be used to remove dissolved reactive phosphorus (Figure 2.1) from wastewater. Biological phosphorus removal provides phosphorus accumulating organisms (PAOs) with a competitive advantage over other heterotrophic carbon oxidizing bacteria. The competitive advantage is established by introducing wastewater containing readily degradable chemical oxygen demand (rbCOD) 14  and return activated sludge to an anaerobic contact tank. Anaerobic conditions allow POAs to use stored polyphosphates as an energy source for rdCOD uptake and storage (Metcalf & Eddy, 2014). Without an anaerobic contact tank, rapidly growing heterotrophic bacteria compete and deplete PAOs of their food source. During aerobic respiration, PAOs use stored rdCOD for cell growth and dissolved reactive phosphorus uptake. Phosphorus is removed from wastewater by separating and wasting PAOs from the bioreactor. After biological nutrient removal (BNR) (including denitrification), nitrogen and phosphorus content in treated wastewater practicing biological nutrient removal (BNR) (including denitrification) may be expected to be in the range of 3-8 mg-N/L and 1-2 mg-P/L (Metcalf & Eddy, 2014). The biological phosphorus removal process may be considered secondary or tertiary treatment depending on the interpretation of the terminology. 2.1.2.3 Tertiary Treatment    Treated wastewater from secondary treatment, or secondary effluent,  contains residual constituents. Total suspended solids (TSS) and colloidal solids present in secondary effluent can reduce inactivation efficiency of final disinfection processes. Brahmi et al. (2010) demonstrated that disinfection c orrelations used to estimate microbial inactivation kinetics will overcompensate in the presence of 25 mg-TSS/L, a typical value found in secondary effluent (Metcalf & Eddy, 2014). Furthermore, nutrients in secondary effluent practicing carbon removal range between 3-8 mg-P/L and 15-35 mg-N/L, while secondary effluent practicing BNR contains 1-2 mg-P/L and 3-8 mg-N/L (Metcalf & Eddy, 2014). In some situations, further nutrient removal is required to meet discharge regulations. For these reasons, WWTFs will use tertiary treatment processes to remove recalcitrant constituents persisting through secondary treatment. For example, 18 out of 19 WWTFs surveyed used tertiary treatment to meet some of the most stringent nutrient discharge permits in North America (Bott et al., 2012).     Tertiary treatment may use physical, chemical, and/or biological processes to achieve constituent removal. However, most tertiary treatments use physical and chemical processes. The most common tertiary treatment process is filtration in which fluid is passed through a physical membrane that retains 15  and separates particulate material. Using aluminum sulphate as a coagulant and sequential microfiltration (0.22 μm), Arnaldos & Pagilla (2010) removed 69% DON and 72% NRDP from secondary effluent. At the full scale, BNR followed by tertiary disc filtration (1.5 µm) produces tertiary effluent with NRDP and DON concentrations of 0.055 mg/L and 1.57 mg/L, respectively (Galvagno et al., 2016). Although filtration is a common form of tertiary treatment, ion exchange, carbon adsorption, and distillation have been used for advanced wastewater treatment objectives (Kurniawan et al., 2006; Gupta et al., 2009; Metcalf & Eddy, 2014). A commonality with these advanced treatment options is the production of a concentrated waste stream. Filtration requires backwash cycles for membrane recharge and ion exchange requires wash water for resin recharge. Recycling filter backwash wastewater or ion exchange brine to the headworks of the WWTF will risk cycling targeted constituents continuously through the WWTF. The recirculation of constituents has been observed in agricultural water reservoirs (Milstein & Feldlite, 2015).  2.2 Wastewater Treatment Sludge    Wastewater treatment sludge is a multiphase fluid produced by chemical, physical, and biological wastewater treatment processes. PS and thickened waste activated solids (TWAS) are sludge streams originating from primary treatment and secondary treatment, respectively. In the U.S., production of PS and TWAS was estimated to be over 6.5 million metric tons (dry weight) per year in 2004 (Leblanc et al., 2009). The putrescible nature of PS and TWAS demands intensive management strategies to prevent spoiling air and water resources. Management and treatment of sludge can represent 50-60% of the operational expenses of a WWTF (Canales et al., 1994; Appels et al., 2008; WEF, 2012). 2.2.1 Sludge Characteristics    The characteristics of PS and TWAS are summarized in Table 2.2. They contain around 96% liquid by weight which requires dewatering to reduce hauling volumes, increase handling capability, increase the carbon density for incineration and composting purposes, or to satisfy landfilling requirements. 16  Additionally, sludge streams are a significant source of infectious pathogenic organisms (Dudley et al., 1980) and require stabilization for certain end uses. Heavy metals in sludge present a risk to watershed water quality and are a concern when considering the end use. Finally, residual material contains rbCOD that will decompose into odorous compounds and greenhouse gasses in an uncontrolled environment.  Table 2.2 - Characteristics of Wastewater Sludge (Adapted From: Metcalf and Eddy, 2014) Constituent   Unit   PS  TWAS  PS and TWAS Total Solids (TS)   (% w/w)   1-4  0.4-1.2    Volatile Solids   (% of TS)   60-85  60-85    pH       5-8  6.5-8    Alkalinity   mg/L   500-1500  580-1100    Heavy Metals                    Arsenic   mg/kg-TS           1.18-49.2    Cadmium   mg/kg-TS           0.21-11.8    Chromium   mg/kg-TS           6.74-1160    Copper   mg/kg-TS           115-2580    Lead   mg/kg-TS           5.81-450    Mercury   mg/kg-TS           0.17-8.3    Molybdenum    mg/kg-TS           2.51-132    Selenium   mg/kg-TS           1.1-24.7    Zinc   mg/kg-TS           216-8550   Processed and treated PS and TWAS is commonly referred to as biosolids. Biosolids are defined as residual organic material that may be beneficially reused. Biosolids may be considered a resource because of the inherent energy and nutrient content. Depending on source treatment, PS and TWAS can have an energy content of 23000-29000 and 9000-14000 kJ/kg-TS (WEF, 2012); a phosphorus content of 0.008-0.028 and 0.028-0.11 g-P2O5/g-TS (Metcalf & Eddy, 2014); and a nitrogen content of 0.015-0.04 and 0.024-0.05 g-N/g-TS (Metcalf & Eddy, 2014), respectively. Various end uses exist that capitalize on these characteristics.  2.2.2 End Use for Biosolids    Roughly 90% of biosolids in the U.S. were land applied, incinerated, or landfilled in 2004 (NEBRA, 2007). Of the disposal methods, landfilling is a feasible option for smaller municipalities with land and landfill volume to spare. However, as scale increases it becomes difficult to justify the rapid consumption 17  of landfill space (Roy et al., 2011). Furthermore, landfilling (without gas capture) has been modeled to contribute more than twice as much carbon dioxide equivalent greenhouse gas emissions when compared to land application and incineration (Houillon & Jolliet, 2005). Incineration options can capture the bound energy in biosolids and offer options for phosphorus recovery from ash. However, expensive heavy metal removal is typically required if recovered phosphorus is planned to be used as a fertilizer (Xu et al., 2012). Furthermore, incineration requires significant economic commitments when compared to land application, including a much larger initial capital cost and larger net-operational costs (including revenue cash flows) (Lundin et al., 2004; Bolzonella et al., 2007). Land application is an attractive and economic residual management option, but public resistance may restrict this end use (Beecher et al., 2004; Robinson et al., 2012). Land application gains public acceptance after wastewater sludge streams have been stabilized to reduce pathogenic organisms and putrescible material (Beecher et al., 2004).   2.2.3 Anaerobic Digestion    Anaerobic digestion (AD) is an established method used to stabilize organic waste, such as municipal and industrial sludge, animal manure, and food waste. The AD process is a biological treatment in which organic matter is converted to carbon dioxide and methane in a multi-step oxidation process that is simplified and summarized in Figure 2.4. The complex multi-stage digestion process has been shown to contain 456 active microbial species capable of digesting 82% (3800 species) of the microbial population in PS and TWAS (Mei et al., 2016). To begin AD, the hydrolysis step uses extracellular enzymes to reduce particulate matter into soluble organic material capable of assimilation through cell membranes. Fermentation and acetogenesis further reduce soluble organics into acetic acid, hydrogen gas, and carbon dioxide. Finally, acetic acid and hydrogen are converted to methane and carbon dioxide during methanogenesis. Overall, complex organics are oxidized into methane gas and carbon dioxide.  18   Figure 2.4 - Intermediate Steps in the Anaerobic Digestion Process    Within the context of wastewater treatment, PS and TWAS are intermittently pumped into mixed and sealed oxygen-free vessels containing anaerobic microorganisms. As time passes, the microorganisms consume the organic portion of PS and TWAS, producing biogas. Biogas is continuously withdrawn from the digestion tank, and the anaerobically digested material (digestate) is removed intermittently. Using AD on PS and TWAS is advantageous for the following reasons (Appels et al., 2008): solids are destroyed, reducing shipping and handling costs; fuel is generated in the form of methane-rich biogas (60-70% methane); pathogenic organisms are digested; and putrescible material becomes less offensive. Despite the benefits, the authors also note a few drawbacks: partial digestion leaves some infectious organisms and putrescible material in the digestate, slow biological growth rates, poor quality of the sequestered fluid (centrate) after dewatering, production of odor compounds in the gas, and increase heavy metal density due to reduced volume of organic matter. 19  2.2.3.1 Anaerobic Digestion Control Parameters     A select few operational parameters can control the performance of an AD process. These include temperature, solid retention time, hydraulic retention time, pH, alkalinity, VFAs, and substrate characteristics. Although substrate characteristics may be considered uncontrollable, some pretreatment methods can be used to render inert compounds more bioavailable for AD. Furthermore, supplements may be added to the substrate to provide nutrient requirements for biological growth.  2.2.3.1.1 Alkalinity, VFAs, and pH The alkalinity and pH is largely a function of the substrate but can be controlled through sodium bicarbonate, lime, or sodium carbonate addition. Chemical addition is often unnecessary because the AD process produces ammonium bicarbonate through the breakdown of proteins (Metcalf & Eddy, 2014). Alkalinity is required to offset the production of VFAs and the carbonic acid production as carbon dioxide solubilizes into the bulk liquid. The ratio of alkalinity to VFAs is a common parameter used to determine the health of an AD process because it represents the buffering capacity of the reactor content before significant pH alterations occur. An operating range of 0.05-0.25 mg-VFA/mg- alkalinity is recommended for optimal AD performance (Metcalf & Eddy, 2014).  2.2.3.1.2 Solid and Hydraulic Residence Time    Volatile solids destruction and pathogen inactivation are the primary goals of AD. To provide adequate oxidation of organics, sufficient contact time between anaerobic microorganisms and substrate is required. The waste sludge stream flowrate is dependent on the upstream WWTF and is generally uncontrollable; therefore, the tank volume is designed to provide the desired solid retention time (SRT) based on the expected flowrate. An increase in SRT is achieved by recycling active biomass back into the AD tank. Solids recycling can be beneficial to the AD process (Young et al., 2013) but is generally not practiced because of the operation complexity. Without biomass recycle, the hydraulic retention time is 20  equal to the SRT and is calculated by dividing the AD effective (liquid) tank volume by the substrate flowrate.    Washout of microbial cultures occurs when the SRT is less than the time required for microbial mitosis, therefore the rate of microbial production is less than the rate of microbial removal and extinction of the microbial species is inevitable. Acidogenic microbes have a low SRT requirement and methanogenic microbes require approximately 5 days to reproduce, but this value varies based on substrate characteristics (WEF, 2012). To meet the EPA 40 CRF Part 503 regulation standard for “Processes to Significantly Reduce Pathogens” residuals must be subjected to at least a 15-day SRT for mesophilic and thermophilic digestion or 60-day SRT for psychrophilic digestion (U.S. EPA, 1993). Typical SRT values for stable AD are above 10 days (Appels et al., 2008). When SRT is lowered below 10 days, a reduction in VS destruction is noticeable. At a SRT value between 5-6 days, methanogenic bacteria begin to get discharged faster than they can reproduce and the AD process eventually fails (Appels et al., 2008; Metcalf and Eddy, 2014).   2.2.3.1.3 Organic Loading Rate    Interestingly, SRT, defined earlier, is a design parameter that is dependent on sludge flowrate and independent of the sludge characteristics (i.e. solids concentration, organic vs. inorganic fraction of solids) within the flow. The presence of inhibitory compounds, such as ammonia, are related to the concentration of organic compounds in the substrate, since they are formed as a by-product of protein degradation (Chen et al., 2008). For this reason, thick sludge feed containing high concentrations of organics may cause process upsets in comparison to dilute sludge containing less organics but operating at the same SRT. Therefore, some AD systems are designed based on organic loading rate, which is estimated by the daily volatile (organic) solid input per volume of reactor (kg-VS/m3/day). Furthermore, some industrial wastewaters contain less VS and more soluble organic compounds such as sugars, so the loading rate is quantified based on COD rather than VS (kg-COD/ m3/day). A recommended volatile solid 21  loading rate for municipal sludge AD is between 1.9 and 2.5 kg-VS/m3/day, with an upper maximum limit of 3.2 kg-VSS/m3/day (WEF, 2012).  2.2.3.1.4 Temperature    The temperature of the AD process greatly affects the microbial kinetic oxidation rates. Furthermore, temperature changes greater than 1°C/day may cause operational problems such as foaming in the AD process (WEF, 2012), so a control range of 0.5°C/day is recommended to avoid upsets (Appels et al., 2008; WEF, 2012; Metcalf & Eddy, 2014). The mesophilic and thermophilic temperature range are common for AD processes. Mesophilic and thermophilic AD are operated at 30-40°C and 50-60°C, respectively. Although the microbial culture changes between operating temperatures, the principle and oxidation reactions of AD remain similar. Benefits and drawbacks of each operating temperature have been subject of discussion. Thermophilic AD performs better at higher OLR but is less tolerant to substrate characteristic changes, whereas mesophilic AD had a more stable performance regardless of the substrate or OLR (Labatut et al., 2014). Furthermore, thermophilic AD achieves larger methane yield which can offset greater heat energy input and produce a net-energy similar to mesophilic AD (Ziemba & Peccia, 2011).  2.2.3.2 Anaerobic Digestion Performance Parameters     Performance of AD is measured using a variety of parameters. The parameters used to describe the performance of AD will depend on substrate characteristics and the end-use objectives. For example, reporting the COD reduction in high strength winery wastewater may be more appropriate than reporting the solid destruction because biomass growth may increase the solids concentration in wastewaters containing soluble COD. When considering AD of slurries with higher solids, such as municipal or industrial sludge, animal manure, lignocellulosic biomass, it is common practice to report AD performance using VS destruction, which is expressed as a percentage of removal across the AD vessel. Pathogen inactivation is also a parameter of interest in municipal sludge treatment because some of the 22  end uses of digested residuals depend on the concentration of fecal coliforms and salmonella count. However, this parameter may mean little for treatment of industrial wastewater from fruit-juice processing.     In most AD applications, methane production is a performance parameter of interest. Although the expression of this parameter may vary between studies, a common method of reporting methane production is in unit volume of methane at standard temperature and pressure (STP, 0°C, 1 atm), per unit of substrate utilized. The unit of substrate utilized may be expressed using COD or VS, depending on the application. Another common method of reporting methane yield is in unit volume of methane at STP per unit effective volume of AD reactor per day (i.e. L-CH4/L/d). However, since this method is again independent of the OLR, the most common unit in the literature is the specific methane yield which is normalized based on the organic substrate load (L-CH4/g VS fed).     Dewatering is a mechanical and chemical treatment that proceeds AD. Often, the dewaterability of AD digestate is measured because a higher dewatering performance may translate into significant cost savings from reduced hauling and shipping costs of biosolids (e.g. stabilized/ dewatered digestate for potential land application). The published methodologies used to measure and quantify dewaterability are subject to critique from the academic community and new methods have been developed in hopes to more accurately quantify the dewaterability of digestate (To et al., 2016). A common method to measure the dewaterability is to quantify the rate of liquid release from the multiphase sludge using a filter paper. This method is known as capillary suction time (CST) and is represented in seconds normalized by the concentration of solids in the multiphase sludge. Another common method is to measure the amount of time required to filter a predetermined volume from the multiphase sludge, called time to filter.     Odor production represents a great concern to stakeholders and may be considered a performance parameter. Modifying the AD process to reduce the production of hydrogen sulphide, methyl mercaptain and other odorous compounds is of interest for the headspace of the AD process (Abbott & Eskicioglu, 2015) and for the storage, transport and land application of biosolids. 23  2.2.3.3 Advanced Anaerobic Digestion    Simultaneous COD removal and energy recovery make AD an appealing technology for wastewater residuals treatment. However, slow microbial growth kinetics require large tanks for thorough digestion. Hydrolysis is collectively considered the rate-limiting reaction for complex substrates such as TWAS (Ariunbaatar, 2014), so there is a well-established research interest in enhancing the microbial kinetics of AD through pretreatment technologies. The technologies are applied prior to AD to enhance hydrolysis by breaking down large particulate material into smaller easier-to-digest material (Carrère et al., 2010; Ariunbaatar et al., 2014; Neumann et al., 2016). As summarized by the aforementioned review articles, biogas production, pathogen inactivation, and solids destruction improvements over conventional AD can be made using thermal hydrolysis, chemical oxidation, and mechanical disintegration. However, a drawback of these technologies is the dependency on energy or chemical input and only thermal hydrolysis technologies have demonstrated positive energy balances (Cano et al., 2015).  2.2.3.4 Biologically Phased Anaerobic Digestion (Acid/Gas-Phase AD)    One practical method of enhancing the AD process is to separate the biological cultures into two separate tanks, a concept that was first proposed by Pohland & Ghosh (1971). The theory behind this practice is to modify the operating environmental conditions for each microbial culture with the purpose of enriching the microbial culture and enhancing its productivity. The first stage is optimized for hydrolytic-acidogenic microorganisms, or acid-phase AD microorganisms, and the second stage is optimized for methanogenic microorganisms, or gas-phase AD microorganisms. Benefits reported in the literature suggest biologically phased AD, or acid/gas-phase AD, is advantageous over single-phase AD because of: increased stability, higher organic consumption rate, higher methane yield, hydrogen yield, increased solids destruction, and more pathogen inactivation (Demirel & Yenigün, 2002; Ariunbaatar et al., 2014; Neumann et al., 2016). However; the disadvantage is a need for more expensive and complex infrastructure.  24       A variety of experiments have been done to analyze the benefits of acid/gas-phase AD. Despite a larger initial capital cost, the pay-back period for acid/gas-phase AD over single-phase AD can be as low as 2 years for expensive biosolids end-use such as incineration or 6 years for cheaper end-use such as land application or landfilling (Bolzonella et al., 2007). While confirming that acid/gas-phase AD provides more net-energy than single-phase AD, Ziemba & Peccia (2011) investigated the pathogen inactivation through a acid/gas-phase AD system and demonstrated that a 60°C or 70°C acid-phase AD process (operated at 5 and 1 hour SRT, respectively) can inactivate pathogenic organisms by two orders of magnitude more than single-phase AD. Using a mesophilic (35°C) acid-phase AD vessel and a sequenced mesophilic gas-phase AD vessel, the improvement in VS destruction of TWAS was found to be 8.7% greater than conventional AD (Bhattacharya et al., 1998). Furthermore, total COD (tCOD) removal in a 30-day mesophilic biochemical methane potential test increased by 26.2-49.4% after fermenting TWAS in a thermophilic (55°C) acid-phase AD vessel for a 4-day SRT. The conclusion that acid/gas-phase AD creates more bioavailable substrate for the gas-phase AD process has been observed in similar studies (Han & Dague, 1997; Gavala et al., 2003; Bolzonella et al., 2007; Lee et al., 2009). Furthermore, Bolzonella et al. (2007) determined the optimal SRT in a hyper-thermophilic (70°C) acid-phase AD vessel occurred between 2-3 day for maximum methane yield in the sequenced mesophilic and thermophilic gas-phase AD vessels. To reinforce the findings, the soluble COD (sCOD) in the hyper thermophilic acid-phase AD vessel was monitored at different SRTs, in which no significant increase in sCOD was observed after a 2 day SRT. To investigate the acid-phase AD reactor at ambient temperature without the addition of external heat energy, Yuan et al. (2011) tracked the production of sCOD in acid-phase AD vessels at 4°C, 14°C and 24.6°C. The results of this study showed that the optimal SRT was 5, 6, and 9 days for 24.6°C, 14°C and 4°C, respectively. Furthermore, sCOD in the 4°C acid-phase AD vessel reached a maximum concentration of 1000 mg/L after the9 day SRT, and the maximum sCOD in the 24.6°C acid-phase AD vessel was 3000 mg/L. This study demonstrates that as acid-phase AD temperature increases, the optimal SRT decreases and the extent of sCOD production increases. It is unclear if sequenced gas-phase AD vessels will experience performance enhancements when using 25  substrate produced by ambient temperature acid-phase AD vessels. One benefit that is clear is the release or “stripping” of nutrients from TWAS during acid-phase AD, even at lower temperature (Yuan et al., 2011; Bolzonella et al., 2012). Controlled release of nutrients in a low pH environment (acid-phase AD) may be used to prevent scaling in the sequenced gas-phase AD vessel, in addition to the other benefits acid/gas-phase AD provides. 2.2.3.5 Scaling    When sludge streams are subjected to AD, fixed organic phosphorus is solubilized into reactive phosphorus. The concentration of released reactive phosphorus during and after AD is often large enough to cause mineral salt precipitation onto internal solid surfaces such as AD tanks, downstream pipes, and dewatering equipment. Conveyance failures in the piping network and equipment downstream of AD are consistently reported at full-scale WWTFs because of scaling issues, which may be observed in Figure 2.5 (Bhattarai et al., 1989; Horenstein et al., 1990; Mamais et al., 1994; Ohlinger et al., 2000; Duckworth et al., 2016; Roussel & Carliell-Marquet, 2016). One study reported a 150 mm discharge line was reduced to 60 mm in 12 weeks because of scaling (Doyle et al., 2000). The reduced cross-section leads to increased pumping requirements and increased time for fluid transport. To remove scaling, acidic solutions are pumped through the affected pipes and left to soak for at least 24 hours, a process that may not completely remove the scaling and is disruptive to WWTF operations. The costs of removing or preventing mineral salt precipitation is estimated between 160,000-880,000 USD/year considering different management strategies and a 20-year normalization (Esping & Merlo, 2011). Operational concerns become more relevant when POAs are used in secondary treatment (Section 2.1.2.2.2).  26   Figure 2.5 - Left: Struvite Scale Build-up in Heat Exchanger Piping (Adapted From: Esping & Merlo, 2011); Right: Vivianite Scale in Anaerobic Digestion Heating Piping (Adapted From: Duckworth et al., 2016)    Mineral fixation onto AD tanks used to digest sludge containing PAOs is well-documented in the literature (Marti et al., 2008; Pastor et al., 2008; Bolzonella et al., 2012) and two major salt species have been identified as the main sources of scaling, struvite (magnesium ammonium phosphate) and hydroxyapatite (calcium phosphate hydroxide) (Marti et al., 2008). Marti et al. (2010) furthered their research to observe the effect of modifying the residual treatment process flow on phosphorus mass flow. The optimal process flow strategy was to elutriate phosphorus from TWAS during a gravity thickening process. In this configuration, a tank is configured for gravity settling with a small retention time to avoid the production of ammonia through hydrolytic-acidogenic (acid-phase AD) oxidation. The excreted phosphorus can bypass gas-phase AD and phosphorus-reduced residuals are subjected to the benefits of the gas-phase AD treatment. This residual management technique has been patented (Baur, 2009) and is commonly referred to as the WASSTRIP™ process. Other methods such as metal salt addition may be used to reduce struvite scaling in AD systems.   2.2.3.5.1 Iron Salt Addition    The addition of iron salts such as ferric chloride or ferrous chloride to the AD process is used to prevent the generation of odorous gases (Abbott & Eskicioglu, 2015). An additional benefit of iron addition was 27  hypothesized to reduce scaling as struvite within an AD process (Dezham et al., 1988) and was later confirmed through bench scale experimentation; Mamais et al. (1994) observed a reduction in struvite precipitation using a dosing ratio of 2.7 g-Fe/g-P. At this dose, phosphate that normally precipitates as struvite was precipitated as vivianite. Vivinite is a partially soluble salt capable of consuming the reactive P that would otherwise be precipitated as struvite. Despite reduction in struvite, the fundamental issues associated with scaling are still relevant to vivianite and have been reported in the literature (Duckworth et al., 2016; Roussel & Carliell-Marquet, 2016), as seen in Figure 2.5. 2.2.3.5.2 WASSTRIP™ Process    The WASSTRIP™ patent claim is a reduction in nuisance phosphorus precipitation and an increased recovery of P as struvite. Two of the ingredients required for struvite precipitation, phosphorus and magnesium, are excreted from TWAS in a fermentation and thickening step. The supernatant containing phosphorus and magnesium is mixed with the dewatered centrate of the downstream AD process. The mixing of the thickened supernatant and ammonia-rich centrate creates a setting in the WWTF where precipitation of struvite may be controlled and recovered. Figure 2.6 contains a process flow diagram of the WASSTRIP™ process that is bordered in red.  Figure 2.6 - WASSTRIP™ Process Flow Diagram 28     The WASSTRIP™ process offers a methodology to reduce scaling in AD. Despite optimization at the pilot scale, 5% of total phosphorus is estimated to precipitate in the AD process while using the WASSTRIP™ configuration (Schauer et al., 2011). The potential for scaling may be explained by the release mechanism of phosphorus from TWAS containing PAOs (Wild et al., 1997). Two separate phosphorus release mechanisms are identified: rapid release of polyphosphates assimilated into the PAOs energy storage mechanism, and slower hydrolysis of phosphorus bound by particulate organic matter. Phosphorus release from PAO polyphosphate storage and a combination of polyphosphate storage and organic phosphorus hydrolysis was determined to be 100 mg-P/L after 4 hours and 250 mg-P/L after 100 hours, respectively, for a mixed PS and TWAS from a BNR process. (Schauer et al., 2011). The WASSTRIP™ thickening process is operated at a short SRT. Therefore, management of phosphorus released from polyphosphates is feasible, but phosphorus released from organic hydrolysis has scaling propensity. Furthermore, sCOD in the phosphorus-rich thickener supernatant bypasses AD and is recycled to headworks in the sidestream return flow, reducing methane production and creating a feedback loop of sCOD in the WWTF.     An acid-phase AD vessel that uses PS and TWAS as substrate and operates at an SRT sufficient for hydrolysis of particulate phosphorus may be an alternative to the WASSTRIP process. Assuming complete oxidation of particulate phosphorus to reactive phosphorus, there is a perceived risk of scaling within the acid-phase AD vessel. Fortunately, acid-phase AD vessels are designed to operate at a pH less than 5.8 (WEF, 2012); at lower pH values such as 5.8, struvite is much more soluble and much less likely to precipitate as scaling (Abbona et al., 1982). For example, Borgerding (1972) observed the solubility of struvite dropped from 5000 mg/L to 100 mg/L during a pH change from 4.5 to 7.5. Therefore, 4.9 g/L of additional scaling may be expected in a single-phase AD vessel in comparison to an acid-phase AD vessel. Therefore an acid/gas-phase AD system with intermediate phosphorus-recovery as struvite may beneficial because it can prevent internal scaling and provide beneficial enhancements of acid/gas-phase AD as discussed in Section 2.2.3.4. 29  2.3 Struvite     Magnesium ammonium phosphate hexahydrate, commonly referred to as struvite, is a partially soluble mineral salt that precipitates from fluids containing Mg2+, NH4+, and PO43- (magnesium, ammonia, and phosphate). Although the crystallization of struvite has a negative connotation with regards to scaling, the formation of solid struvite signifies the removal of reactive nutrients from wastewater. Additionally, controlled recovery of struvite and removal of nitrogen and phosphorus from wastewater can reduce residual solid production by 49% when compared to chemical precipitation of phosphorus (Woods et al., 1999). Controlled struvite precipitation can be beneficial and has driven commercial innovation. Currently, a variety of proprietary struvite precipitation processes are available, including the Airprex®, Crystalactor®, NuReSys®, Pearl®, Phosnix®, and PHOSPAQ™ process (Metcalf & Eddy, 2014).     The proprietary struvite precipitation processes listed vary considerably in design. AD centrate represents less than 1% of the total WWTF flowrate but amounts to 20% and 30% of the nitrogen and phosphorus load, respectively (Bilyk et al., 2012). The low volume and high nutrient concentration makes AD centrate an ideal candidate for struvite precipitation and many of the precipitation technologies target AD centrate. However, the AirPrex® process is designed to precipitate struvite directly from AD digestate (without dewatering) (Nieminen, 2010). Further a fundamental difference between the precipitation technologies may be realized after considering the chemical properties governing the precipitation of struvite.  2.3.1 Struvite Solubility    Struvite is a partially soluble salt containing Mg2+, NH4+, and PO43- in equal molar ratios. The chemical precipitation equation is presented in Equation 2.1Error! Reference source not found.. As shown in Equation 2.1 hydrogen ions are released during the precipitation of struvite. Therefore, struvite precipitation is thought to occur naturally in sections of high turbulence and continuous mass flow; continuous removal of hydrogen keeps the solution pH from lowering and favours struvite precipitation.  30  𝐌𝐠𝟐+(𝐚𝐪) + 𝐍𝐇𝟒+(𝐚𝐪) + 𝐇𝐧𝐏𝐎𝟒𝐧−𝟑(𝐚𝐪) + 𝟔𝐇𝟐𝐎(𝐥) ←→ 𝐌𝐠𝐍𝐇𝟒𝐏𝐎𝟒●𝟔𝐇𝟐𝐎(𝐬) + 𝐧𝐇+(𝐚𝐪) Equation 2.1    The partitioning between ionic species on the left side and the solid crystal on the right side of Equation 2.1 is described using solubility. A state of equilibrium exists between the solid crystal and the aqueous ions in which the rate of solubilisation is equal to the rate of precipitation. This state of equilibrium is described by the solubility product, Ksp. The solubility product is calculated from the concentrations of ionic activities, an effective measurement of an ions chemical concentration. However, calculating ionic activity is difficult in the presence of other interfering reactions. When the calculation of ionic activities is infeasible, the conditional solubility product, Ps, is a more practical estimate of the equilibrium state for a unique substrate (Ohlinger, 1999; Britton et al., 2005). The Ps value may be calculated as shown in Equation 2.2 based on the molar concentration of ions in a solution. The effect of pH on the solubility of struvite is shown in Figure 2.7 using Ps values. As the pH increases from 6.5 to 9 the Ps value decreases from 0.0015 to 0.0002. The minimum solubility for struvite is expected to be around 10.7 (Le Corre et al., 2009).                                                                           𝐏𝐬 = [𝐌𝐠𝟐+] ∗ [𝐍𝐇𝟒+] ∗ [𝐏𝐎𝟒𝟑−] Equation 2.2  Figure 2.7 - Struvite Conditional Solubility Product in Anaerobic Digestion Dewatering Centrate (Adapted From: Britton et al. (2005)) 31  2.3.2 Precipitation Mechanisms    When the product of ionic species is greater than Ksp or the product of molar concentrations is greater than Ps, the solution is considered saturated and precipitation is expected to occur. Nucleation and crystal growth are two possible precipitation mechanisms. Nucleation is described as the combination of ions that form crystal embryos while crystal growth is a surface integration reaction in which solute ions are incorporated into a crystal lattice (Mullin, 2001). Nucleation precipitation is further categorized into two separate types of nucleation, homogeneous nucleation and heterogeneous nucleation. If nuclei are precipitating onto crystal embryos, the nucleation is considered homogenous, and if the nuclei are precipitating onto a foreign material, the nucleation is considered heterogeneous. The extent of solution saturation governs which crystal precipitation mechanism or combination of crystal precipitation mechanisms occur.      Each form of crystal production requires a different extent of solution saturation. The saturation value, or supersaturation ratio (SSR), is a function of the concentrations of ions in a solution. The SSR is defined as the ratio of the activity of the ionic species of the sample to the equilibrium solubility product (Mullin, 2001; Le Corre et al., 2009), or the SSR may be estimated using the ratio of sample conditional solubility product to equilibrium conditional solubility product, as shown in Equation 2.3 (Britton et al., 2005). In Figure 2.8, the metastable region is displayed between the supersaturation curve and the solubility curve. The metastable region describes the SSR in which crystal growth may occur in the presence of an established crystal lattice, but nucleation will not occur (Mullin, 2001). Without the presence of a crystal lattice, the solution will not precipitate and will remain in a state of saturation. The metastable region is a desirable state for engineering processes targeting large pelleted crystals, as significant crystal growth is required without losing mass to the formation of new crystal embryos. A solution is described as supersaturated when the saturation conditions are extending beyond the metastable zone and into the labile zone. A supersaturated solution has the propensity to nucleate either through homogeneous nucleation or heterogeneous nucleation. The literature suggests that heterogeneous nucleation occurs at 32  lower SSRs than homogeneous nucleation (Mullin, 2001; Mehta and Batstone, 2013; Metcalf & Eddy, 2014).       𝐒𝐒𝐑 = 𝐏𝐬−𝐒𝐚𝐦𝐩𝐥𝐞/𝐏𝐬−𝐄𝐪𝐮𝐢𝐥𝐢𝐛𝐫𝐢𝐮𝐦  Equation 2.3   Figure 2.8 - Thermodynamic States of Crystal Precipitation, Where: A, Solution is in Undersaturated Zone and will not Precipitate; B, Solution is in the Metastable Zone and will Precipitate Through Crystal Growth; C, Solution is in the Labile Zone and will Precipitate by Nucleation or Crystal Growth (Adapted From: Mullin, 2001)     Based on saturation principles, struvite precipitation has been operated using two main crystal production methodologies. A visual description of each type of precipitation mechanism is shown in Figure 2.9. Often, as seen in the bottom row of Figure 2.9, crystallization processes are controlled within the metastable zone to encourage the production of a pure crystal product. In some cases, the operational complexity of substrate recycle and process control is avoided for rapid and convenient crystal production, as seen in the top row.  33   Figure 2.9 - Struvite Precipitation Mechanism (Top Row: Nucleation Precipitation; Bottom Row: Growth Precipitation)     Benefits and drawbacks are apparent for each precipitation strategy. The purity of struvite, presence of trace contaminants, and handling capacity are features that determines the resale value of the struvite. The growth style precipitation condition favors the production of pure struvite. On the other hand, nucleation style crystal production can provide benefits to the WWTF in the form of reduced operational requirements and colloid removal. In both cases the goal of removing and recovering reactive phosphorus from wastewater is achieved.  2.3.3 The Phosphorus Cycle     Phosphorus is a fundamental component of all life. Basic societal functions such as fertilizer production and detergent manufacturing are dependent on phosphorus. As a raw material, phosphorus is extracted from igneous and sedimentary rock as phosphate ore. Phosphorus may end up as sediment at the bottom of lakes and oceans, or it may be continuously recirculated through an agricultural cycle (Figure 2.10). Approximately 34% of all anthropogenic phosphorus consumption is discharged into raw wastewater (Rahman et al., 2014). Recovery and reuse of phosphorus in WWTFs can reduce stress on phosphate ore reserves and mitigate dependence on limited phosphorus resources. 34   Figure 2.10 - Phosphorus Cycle Considering Human Impacts (Adapted From: Cornel & Schaum, 2009)    Approximately 7000 million tons of phosphate ore is economically feasible to mine, and human consumption of phosphate ore amounts to 40 million tons per year (Shu et al. , 2006). Global consumption of phosphate ore is likely to continue to grow by 1.5% each year; however, if demand grows at 3%, phosphate reserves are predicted to deplete by 2065 (Steen, 1998). More recent estimates claim that only 20-35% of phosphate ore reserves will be depleted by 2100, or 40-60% depleted under a worst-case scenario (Vuuren et al., 2010), yet the critical time period of a finite resource is when high-quality easily-accessible reserves have been depleted. At this point production declines while demand continues  to increase, representing peak phosphorus (Cordell et al., 2009). Considering historic phosphorus production and the known amount of global phosphate ore reserves, Cordell et al. (2009) predicted that peak phosphorus could occur by 2033, illustrated in Figure 2.11.  35   Figure 2.11 - Modeled Global Phosphate Ore Extraction (Red Line) Using Known Phosphate Ore Reserves and Historic Extraction (Data Points), (Adapted From: Cordell et al., 2009)    Phosphorus is a limited resource and an irreplaceable nutrient for biological functions. Partial depletion of phosphate ore reserves is a relevant concern to human society because it may impact the global production of food. Therefore, reusing and recovering phosphorus from wastewater will play an important role in reducing dependence on limited phosphate ore reserves. 2.3.4 Suitability of Struvite  as Fertilizer    Application of struvite recovered from wastewater as fertilizer has been practiced at the full scale in Japan since 2000 (Ueno & Fujii, 2001). Under identical phosphorus application rates of 0, 8, 16, 24, 32, and 40 mg-P/kg-soil, bioaccumulation of phosphorus into plant matter was studied using struvite and other conventional phosphate rock sources on ryegrass; ryegrass fertilized with struvite showed slightly greater phosphorus uptake than other conventional fertilizers (Gonzalez Ponce & Garcialopez, 2007). Furthermore, struvite fertilizer performance was similar to organic rock phosphate in alkaline and slightly acidic soils in terms of plant dry matter production and uptake (Massey et al., 2009). The low solubility characteristic of struvite allows for long-term single applications of fertilizer (de-Bashan & Bashan, 2004) 36  2.3.4.1 Heavy Metals    Heavy metals are a common regulatory standard for the determination of fertilizer quality.  Regulatory limits are set to protect public health from chronic poisoning that may be associated with high levels of heavy metal contamination in fertilizer used to grow agriculture products. In almost all cases of struvite production from a variety of municipal sources, heavy metal concentrations have satisfied the regulatory standards of the Canada Fertilizer Act, see Table 2.3. The exception being the nickel contamination that Ronteltap et al. (2007) reported for spiked human urine in a batch precipitation process. 37  Table 2.3 - Reported Heavy Metal Contamination in Struvite   Source Substrate Description Crystalization Description Heavy Metal Concentration (mg/kg of dry weight)1,2       As  Cd  Cu  Hg  Mn  Ni  Pb  Zn Liu et al., 2011 Screened (<0.5 mm) swine wastewater Upflow crystalizer with MgCl2 addition, similar to Phosnix® nr  0  17.9  nr  nr  0  0  95.5 Di laconi et al., 2010 Mature landfill leachate 1 L stirred batch containers using MgCl2 and NaOH nr  <dl  <dl  nr  80  45  <dl  7 Ronteltap et al., 2007 Unhydrolyzed urine spiked with heavy metals 0.5 L stirred batch containers using MgCl2 and NaOH nr  6.9  387  nr  nr  288  45.4  667 Uysal et al., 2010 Liquid portion of AD effluent 0.4 L stirred batch containers using MgCl2 and NaOH nr  nr  nr  4.23  nr  <1.29  nr  13 Latifian & Mattiasson, 2012 Liquid fraction of AD effluent 1 L stirred batch containers with added PO4, MgCl2 and NaOH 0.5  0.02  7.26  0.01  nr  nr  0.3  11 Fattah et al., 2009 Liquid fraction of AD effluent Upflow fluidized bed reactor with MgCl and NaOH, similar to Pearl® 1  0.5  67  nr  145  8.7  nr  29 Canada, Fertilizer Act3     75  20  757  5  nr  181  505  1868 1:  nr, not reported; dl, detection limit 2: bolded numbers represent values exceeding the regulation 3: according to the T-4-93 Safety Guidelines for Fertilizers and Supplements (10/22/2017) 38  2.3.4.2 Trace Organic Constituents     In addition to heavy metals, struvite may contain other TOCs. Compounds such as azithromycin, carbamazepine, diclofenac, ibuprofen, triclosan (TCS), and triclocarban (TCC) are used to treat human patients for microbial infections and symptoms of pain relating to inflammation and neurological disorders. These TOCs have been consistently detected in domestic wastewater (Luo et al., 2014), and may end up in struvite precipitated from wastewaters. The chemical properties of the pharmaceuticals are shown in Table 2.4. The octanol-water partitioning coefficient characterizes the likelihood of a compound to be in the organic or liquid phase. In general, a high affinity for organics is characterized by a log(Kow) value of greater than 4 (Roger, 1996). Therefore, the TOCs listed in Table 2.4 will tend to partition onto organic colloids. These colloids may in turn act as a nucleation site for struvite precipitation and a potential source of TOC contamination in struvite.  Table 2.4 - Chemical Properties of Pharmaceuticals Source Compound Medical Application Molecular Formula Molecular Weight, (g/mole) Octanol-Water Partition, log(KOW) Gros et al., 2006 Azithromycin Antibiotic C38H72N2O12 748.99 4.02 Gros et al., 2006 Carbamazepine Psychiatric C15H12N2O 236.27 2.47 Gros et al., 2006 Diclofenac Anti-inflammation C14H10Cl2NO2 296.16 4.51 Gros et al., 2006 Ibuprofen Anti-inflammation C13H18O2 206.23 3.97 Halden & Paull, 2005 Triclocarban Antibiotic C13H9Cl3N2O 315.59 4.9 Halden & Paull, 2005 Triclosan Antibiotic C12H9Cl3O2 289.55 4.8    Although TOCs are not subjected to fertilizer regulations as of 2017, studies have shown that their presence can impede the productivity of a crop. For example, Herklotz et al. (2010) discovered that TCS, 39  a common disinfectant used in soaps, can inhibit normal plant germination. Using struvite as fertilizer may introduce TOCs into a soil matrix and negatively affect crop production.      On the other hand, vegetable uptake of 118 different TOCs was documented by Sabourin et al. (2012). Using biosolids as a fertilizer, only 8 of the 114 TOCs were detected in vegetables with a maximum concentration of 6.25 ng/g-dry-matter. Analytes were not detected in all three vegetable replications at the relatively low concentrations (Sabourin et al., 2012), suggesting that bioaccumulation of TOCs into plant matter is low when biosolids are used as fertilizer. However, a recent study has shown that the organic matter in biosolids acts as a buffering agent to reduce the uptake of TCS and TCC in carrots by 85-89% by redirection of contaminants to the organic matter naturally present in biosolids (Fu et al., 2016). In the situation where struvite is used as a fertilizer, a large volume of organic matter will not be available to adsorb TOCs. Therefore, plant uptake of TOCs may be substantially larger when using struvite as a fertilizer.    To facilitate discussion, an understanding of the concentration of TOCs in struvite precipitated from wastewater is needed. Previous studies have found negligible amounts of TOCs in struvite (Huang et al., 2016; Butkovskyi et al., 2015). In the study done by Huang et al. (2016), synthetic urine was used to analyze the fate of estrogen through a struvite precipitation process. However, the use of synthetic urine was defended: “Synthetic urine was used to avoid any matrix complexity caused due to the presence of salts, amines and lipids present in real urine”, but as mentioned above some TOCs have a strong affinity for organic materials such as lipids and other trace organic colloids that are excreted into the matrix of natural urine (Hirom et al., 1976). Organic lipids may be a surrogate source of TOC contamination in struvite that may have been overlooked. Furthermore, at the full scale, Butkovskyi et al. (2015) extracted TOCs from pelleted struvite produced from a struvite reactor designed to operate in the metastable saturation zone. The TOC content may have been low due to the growth style struvite precipitation. Butkovskyi et al. (2015) attempted to extract TOCs from pelleted struvite which may have kept internal 40  impurities sheltered from the extraction solution. Further investigation is required to characterize the fate of TOCs in struvite, especially for struvite precipitation processes favouring heterogeneous nucleation. 2.4 Summary    Wastewater treatment is a crucial service that protects human and environmental health. Within the context of nutrient treatment, dissolved colloidal nutrient species (NRDP and DON) present a technological challenge because of the small and biologically inert characteristics. Return flows such as AD dewatering centrate contribute significantly to the colloidal nutrient load within a WWTF. In addition to NRDP and DON, AD centrate also contains a large quantity of dissolved reactive nutrients in a small volumetric flowrate. The hydraulic and nutrient constituent concentrations offer an opportunity to efficiently treat NRDP and DON with reactive nutrients by using a heterogeneous struvite precipitation process. As a consequence of targeted colloid removal, it is unclear what fate heavy metals and TOCs will have when passing through the heterogeneous struvite precipitation process.     Further, the strategy of dissolved nutrient removal through heterogeneous struvite precipitation has potential to be incorporated into an acid/gas-phase AD system. Heterogeneous struvite precipitation may be applied upstream of gas-phase AD vessel with the intention of avoiding scaling in the gas-phase AD vessel and dewatering equipment. However, the performance of the gas-phase AD process post struvite precipitation is uncertain.   41  Chapter 3: Materials and Methods    Raw materials, equipment, and chemical reagents used to complete the experiments are listed in this chapter. In addition, descriptions of the experimental design and procedure are provided.  3.1 Materials    The chemical materials used for this research are tabulated in Table 3.1. Table 3.1 - List of Chemical Materials Materials   Manufacturer  Purity Potassium Nitrate   Sigma-Aldrich  99.99% Glutamic Acid   Acros Organics  99% Potassium Persulphate  Fisher Scientific  ASC Grade Sodium Hydroxide  Fisher Scientific  ASC Grade Ammonium Molybdate  MP Biomedicals  ASC Grade Antimony Potassium Tartate  Fisher Scientific  ASC Grade Ascorbic Acid  Fisher Scientific  ASC Grade Sulphuric Acid  BDH  ASC Grade Potassium Dihydrogen Phosphate  Fisher Scientific  ASC Grade Sodium Tripoly Phosphate  Fisher Scientific  ASC Grade Distilled Water  -  0.0 ppm salt Potassium Sodium Tartrate  Sigma-Aldrich  99% Sodium Citrate, Dihydrate  Fisher Scientific  99% Phenol  Fisher Scientific   Sodium Hypochlorite  Ricca  5% Sodium Nitroferricyanide  Fisher Scientific  99.9% Potassium Acid Phalate  Fisher Scientific  ASC Grade Mercuric Sulphate  Fisher Scientific  ASC Grade Potassium Dichromate  Fisher Scientific  ASC Grade Triclocarban   Sigma-Aldrich   99% Triclocarban-d4   CDN Isotopes   99.2% Triclosan   Sigma-Aldrich   99.1% Propionic Acid  Sigma-Aldrich  99.95% Isobutyric Acid  Sigma-Aldrich  99.5% Butyric Acid  Sigma-Aldrich  99% Acetic Acid  Sigma-Aldrich  99% Acetonitrile  Fisher Scientific  Optima Methanol  Fisher Scientific  Optima 42  Biological materials used in the experiment include waste sludge and inoculum.  3.1.1 Sludge Substrate     Sludge samples were taken from the Kelowna Wastewater Treatment Facility (KWWTF), B.C., Canada. The KWWTF operates primary, secondary, and tertiary treatment processes. Screening and grit removal precede primary gravity clarification. PS is fermented during clarification to elutriate VFAs for downstream BNR and forms fermented PS at the bottom of the clarifier. Primary effluent is blended with return activated sludge in an anaerobic contact tank followed by an extended aeration tank for carbon utilization and nitrification. A final anoxic zone accomplishes denitrification with carbon supply from primary effluent. WAS is thickened using dissolved air floatation (DAF). For this research, TWAS and fermented PS were sampled weekly from their respective sampling ports and blended together at the research facility using a 66:33% volumetric ratio of TWAS to fermented PS to represent KWWTF operation. Characterization of mixed sludge substrate is presented in Table 3.2. Characterization of sludge substrate were conducted bi-weekly.  Table 3.2 - Sludge Substrate Characterization Parameter1  Value2 pH   5.50 (0.36, 7) COD (mg/L)   54,129 (2802, 7) Alkalinity (mg/L)   489 (117, 5) TS (% w/w)   4.26 (0.33, 11) VS (% w/w)   3.58 (0.31, 11) VFA (mg/L)   1,088 (236, 5) 1: COD, chemical oxygen demand; TS, total solids; VS, volatile solids; VFA, volatile fatty acid 2: average value (st.dev., number of replicates) 3.1.2 Anaerobic Inoculum    Mesophilic inoculum was sampled from the effluent of a 7 L New Brunswick automated fermenter. The bench-scale fermenter had been operating using KWWTF mixed sludge for over three years. Originally, the mesophilic inoculum was sourced from the City of Penticton’s (B.C., Canada) mesophilic AD tanks. Thermophilic inoculum was obtained from AD effluent produced from a bench-scale AD experiment 43  operating for 8 months. The experiment used mixed sludge from the Westside Regional Municipal Treatment Facility (B.C., Canada) which operates similarly to the KWWTF. Originally, the thermophilic inoculum for the experiment was sourced from full-scale AD tanks at the Annacis Island Municipal WWTF (B.C., Canada).  3.2 Equipment    A list of the equipment used for data acquisition and experiment management is presented in Table 3.3. Table 3.3 - List of Equipment Equipment Used   Type, Manufacturer Spectrophotometer    Genesys 10, Thermo Electron Corporation Thermotron    S-1.5-3200, Thermotron pH Probe    13-636-XL25, Fisher Scientific Gas Monometer    Custom Built Balance    XS204DR, Mettler Toledo Pipettes    Various, Fisher Scientific Liquid Chromatograph/Mass Spectrometer    ACQUITY, Waters Gas Chromatograph – A    7890A, Agilent Gas Chromatograph – B    7820A, Agilent Capillary Suction Apparatus    440, Fann Muffle Furnace     W-13, Paragon Industries Incubator/Shaker    Innova 44R, New Brunswick Scientific  Centrifuge    Sorvall Lengend XT, Thermo Scientific  3.2.1 AD Vessels    A model AD vessel is presented in Figure 3.1. Vessels were created from Erlenmeyer flasks. The neck of the flask was sealed from the ambient air using a rubber stopper and silicon. Two holes were bored through the rubber stopper and glass tubes were inserted into the holes and sealed with silicon. The lower end of a short glass tube was positioned in the headspace of the vessel and the other end was connected to a Tedlar® bag for gas collection. The second, longer glass tube was positioned so that the lower end reached 50 mm from the flask bottom. Effluent was drawn from this line using plastic syringes. Feed sludge was added through the flask arm using a plastic syringe. Metal clamps were used to seal the effluent and feed lines when not in use.  44   Figure 3.1 - Image of a Bench-Scale Anaerobic Digestion Vessel 3.3 Experiment Layout    The experimental layout began with eight AD vessels. Four AD vessels were operated for mesophilic and four AD vessels for thermophilic temperature, respectively. The types of AD processes operated in each temperature condition are: one single-phase AD vessel, one acid-phase AD vessel, and two gas-phase AD vessels (Figure 3.2). Single-phase AD vessels simulated single-stage conventional AD with acidogenesis and methanogenesis co-occurring in the same vessel. Acid- or gas-phase AD vessels simulated only the acidogenesis or methanogenesis phase of AD (see Figure 2.4), respectively. Control AD vessels for gas-phase AD and single-phase AD were operated in each temperature condition for comparison and analysis of the effect that upstream struvite precipitation has on the performance of gas-phase AD. Digestate from the acid-phase AD and the single-phase AD vessels were used to produce centrate for a struvite precipitation experiment.  45     The AD phase separation of acidogenic and methanogenic microorganisms for acid-phase AD and gas-phase AD has been achieved/controlled by SRT. A 20 day SRT was selected for single-phase AD vessels based on typical values used in the industry (Metcalf and Eddy, 2014), and a 2 day SRT was selected for acid-phase AD vessels based on preliminary nutrient release data. The objective was to select an SRT that produced maximum release of phosphorus and sufficient amount of ammonia for downstream struvite precipitation. The nutrient release data for each operating temperature is plotted in Appendices Appendix A - Preliminary Fermentation Experimental Results.    Figure 3.2 - AD Vessel Substrate Flow 3.3.1 Centrate Production    The effluent sampled from AD vessels was stored at 4°C and dewatered in 0.5 L batch volume. A Thermo Sorvall XT centrifuge was used to extract centrate from AD effluent. The centrifuge was operated at 4700 rpm (4700 x g) for 30 min dewatering cycles. A solution of 2.5% (w/w) polyacrylamide 46  flocculating polymer (Zetag 7553) was hand mixed into AD effluents prior to centrifuging. Mixing was done by gently rotating the dewatering containers until the effluent was observed to flocculate (approximately 1 minute of rotational mixing). Two doses of polymer were applied to each AD effluent. The first dose (low dose) was optimized to produce a dewatered sludge cake with the highest measured TS content. The second dose (high dose) was optimized to produce centrates with minimal TSS content. The polymer dose for each AD effluent was selected during a preliminary optimization experiment in which incremental increases in polymer dose were applied to 50 mL AD effluent aliquots and then dewatered using 4700 x g for 30 min dewatering cycles. The product cake was measured for TS and the product centrate was measured for TSS. Preliminary experiment results are presented in Appendix B - Preliminary Polymer Dose Optimization Results and the optimal dose is summarized in Table 3.4. Table 3.4 - Summary of Polymer Dose Used for Dewatering AD Effluent   Polymer Dose (g/kg-TS)     Low Dose  High Dose Mesophilic acid-phase AD   3.9  4.0 Thermophilic acid-phase AD   4.7  9.4 Mesophilic single-phase AD   11.1  16.0 Thermophilic single-phase AD   18.8  24.4 3.4 Experiment Design    Two designs of experiments were realized using the experimental layout. The first experiment monitored AD performance parameters and the second experiment monitored pollutant fate through a struvite precipitation procedure. 3.4.1 Anaerobic Digestion Experiment    The AD design of experiments is shown in Table 3.5. Sludge substrate used in the experiment was subjected to variability as a result of process changes or seasonal load patterns at the KWWTP. Furthermore, the time required to collect the required data for one run was around three months. Randomization and replication of the experimental runs was unfeasible given the time constraint of the research program and the variable bias of substrate. Instead, each run was operated in parallel and time-47  dependant (not independent) replications in the data from each experimental run were measured during steady-state conditions. Steady-state conditions were defined as less than 10% variation in biogas production and VS removal.  Table 3.5 - Anaerobic Digestion Experimental Design Run   Factor (level)    Anaerobic Digestion Temperature (°C)  Intermediate Struvite Precipitation (Yes/No) 1   35  Yes 2   35  No 3   55  Yes 4   55  No    The main objective of the AD experiment was to determine the effect an upstream precipitation process would have on the acid/gas-phase AD system performance. Considering removal of colloidal recalcitrant pollutants during struvite precipitation, it is important to observe downstream gas-phase AD effluent will not contain similar pollutants and render the precipitation process obsolete. This query was studied at mesophilic and thermophilic operating temperature. The intention is to observe if operating temperature will have an impact on the production of recalcitrant colloidal material during the gas-phase AD process. The sampling points for this experiment are shown in Figure 3.3. 48   Figure 3.3 - Anaerobic Digestion Experiment Sampling Points    The output from this experiment is tabulated in Table 3.6.     49  Table 3.6 - Anaerobic Digestion Experiment Data Output Data Series1   Measurement Period   Frequency Effluent Characteristics            VS   10/05/2016 - 20/12/2016   Bi-Weekly    TS   10/05/2016 - 20/12/2016   Bi-Weekly    COD   01/10/2016 - 20/12/2016   Bi-Weekly    OP   01/10/2016 - 20/12/2016   Bi-Weekly    NRDP   01/10/2016 - 20/12/2016   Bi-Weekly    TP   01/10/2016 - 20/12/2016   Bi-Weekly    NH3+   01/10/2016 - 20/12/2016   Bi-Weekly    DON   01/10/2016 - 20/12/2016   Bi-Weekly    TN   01/10/2016 - 20/12/2016   Bi-Weekly    Alkalinity   01/10/2016 - 20/12/2016   Bi-Weekly    VFA   01/10/2016 - 20/12/2016   Bi-Weekly    pH   01/10/2016 - 20/12/2016   Bi-Weekly    Dewaterability  01/10/2016 - 20/12/2016   Monthly Biogas Characteristics            Volumetric Production   10/05/2016 - 20/12/2016   Daily    Gas Composition   01/10/2016 - 20/12/2016   Bi-Weekly 1: VS, volatile solids; TS, total solids; COD, chemical oxygen demand; OP, orthophosphate; NRDP, non-reactive dissolved phosphorus;  TP, total phosphorus; NH3+, ammonia; DON, dissolved organic nitrogen; TN, total nitrogen; VFA, volatile fatty acid 3.4.2 Struvite Precipitation Experiment    A full factorial experiment was designed to analyze the effect that AD temperature, AD configuration, and polymer dose has on pollutant removal efficiency using struvite crystallization methodology. A summary of the designed experimental runs and the respective factors used in the experiment are presented in Table 3.7 and a flow diagram of the experimental levels is shown in Figure 3.4. Due to the novel methodology, it was important to verify that the process works on most of the popular AD processes used to digest municipal sludge. This meant testing the methodology on acid-phase AD and single-phase AD at mesophilic and thermophilic operating temperature. Furthermore, the effect of centrate TSS content was a predictable effector of the methodology, so polymer dose was used to control the centrate TSS content and was considered an experimental level.   50  Table 3.7 - Struvite Precipitation Experiment Design Run  Factor (level)   Anaerobic Digestion Temperature (°C)  Solid Retention Time of Anaerobic Digestion Vessels (day)  Centrifugal Dewatering Polymer1 Dose (Dose Level)2 1  35  2  High 2  35  2  Low 3  35  20  High 4  35  20  Low 5  55  2  High 6  55  2  Low 7  55  20  High 8  55  20  Low 1: Zetag 7553 used as the flocculating polymer 2: Detailed polymer dose found in Table 3.4  Figure 3.4 - Struvite Precipitation Experimental Flow Diagram for Mesophilic and Thermophilic Anaerobic Digestion Operating Temperatures    Randomization of the experimental runs for the precipitation experiment was not feasible given the dependence on the upstream AD effluent. Instead, precipitation runs were done when enough centrate was 51  available to operate a one liter batch precipitation run for all four AD effluents. This procedure was followed for the low polymer dose level. For the high polymer dose level, effluent was stored in bulk until enough centrate could be produced for four batch precipitation runs from the four AD effluents. A jar test apparatus was used to precipitate struvite (Figure 3.5). The sampling points for the struvite precipitation experiment are shown in Figure 3.6.  Figure 3.5 - Jar Tester Used for Struvite Precipitation  Figure 3.6 - Struvite Experiment Sampling Points    52  The data outputs from the experiment are tabulated in Table 3.8. Table 3.8 - Struvite Experiment Data Output Data Series1   Measurement Period   Frequency  Influent Centrate            OP   10/05/2016 - 20/12/2016   Monthly    NRDP   10/05/2016 - 20/12/2016   Monthly    PolyDP   10/05/2016 - 20/12/2017   Monthly    OrgDP   10/05/2016 - 20/12/2018   Monthly    NH3+   10/05/2016 - 20/12/2019   Monthly    DON   10/05/2016 - 20/12/2020   Monthly    TSS   10/05/2016 - 20/12/2021   Monthly Effluent Centrate           OP   10/05/2016 - 20/12/2021   Monthly    NRDP   10/05/2016 - 20/12/2022   Monthly    PolyDP   10/05/2016 - 20/12/2023   Monthly    OrgDP   10/05/2016 - 20/12/2024   Monthly    NH3+   10/05/2016 - 20/12/2025   Monthly    DON   10/05/2016 - 20/12/2026   Monthly    TSS   10/05/2016 - 20/12/2027   Monthly Struvite Precipitates            TS   10/05/2016 - 20/12/2026   Monthly    Heavy Metals   10/05/2016 - 20/12/2027   Monthly    TOCs   10/05/2016 - 20/12/2028   Monthly 1: OP, orthophosphate; NRDP, Non-reactive dissolved phosphorus; PolyDP, Dissolved polyphosphorus; OrgDP, Organic dissolved phosphorus; NH3+, Ammonia; DON, Dissolved Organic Nitrogen; TSS, Total Suspended Solids; TOC, Trace Organic Constituent 3.5 Analytical Methods for Sample Characterization    The analytical methods used in both experiments are described below.  3.5.1 Phosphorus    Dissolved total phosphorus, dissolved acid-hydrolysable phosphorus and dissolved reactive-orthophosphorus was measured on samples filtered through 0.45 µm nylon filtration membranes followed by dilution with distilled water to match the 0.3 - 5 mg/L test range. Samples containing obtrusive solids were centrifuged to facilitate the filtration step. Dissolved organic phosphorus was digested into ortho-phosphorus using adaptations to the persulphate digestion method (APHA, 4500-P B.5; Galvagno et al., 2016). Dissolved acid-hydrolysable phosphorus digestion was completed using the procedure described in 53  APHA, 4500-P B.2. Undigested samples were measured for reactive-orthophosphorus using 1 mL aliquots of the same sample that was used to measure dissolved total phosphorus and dissolved poly-phosphorus. All sample aliquots were diluted to a final 5 mL volume using distilled water and analyzed using the ascorbic acid method (APHA, 4500-P E) at a light path of 1.0 cm and a wavelength of 660 nm. Dissolved organic phosphorus was interpreted as the difference between the measured dissolved total phosphorus and dissolved poly-phosphorus. Dissolved poly-phosphorus was interpreted as the difference between dissolved acid-hydrolysable phosphorus and reactive-orthophosphorus. All glassware used for phosphorus testing was vigorously cleaned using heated 0.2 N hydrochloric acid solution, reverse osmosis water, and distilled water. Appendix C - Sample Calibration Curves includes sample calibration curves for orthophosphate, total phosphorus, and acid-hydrolysable phosphorus testing.   3.5.2 Nitrogen    Dissolved total nitrogen, nitrite, nitrate, and ammonia were measured on samples filtered through 0.45 µm nylon filtration membranes followed by dilution with distilled water to match the test range. Samples containing high solids were first centrifuged to facilitate the filtration step.     Dissolved total nitrogen was measured using the persulfate digestion method (APHA, 4500-P J). Using this method, 4 mL of sample was autoclaved at 140-170 kPa for 1 hour in the presence of an alkali 0.2 M persulphate solution (from ASC grade K2S2O8). Undigested samples and digested samples were measured for nitrate using a Dionex Ion Chromatograph, Model ICS-2100 equipped with an AS11-HC column (4 x 250 mm), a matching guard column, KOH eluent, and a thermal conductivity detector. Nitrite was measured using the colourimetric method APHA-NO2- B, and ammonia was measured using APHA 4500-NH3 F. DON was interpreted as the difference between dissolved total nitrogen and the sum of dissolved nitrite, nitrate, and ammonia. Particulate organic nitrogen was measured similarly, but without preliminary filtration, and particulate organic nitrogen was interpreted as the difference between total organic nitrogen and dissolved total nitrogen. All glassware used for nitrogen testing was vigorously cleaned using heated 0.2 N hydrochloric acid solution, reverse osmosis water, and distilled water. 54  Appendix C - Sample Calibration Curves includes sample calibration curves for total nitrogen and ammonia testing.  3.5.3 Colloidal Solids and Zeta Potential    Colloidal solids were characterized by the sequenced measurement of TSS and the intensity-size distribution of the remaining colloids in the filtrate. TSS was measured according to the Standard Method procedure 2540D (APHA, 2005) using prewashed glass microfiber filters that have a pore size of 1.5 µm (VWR; Cat. No.: 28333-129). A Malvern Zetasizer (Nano ZS) was used to characterize the colloids in the filtrate of the TSS measurement. The Zetasizer uses dynamic light scattering principle to characterize the size and intensity of colloidal particles, the instrument is calibrated to measure particle sizes between 0.01 µm and 10 µm. Furthermore, the Zetasizer was used to measure the average zeta potential in filtrate. All Zetasizer measurements were done at 25°C. 3.5.4 Struvite and Centrate Trace Organic Constituents     TCS and TCC were measured in both precipitates and post-precipitation treated centrate samples. Blank samples were carried through the sample preparation procedure in replicates of three. One blank replicate was left unchanged to ensure the sample preparation procedure or column carryover did not introduce contamination. The second replicate was spiked with 100 µL of 1.5 mg/mL of isotopically labelled TCC (TCC-d4) as an internal standard, and the third replicate was spiked with 100 µL of 1.5 mg/mL TCC-d4 internal standard and 100 µL of 500 ng/mL of native TOCs in high-purity methanol. Sand that had been preheated to 550°C for 30 minutes was used as the blank for precipitates. The 5 µm filtered fraction of peat moss extract was used as the blank for centrate samples.    The precipitate samples were dried in a Thermotron™ environmental chamber for several days at 50 ± 1.1°C to avoid mass loss from the crystal structure; above a temperature of 50°C, ammonia and moisture has been observed to volatilize from struvite (Ali, 2007). Once at steady mass, samples were crushed using an agate mortar and pestle that was thoroughly cleaned between sample preparations by washing 55  with weak hydrochloric acid, methanol, and acetonitrile. A 0.5 g mass of sample was suspended in 15 mL of acetonitrile (BDH Chemical, ACS grade) within high-density polyethylene containers. At this point, 100 µL of 1.5 mg/mL TCC-d4 solution was added to each sample and 100 µL of 500 ng/mL natural analyte was added to one replicate of each sample. The solution was agitated in a vortex mixer for 30 minutes and then left to settle overnight. The supernatant was transferred to glass cuvettes, submersed in a 55°C water bath, and evaporated to near dryness under nitrogen gas. Finally, the extract was reconstituted in methanol to 1.5 mL and filtered into injection vials through a syringe style 0.22 µm PTEP filter membrane. Samples were then stored in the freezer (-20°C) for batch analysis.      The centrate effluent samples were dosed with internal standard and natural analyte, similar to precipitate samples. Analyte in each sample was concentrated in a conditioned Waters Oasis solid-phase extraction HLB cartridge. After concentration, each cartridge was washed with ultra-pure water and then eluted with high-purity methanol. Methanol extract was evaporated to near dryness under nitrogen gas in a 55°C water bath and then reconstituted to a total volume of 1.5 mL using methanol and filtered through 0.22 µm PTEP filter membranes.    Analyte recovery for each sample matrix was assessed according to Equation 4.1 by using the difference in concentrations (Ternes & Joss, 2007).  Recovery (%) =Cspike − CsampleCadded∗ 100 Equation 3.1    Where Cspike is the concentration in the spiked sample, Csample is the concentration in the natural sample, and Cadded is the concentration dosed to the spiked sample. The recovery of each matrix is documented in Table 3.9.   56  Table 3.9 - Triclocarban and Triclosan Recovery through Extraction and Analysis    Precipitates  Centrate Effluent   Triclosan  Triclocarban  Triclosan  Triclocarban Mesophilic acid-phase AD  45.0%  82.6%  118.4%  106.8% Thermophilic acid-phase AD  74.0%  108.7%  74.9%  100.0% Mesophilic single-phase AD  82.2%  98.7%  47.9%  66.1% Thermophilic single-phase AD  78.4%  107.2%  96.2%  107.5%    After sample preparation, each sample was analysed using a Waters ACQUITY™ ultra-performance liquid chromatograph attached to a triple quadrupole mass spectrometer (UPLC-MS/MS) equipped with an electrospray ionization probe running in negative ionization mode (ESI-). Separation was achieved using a Waters BEH C18 1.7 µm column using a gradient elution with 5 mM ammonium acetate in HPLC grade water and LC-MS grade methanol provided by Sigma-Aldrich and Fisher Scientific respectively. High purity reference standards were used to optimize each analyte’s response at the beginning of the project and to prepare a multi-point calibration curve to quantify each compound (see Appendix C - Sample Calibration Curves). Isotope dilution was used to account for analyte losses during sample preparation and for instrumental error during analysis. Detection of analyte was confirmed by retention time off the column, the mass of the precursor ion, and the subsequent detection of at least one fragment ion. A summary of the chemical supplier, compound purity, retention time, precursor ion mass, and fragment ion mass(es) used for analyte quantification are summarized in Table 3.10.  Table 3.10 - List of Trace Organic Compounds Quantified in Struvite with Respective Quantification Settings Compound  Supplier  Purity  Retention Time (min)  Precursor Mass  Fragment Mass  Cone Voltage (V) Triclocarban  Sigma-Aldrich  99%  2.5  312.9  125.9/159.9  22/31 Triclocarban-d4  CDN Isotopes  99.2%  2.5  317  130/160  22/31 Triclosan  Fluka  99.1%  2.6  286.8  34.9  20 57  3.5.5 Struvite Precipitate Heavy Metal Content     Struvite samples examined for heavy metals were dried, and crushed in a similar similarly to the TOC sample preparation. However, the analytical testing was done in a commercial lab using nitric and hydrochloric acid hot block digestion followed by inductively coupled plasma-mass spectroscopy.   3.5.6 Total solids and Volatile Solids    Solids analysis for AD digestate and AD substrate (influent) samples was done using ceramic crucibles (CoorsTek™ porceline), an analytical balance (Thermo Fisher Scientific, Mettler-Toledo Excellence XA-105), convection oven and a muffle furnace. TS and VS were measured according to the procedure outlined in APHA 2540 B and 2540 E, respectively (APHA, 2005). Before use, crucibles were prepared for analysis by being soaked in a 20% sulfuric acid solution for one day; then scrubbed, rinsed and air dried; heated to 550°C for 30-min; and left in a desiccator to cool to room temperature. Roughly 20 g of sample was added onto crucibles, and then samples were dried at 98°C ± 2°C overnight followed by a 1 hour drying period at 105°C ± 2°C the following morning. The final 550°C burning step was completed for a period of at least 1-hour. All sample masses were recorded after the crucibles had equilibrated to room temperature inside a desiccator. The results for TS and VS analysis were expressed in units of % by weight. The equation used to calculate TS and VS is provided in Equation 3.2 and Equation 3.3.  𝐓𝐨𝐭𝐚𝐥 𝐒𝐨𝐥𝐢𝐝𝐬 (%,𝐠𝐠) = (𝐰𝐞𝐭 𝐦𝐚𝐬𝐬 (𝐠) − 𝐝𝐫𝐲 𝐦𝐚𝐬𝐬 (𝐠)𝐰𝐞𝐭 𝐦𝐚𝐬𝐬 (𝐠)) ×  𝟏𝟎𝟎 Equation 3.2  𝐕𝐨𝐥𝐚𝐭𝐢𝐥𝐞 𝐒𝐨𝐥𝐢𝐝𝐬 (%,𝐠𝐠) = (𝐝𝐫𝐲 𝐦𝐚𝐬𝐬 (𝐠) − 𝐛𝐮𝐫𝐧𝐞𝐝 𝐦𝐚𝐬𝐬 (𝐠)𝐰𝐞𝐭 𝐦𝐚𝐬𝐬 (𝐠)) ×  𝟏𝟎𝟎 Equation 3.3 3.5.7  Chemical Oxygen Demand (COD)     Materials used for COD analysis include 12 mL glass vials, standard solutions for calibration, digestion solution, catalyst solution, an analytical balance (Thermo Fisher Scientific, Mettler-Toledo Excellence XA-105), a spectrophotometer (Thermo Fisher Scientific, GENESYS™ 10S UV-Vis Spectrophotometer), 58  an oven (Thermotron Industries, Thermotron S-1.5C), and miscellaneous glassware used for volumetric measurements. The measurement procedure was performed according to APHA 5220 D (APHA, 2005). Sample preparation started by pipetting an aliquot of effluent or substrate into a 50 mL centrifuge tube that was positioned on an analytical balance. After recording the mass, an adequate amount of dilution water was added to the centrifuge tube to place the diluted COD value of the sample below the upper limit of the calibration. The sample was homogenized (Kinematica™ Polytron™, PT 10-35 GT) for 1 minute at 8000 rpm. Afterwards a 2.5 mL aliquot of the homogenized diluted sample was pipetted into 12 mL glass vials for subsequent digestion. Digestion was done by incubating a mixture of diluted sample with 3 mL of digestion solution and 1.5 mL of  catalyst solution for 3 hours at 150°C ± 0.1°C. Digestion solution contained mercuric sulfate, potassium dichromate and concentrated H2SO4 (>98%). The catalyst solution contained silver sulphate and concentrated H2SO4 (>98%). Once digestion was complete and samples were cooled to room temperature, the sample absorbance was measured at 600 nm wavelength (1 cm light path) and compared to the calibration standards for COD determination. Calibration standards were prepared using potassium hydrogen phthalate (>99.95%, Sigma BioXtra). A sample calibration curve is provided in Appendix C - Sample Calibration Curves. All reagents used were ASC grade or of better quality.  3.5.8 Capillary Suction Time    CST is a commonly referenced method to interpret the dewatering rate of multi-phase fluids. Materials used for CST include chromatography paper (Whatman®, Type 17), and a CST apparatus (Fann Instrument Company, Model 440). The procedure used to measure CST was adapted from APHA 2710 G (APHA, 2005). An exception to the standard method was the reduction of sample volume to 5 mL. This adaptation was justified with the argument that a volumetric measurement of 5 mL using a 5 mL injection syringe is more accurate than 6.4 mL using a 10 mL syringe. The measurement was made by filling a reservoir with 5 mL of sample. Fluid drains from the reservoir through a filter paper underneath by capillary action and static fluid pressure. The flowrate through the filter paper signifies the CST and is 59  measured by a digital timer that is connected to the filter paper. Since TS is an influencing factor of CST, the CST is normalized by TS to allow for comparison between referenced studies. All CST measurements were made on a level surface, in an undrafted environment, and at 22°C.   3.5.9 Alkalinity    Alkalinity was measured using a burette with 0.1 N H2SO4 solution. The procedure was followed according to APHA 2320 B (APHA, 2005). Effluent samples were centrifuged to extract a supernatant that was used for testing. The supernatant was titrated to a pH of 4.6 and the acid volume was used to determine the alkalinity.  3.5.10 Volatile Fatty Acids     The method for VFA analysis was adapted from a method developed by Ackman (1972). Effluent and substrate samples were centrifuged and filtered through 0.2 μm nylon filters. 0.5 mL of sample filtrate was mixed with 0.5 mL internal standard of isobutyric acid and stored at -20°C for batch analysis. In batches, samples were injected into a gas chromatograph (Agilent, 7890A) equipped with a 25-meter column (Agilent, 19091F-112) and a flame ionization detector. Helium was used as a carrier gas at a flow rate of 40 mL/min. A standard containing 2000 mg/L of acetic, propionic, and butyric acid was injected with each batch to test for recovery.  3.5.11 Biogas Volume    Gas samples collected from bench-scale AD vessels were analyzed for volume. Each day, a gas volume from the AD vessels was pumped through a monometer to measure the gas production. Once measured, gas volumes were adjusted to standard temperature and pressure (0°C, 1 atm). Pressure was measured by local airport and the room temperature was measured with a thermometer. The monometer was calibrated by injecting a known volume into the monometer and recording the height response to the injection.  60  3.5.12 Biogas Composition    The methane, carbon dioxide, nitrogen, and oxygen gas in the biogas mixture were quantified using a method developed by Huyssteen (1967). The gas chromatograph (Agilent, 7820) used was equipped with a packed column (Agilent, G3591-8003/80003) and a thermal conductivity detector. 0.7 mL gas volumes were manually injected into the gas chromatograph. Helium was used as the carrier gas at a flow rate of 25 mL/min. The GC was calibrated by injecting a standard with known concentrations of gas. The calibration standard contains 20% carbon dioxide, 7% nitrogen, and 73% methane. 3.5.13 Statistical Analysis of Data    Data from the struvite precipitation experiment contained 3 experimental levels and was considered a full factorial experiment. The levels considered were: AD operating temperature, AD SRT, and polymer dosage. To assess the effect of each experimental level, Minitab 18 statistical software was used for run multi-factor analysis of variance. In the analysis of variance, experimental levels were assumed to have linear effect on experimental outputs.      Data from the AD experiment contained two sets of two experimental levels. Two controls, conventional single-phase AD and conventional acid/gas-phase AD were used in the analysis of the acid/gas-phase AD with struvite system. Under both control regimes, temperature and struvite precipitation were used as the experimental levels. Pair-wise comparisons were done using paired t-tests for all experimental levels in the AD experiment.      The residual plots for methane production data is shown in Appendix D - Residual Plots for Methane Production. The data appear to be randomly distributed.   61  Chapter 4: Results & Discussion    The results and discussion are separated into two subsections to present each of the experimental designs identified in Chapter 3. The struvite precipitation experiment is presented first and the AD experiment is presented second. The AD experiment is presented second because an understanding of the results from the struvite experiment is crucial in the discussion of the AD experiment.  4.1 Using Heterogeneous Struvite Nucleation to Remove Recalcitrant Nutrients from Anaerobic Digestion Dewatering Centrate     The primary objective of this experiment was to remove recalcitrant nutrients from anaerobically digested sludge dewatering centrate. A struvite precipitation methodology is proposed where salt crystals  (struvite) are encouraged to ballast colloidal particles through heterogeneous nucleation. The secondary objective was to assess presence of TOCs in precipitates. Four biologically unique dewatering centrates were used to test the precipitation methodology on the variety of AD configurations that can be expected from a municipal wastewater treatment plant. The effect of SRT (2 day, 20 day) and AD temperature (35°C, 55°C) on the removal of NRDP and DON was monitored. Averaged across all four centrates, the precipitation methodology resulted in NRDP and DON removal of 82.4% and 66.6%, respectively. Antimicrobial contaminants (TCS, TCC) were observed in the precipitates at minute concentrations (<18 ng/g-dry solids). Therefore, heterogeneous struvite nucleation can provide a means of recalcitrant nutrient treatment and reactive nutrient recovery without the TOC burden of biosolids land application.    4.1.1 Centrate Characterization    Centrate characterization is presented in Table 4.1. Comparison among centrates shows significantly higher OP concentrations in acid-phase AD centrate. Further, mesophilic AD has a higher OP concentration then the respective thermophilic acid-phase AD and single-phase AD centrates. This may be explained by the uncontrolled precipitation of struvite and calcium phosphate within the AD vessel 62  (Marti et al., 2008). Aage et al. (1997) determined that the struvite solubility increased from 1.6 × 0-14 to 3.4 × 10-14 between a temperature of 30°C and 50°C, and Borgerding (1972) observed similar solubility trends with regards to temperature. Furthermore, Borgerding (1972) reported that struvite solubility in a solution at a pH of 4.5 was greater than 5000 mg/L which rapidly decreased to 100 mg/L at a pH change of 7.5. The temperature and pH solubility trends found in the literature are supportive of the OP data observed in the centrate characterization. From a scaling prevention and struvite recovery perspective, the mesophilic acid-phase AD effluent offers the most optimal centrate to perform struvite precipitation on.     Comparison of the high and low polymer dose centrates shows a reduction in NRDP and a production of DON coincides with an increasing polymer dose. This is true for all four experimental centrates. It is presumed that the increased polymer dose leads to an increase in colloidal solids capture rate and a reduction in NRDP; this is reflected by the TSS measurements in Table 4.1. However, the production of DON is stipulated to be a by-product of the dewatering polymer. The dewatering polymer, Zetag ® 7553, is marketed as a polyacrylamide flocculent. One of the functional groups in polyacrylamide flocculants are nitrogen-containing amines. Through centrifugal dewatering, amines and fractions of the polymer backbone may be sheared and solubilized into solution which contribute to higher DON concentrations. The comparison between high and low polymer dosed centrates suggests that NRDP will be reduced and DON will be increased by increasing the polyacrylamide polymer dosing rate.  63  Table 4.1 - Centrate Characterization¹     Low Polymer Dose  High Polymer Dose     M-Acid  T-Acid  M3-Acid/Gas  T3-Acid/Gas  M-Acid  T-Acid  M-Acid/Gas  T-Acid/Gas Total Suspended Solids   125  222  118  289  99  117  21  76   ± 27²  ± 35  ± 59  ± 74  ± 11  ± 7  ± 1  ± 9                   Non-reactive Dissolved Phosphorus  25.6  27.7  10.3  14.9  8.7  9.8  3.2  7.3  ± 6.8  ± 6.7  ± 5.5  ± 3.0  ± 1.1  ± 3.1  ± 2.2  ± 3.0                  Dissolved Organic Phosphorus   15.1  13.3  5.6  8.4  4.4  6.4  1.5  4.7   ± 7.9  ± 6.1  ± 3.4  ± 3.3  ± 1.1  ± 3.1  ± 1.5  ± 4.0                   Dissolved Polyphosphates   8.4  11.3  4.5  5.5  4.6  3.5  0.6  2.0   ± 4.6  ± 5.5  ± 2.8  ± 2.6  ± 1.5  ± 2.9  ± 0.1  ± 1.6                   Orthophosphate   842.0  764.1  511.0  457.1  807.5  771.3  494.9  449.7   ± 38.5  ± 97.4  ± 12.3  ± 36.1  ± 6.6  ± 5.3  ± 3.6  ± 1.6                   Dissolved Organic Nitrogen   63.9  98.4  37.3  69.9  114.9  98.0  116.9  71.5   ± 21.0  ± 29.7  ± 5.3  ± 26.0  ± 14.3  ± 11.9  ± 14.6  ± 12.7                   Ammonia   532.3  948.3  727.8  1120.3  685.4  1164.1  954.1  1206.9   ± 48.1  ± 152.2  ± 113.6  ± 65.6  ± 9.4  ± 10.1  ± 15.1  ± 16.9 1: all values in mg/L 2: reported error is one standard deviation (n = 4) 3: T: thermophilic, M: mesophilic  64        One final remark on the centrate characterization is the comparison of the NRDP concentration in acid-phase AD and single-phase AD. Higher NRDP concentration in the acid-phase AD centrates are consistent with the concept that hydrolysis reactions are expected to solubilize particulate solids (Barker & Stuckey, 1999); however, NRDP concentration in the single-phase AD centrate is lower and suggests that there is a removal mechanism during methanogenesis (Table 4.1). Kuo and Parkin (1996) have demonstrated that anaerobic biological growth (SRT = 40 day) produces soluble microbial products that chelated 44 mg/L of nickel from solution. It is possible that as anaerobic SRT increases from 2 day (acid-phase AD) to 20 day (single-phase AD) soluble microbial products are produced that chelate NRDP from solution. However, the single-phase AD effluent is dosed with significantly more polymer than acid-phase AD, which may be the contributing factor to lower NRDP concentration.  4.1.2 Solids Analysis    A visual example of turbidity removal is shown in Figure 4.1. The image shows the same thermophilic acid-phase AD centrate prior to precipitation and 5 minutes after the precipitation procedure (right). The reduction in turbidity is clear because the baseboard in the fumehood is visible in the treated centrate. Turbidity reduction is an indication that light-blocking solid particles are being removed.  65   Figure 4.1 - Thermophilic Acid-Phase Anaerobic Digestion Centrate Before (Left) and After (Right) Struvite Precipitation    Analysis of TSS on treated and untreated centrate samples indicates that the struvite precipitation methodology can be used as a solids removal strategy. Figure 4.2 shows paired TSS concentrations of each centrate before and after precipitation. It appears that the greatest reduction in TSS occurs at an initial TSS concentration of around 0.08 and 0.15 g/L, and a reduction below 0.02 g-TSS/L was not achieved in this experiment. TSS data can be an indication of removal of nutrients bound to solid sizes greater than 1.45 µm.    66   Figure 4.2 - TSS Removal through Precipitation on Centrates with a) Low Polymer Dose, and b) High Polymer Dose (n = 4)    One mechanism for removal of colloidal particles is surface charge neutralization. Zeta potential was measured on treated and untreated centrate in order to confirm the mechanism of solids removal was heterogeneous nucleation and settling rather than surface charge neutralization (Figure 4.3). In most samples the zeta potential became stronger (e.i., deviated further from neutral). However, in the only case where the zeta potential became weaker through precipitation, the difference was found to be statistically insignificant using a paired t-test (P = 0.184 > 0.05). The zeta potential data help reinforce the hypothesis that heterogenious nucleation is mechanism causing solids removal.  0.000.050.100.150.200.250.30TSS (mg/L) TSS Untreated TSS Treateda) 0.000.050.100.150.200.250.30TSS (mg/L) TSS Untreated TSS Treatedb) 67   Figure 4.3 - Zeta Potential of Untreated and Treated Samples (n = 4) 4.1.3 Colloidal Nutrient Removal    The removal of NRDP from AD centrate using heterogeneous nucleation was successfully demonstrated through nutrient testing on the untreated and treated centrates (Figure 4.4). By comparing untreated and treated centrate concentrations, average NRDP removal was found to be 81.6% for 32 batch runs (± 8.58%). Speciation of NRDP showed that the PolyDP fraction displayed a reproducible removal of 92.8% (± 5.6%) while OrgDP removal displayed lower average removal of 74.4% with considerable variability (± 12.4%). A linear analysis of variance between the factors involved in the experiment revealed that the AD operating temperature (P = 0.094 > 0.05) and SRT (P = 0.356 > 0.05) did not significantly influence NRDP removal. On the other hand, polymer dose (P = 0.004 < 0.05) had a significant influence on the NRDP removal; lower polymer dose yielded a greater NRDP removal. As summarized in the discussion of Table 4.1, untreated centrate dewatered with higher polymer dose experiences a reduction in NRDP in comparison to a lower polymer dose. This may explain the significant effect of polymer dose, as a reduction in untreated centrate NRDP will reduce the magnitude of NRDP removal during struvite precipitation. In general, the precipitation methodology can be applied to any of the four centrates used in this study without sacrificing NRDP removal efficiency, but the upstream dewatering polymer dose optimization will affect the NRDP removal performance. -18-16-14-12-10-8-6-4-20Meso Acid-PhaseADThermo Acid-PhaseADMeso Acid/Gas-Phase ADThermo Acid/Gas-Phase ADZeta Potential (mV) Zeta Untreated Zeta Treated68   Figure 4.4 - Non-Reactive Dissolved Nitrogen Removal through Precipitation on Centrates with a) Low Polymer Dose, and b) High Polymer Dose (n = 4)     The robust removal of PolyDP is assumed to be because of hydrolysis reactions that are catalyzed by the abundance of magnesium ions. Divalent cations such as magnesium can cleave phosphate from a polyphosphate molecule through a chelation mechanism (Rashchi & Finch, 2000). Although the PolyDP removal is consistent for the single-phase AD centrates between the high and low polymer dose, the PolyDP removal occurring in acid-phase AD centrates is affected by polymer dose. It is unclear why the acid-phase AD centrate is affected by polymer dose more than the single-phase AD centrates and further research is required to investigate this topic.      Removal of DON was also demonstrated using the precipitation methodology (Figure 4.5). The average removal efficiency was 65.7% (± 20.1%). In comparison to NRDP, DON experienced consistently lower removal rates through the struvite precipitation process. This observation is assumed to be caused by the hydrophilic and hydrophobic characteristics of NRDP and DON. Qin et al. (2015) characterized the hydrophobic properties of final effluent NRDP and DON using resin separation and found that DON and NRDP was 72.2% and 19.3% hydrophilic, respectively. This means DON, as a collective molecular 0102030405060708090100Removal Efficiency (%) DOrgP PolyDP NRDPa) 0102030405060708090100Removal Efficiency (%) OrgDP PolyDP NRDPb) 69  species, exists more as a solute in solution and less as dissolved colloidal material. NRDP, on the other hand is more likely to exist as dissolved colloidal material. Therefore, the greater removal of NRDP is expected considering that the theoretical removal mechanism used is heterogeneous struvite precipitation. The SRT (P = 0.166) was not found to be significantly effecting the DON removal. However, temperature (P = 0.001) and polymer dose (P = 0.049) did have a significant effect on DON removal. In a study done by Ahuja et al. (2015), the temperature of thermal hydrolysis prior to AD was found to influence the hydrophilic fraction of DON in the centrate sidestream. It is presumed that the temperature of AD will similarly impact the fraction of hydrophilic and hydrophobic DON and subsequently influence the removal rates observed during heterogeneous struvite precipitation. A definitive characterization of the hydrophobic and hydrophilic fraction of DON in mesophilic and thermophilic AD vessels is required to verify this assumption.   Figure 4.5 - Dissolved Organic Nitrogen Removal through Precipitation on Centrates with a) Low Polymer Dose, and b) High Polymer Dose (n = 4) 4.1.4 Trace Organic Constituents     Trace organic constituents (TOCs) were measured through the high polymer dose experiment for precipitate and treated centrate samples. The measured concentrations of TCS and TCC are presented in 020406080100Removal Efficiency (%) a) 020406080100Removal Efficiency (%) b) 70  Figure 4.6. The magnitude of concentration for TCS and TCC was similar within the precipitate matrix (Figure 4.6a). However, TCS concentration in the treated centrate samples was an order of magnitude greater than the TCC concentration (Figure 4.6b). TCS and TCC concentrations were higher in the acid-phase AD treated centrate samples. It is theorized that the antimicrobials which were originally part of the sludge structure were released into the liquid (soluble) phase as a result of hydrolysis, but did not have sufficient retention time to undergo biodegradation in acid-phase AD vessels.  Figure 4.6 - Presence of Triclosan and Triclocarban in a) Precipitate Samples (n = 5) and b) Treated Centrate Samples (n = 2)    There are multiple methods of handling struvite post-precipitation. Two common methods to remove water from struvite slurry include thermal drying and dewatering/thickening, although chemical drying has also become an established method (Bowers, 2016). At the full-scale, the fate of TOC contamination will change depending on the method of water removal. The TOC contamination found in Figure 4.6a is representative of struvite TOC contamination using thermal drying techniques. However, if interstitial 05101520253035Concentration (ng/g-dry) Triclosan Triclocarbana) 01002003004005006007008009001000Concentration (ng/L) Triclosan Triclocarbanb) 71  fluid is removed via centrifugation, TOCs are expected to be removed. Using the average TOC concentrations in precipitates (Figure 4.6a), treated centrate (Figure 4.6b), and TS data (Table 4.2), a prediction of the TOC contamination that would be found in struvite precipitates that are dewatered prior to drying may be reported according to Equation 4.1 where C denotes the concentrate of TCS or TCC and %TS represents the solids concentration of the struvite precipitates prior to drying.  𝐂𝐩𝐫𝐞𝐜𝐢𝐩𝐢𝐭𝐚𝐭𝐞𝐬−𝐝𝐞𝐰𝐚𝐭𝐞𝐫𝐞𝐝 = 𝐂𝐩𝐫𝐞𝐜𝐢𝐩𝐢𝐭𝐚𝐭𝐞𝐬−𝐝𝐫𝐢𝐞𝐝 − 𝐂𝐭𝐫𝐞𝐚𝐭𝐞𝐝 𝐜𝐞𝐧𝐭𝐫𝐚𝐭𝐞 ∗ (𝟏 − %𝐓𝐒𝐬𝐭𝐫𝐮𝐯𝐢𝐭𝐞 𝐬𝐥𝐮𝐫𝐫𝐲) Equation 4.1    The calculation yields concentrations of TCS and TCC between -2 and 18 ng/g-dry (Figure 4.7). The negative TOC contamination in the mesophilic single-phase AD sample is presumed to be a result of the low TSS removal and small initial TSS concentration in the centrate sample (Figure 4.2b). In comparison, TCS and TCC can be found in biosolids being land applied to food crops in concentrations of 9280 and 7060 ng/g-dry for TCC and TCS, respectively (Cha & Cupples, 2009). Therefore, mass flux of TOC contamination to agricultural lands can be reduced by applying struvite fertilizer instead of biosolids and dispose biosolids via different methods (e.g. incineration).  Figure 4.7 - Precipitate Sample TOC Concentration Prior to Drying, Interpreted from a Mass Balance of Averaged Values   -505101520Meso Acid-Phase AD Thermo Acid-PhaseADMeso Acid/Gas-Phase ADThermo Acid/Gas-Phase ADConcentration (ng/g-dry) Triclosan Triclocarban72  Table 4.2 - Precipitate Sample Concentration1 Sample  TS (% by weight) Mesophilic acid-phase AD 1  2.70% Mesophilic acid-phase AD 2 +Spike  2.70% Mesophilic acid-phase AD 3  2.50% Mesophilic acid-phase AD 4  2.50% Mesophilic acid-phase AD 5  2.90% Mesophilic acid-phase AD 6  2.90% Thermophilic acid-phase AD 1  10.40% Thermophilic acid-phase AD 2 + Spike  10.40% Thermophilic acid-phase AD 3  10.20% Thermophilic acid-phase AD 4  10.20% Thermophilic acid-phase AD 5  8.80% Thermophilic acid-phase AD 6  8.80% Mesophilic single-phase AD 1  5.90% Mesophilic single-phase AD 2 +Spike  5.90% Mesophilic single-phase AD 3  6.90% Mesophilic single-phase AD 4  6.90% Mesophilic single-phase AD 5  6.80% Mesophilic single-phase AD 6  6.80% Thermophilic single-phase AD 1  6.80% Thermophilic single-phase AD 2 + Spike  6.80% Thermophilic single-phase AD 3  8.70% Thermophilic single-phase AD 4  8.70% Thermophilic single-phase AD 5  7.60% Thermophilic single-phase AD 6  7.60% 1: M, Mesophilic; T, Thermophilic; 1-6 indicate replicates 4.1.5 Precipitate Heavy Metal Contamination    As discussed in the literature review chapter, heavy metal contamination is a well-researched topic (Table 4.3). Unlike other work, this research has intentionally targeted impurities in centrate for removal, with success. More impurities in the fertilizer product may lead to larger concentrations of heavy metals. If heavy metal levels are too high, the fertilizer product will lose value and make the overall heterogeneous struvite precipitation process less feasible. Therefore, quantification of heavy metals in the resultant precipitates was pursued. Table 4.3 lists the concentrations of heavy metals found in struvite based on the upstream AD process. 73  Table 4.3 - Heavy Metal Concentrations in Struvite Precipitates 1  Mesophilic acid-phase AD Thermophilic acid-phase AD Mesophilic single-phase AD Thermophilic single-phase AD Arsenic (As) 0.5 < 0.4 < 0.4 < 0.4 Cadmium (Cd) < 0.04 < 0.04 < 0.04 < 0.04 Chromium (Cr) < 1.0 < 1.0 < 1.0 < 1.0 Cobalt (Co) 0.3 0.2 < 0.1 < 0.1 Copper (Cu) 8.4 19.3 5.4 8.3 Lead (Pb) < 0.2 0.2 < 0.2 < 0.2 Manganese (Mn) 173.3 47.3 5.0 4.0 Mercury 0.1 0.1 0.1 0.4 Molybdenum (Mo) 0.2 0.4 0.1 0.2 Nickel (Ni) 2.8 1.9 2.2 1.0 Titanium (Ti) < 2 < 2 < 2 < 2 Vanadium (V) < 0.4 < 0.4 < 0.4 < 0.4 Zinc (Zn) 4.3 13.0 3.0 5.0 1: data represents average of three time independent samples, in mg/kg-dry weight    The values presented above were averaged across all four categories and compared to the T-4-93 Safety Guideline for Fertilizers and Supplements (Canadian Food Inspection Agency, 2017), the heavy metal concentrations do not use more the 3% of the allowable heavy metal emission capacity for any given metal constituent at a fertilizer application rate of 4,400 kg/ha-yr. The used capacity is graphed visually in Figure 4.8.  Figure 4.8 - Heavy Metal Capacity Use of Resultant Struvite Precipitates for a 4,400 kg/ha-yr Application Rate 0%10%20%30%40%50%60%70%80%90%100%Heavy Metal Capacity Use RemainingCapacityUsedCapacity74  4.2 Incorporating Struvite Ballasted Coagulation into Acid/Gas -Phase Anaerobic Digestion    The results and discussion in Section 4.1 demonstrate heterogeneous struvite precipitation removes colloidal material. Removal of colloidal material and nutrients from acid-phase AD centrate signifies a solution to issues associated with AD of BNR sludge. In particular, internal scaling and sidestream recycle of reactive and recalcitrant nutrients may be avoided. However, some organic constituents such as rbCOD in the centrate persist through the precipitation process. Similar to sidestream recycle of OP in a BNR facility, recycle of rbCOD to the headworks of the WWTF is counterintuitive.     A plausible solution is to use treated acid-phase AD centrate as substrate for a gas-phase AD process. This way the benefits of an acid/gas-phase AD process discussed earlier in Section 2.2.3.4 may be realized without the issues associated with scaling and nutrient recycle discussed in Section 2.2.3.5.2. Furthermore, through the precipitation process colloidal material was removed. Colloidal material has been linked to poor dewaterability (Neyens & Baeyens, 2003). Removal of colloids prior to gas-phase AD may signify enhanced effluent quality and greater AD performance. In this research, the leftover fermented sludge cake was reconstituted into the treated centrate to provide additional substrate for the gas-phase AD process and to observe if the effluent quality will be impacted through potential enzymatic pathways such as hydrolysis of particulate nutrients.     The results presented characterize the AD performance, effluent quality, and biogas production of an acid/gas-phase AD process that incorporates struvite ballasted coagulation on the acid-phase AD centrate. The discussion of the results focuses on the enhancement of AD performance and improvement in AD effluent quality. Effluent quality data for each AD system is tabulated in Table 4.4 for reference during the discussion.    75  Table 4.4 - Effluent Anaerobic Digestion Characterization1,2   Mesophilic  Thermophilic Parameter Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase AD with Struvite  Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase AD with Struvite  Overall Loading Conditions SRT (day) 20 20 20  20 20 20 OLR (g-VS/L/d) 1.83 1.77 1.77  1.83 1.77 1.77 OLR (g-tCOD/L/d) 2.71 2.68 2.68  2.71 2.68 2.68  Overall Removal Efficiency TS Removal (%) 43.0 50.5 46.3  50.3 55.5 46.4 (0.8, 6) (3.0, 6) (2.5, 6)  (1.9, 6) (2.1, 6) (3.1, 6) VS Removal (%) 49.0 54.6 56.1  56.7 60.4 58.9 (1.4, 6) (2.1, 6) (2.2, 6)  (1.9, 6) (1.3, 6) (2.4, 6) tCOD (%) 53.5 55.8 55.3  59.0 57.2 55.7 (3.0, 7) (2.0, 7) (4.6, 7)  (3.8, 7) (3.5, 7) (4.2, 7)  Effluent Characterization OP (mg-P/L) 484.7 480.2 21.3  470.4 471.3 11.9 (24.5, 7) (24.2, 7) (7.4, 7)  (20.1, 7) (22.9, 7) (5.1, 7) NRDP (mg-P/L) 10.9 9.4 1.3  14.0 12.0 3.0 (2.6, 7) (1.1, 7) (0.7, 7)  (4.6, 7) (3.3, 7) (0.5, 7) TP (mg-P/L) 893.2 861.7 365.6  1,117.4 1,195.1 561.5 (13.6, 4) (34.9, 4) (30.6, 4)  (136.3, 4) (32.9, 4) (13.4, 4) Ammonia (mg-N/L) 965.9 1,214.2 1,037.8  1,130.9 1,463.3 1,154.0 (46.5, 7) (33.4, 7) (35.9, 7)  (49.5, 7) (33.7, 7) (166.7, 7) DON (mg-N/L) 22.1 13.5 14.8  41.3 31.9 22.7 (11.2, 7) (10.8, 7) (8.1, 7)  (21.6, 7) (16.9, 7) (12.3, 7) TN (mg-N/L) 2,426.1 2,381.2 2,239.3  2,475.6 2,573.4 2,263.8 (72.0, 4) (49.2, 4) (21.8, 4)  (9.8, 4) (62.1, 4) (54.2, 4) CST (second/%TS) 625 536 256  747 728 735 (20, 3) (59, 3) (6, 3)  (8, 3) (19, 3) (43, 3) Alkalinity                (mg-CaCO3/L) 3,048 3,789 5,335  4,480 4,234 5,541 (482, 5) (661, 5) (336, 5)  (386, 5) (931, 5) (256, 5) VFA (mg/L) 41 198 119  55 87 69 (13, 5) (166, 5) (43, 5)  (25, 5) (42, 5) (13, 5) pH 7.2 7.4 7.6  7.8 7.8 7.9 (0.2, 8) (0.2, 8) (0.2, 8)  (0.1, 8) (0.2, 8) (0.1, 8) 1: data represents arithmetic mean of measurements (standard deviation, number of data points) 2: SRT, solid retention time; OLR, organic loading rate; VS, volatile solids; tCOD, total carbonaceous oxygen demand; TS, total solids; OP, orthophosphate; NRDP, non-reactive dissolved phosphorus; TP, total phosphorus; DON, dissolved organic nitrogen; TN, total nitrogen; CST, capillary suction time; VFA, volatile fatty acids  76  4.2.1 Total Solids Removal    The extent of TS removal through each AD system is presented in Figure 4.9. Acid/gas-phase AD systems removed a larger portion of the substrate TS in comparison to the single-phase AD systems for both mesophilic and thermophilic temperature ranges. However, the acid/gas-phase AD with struvite system performed slightly worse than the respective acid/gas-phase AD system in terms of TS removal. It is likely that the reduced TS removal is a result of an increase in dissolved fixed solids that were dosed to the acid-phase AD centrate to precipitate struvite. A lower TS removal can lead to the assumption that the sludge cake production will increase; however, this may not be the case because a larger portion of the TS dissolved fixed solids may be separated from AD digestate during dewatering. Ultimately, dissolved fixed solids are returned to the headworks of a WWTF and removed in the final effluent. In situations where final effluent chloride or sodium concentrations are restricted, magnesium hydroxide may be used as the magnesium and hydroxide source.  Figure 4.9 - Anaerobic Digestion Total Solids Removal (n = 6) 4.2.2 Volatile Solids Removal       The VS removal data is representative of organic degradation and may be used to rate the performance of an AD system. As seen in Figure 4.10, acid/gas-phase AD performed significantly better than the single-phase AD for mesophilic and thermophilic temperature ranges. Both the conventional single-phase AD and conventional acid/gas-phase AD for mesophilic and thermophilic AD were within 404550556065Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase AD withStruviteTS Removal (%) MesophilicThermophilic77  the reported standard deviations found in previous studies using similar substrate and operating conditions (Wahidunnabi & Eskicioglu, 2014). The acid/gas-phase AD with struvite system did not differ significantly from the respective acid/gas-phase AD system (P = 0.113, mesophilic; P = 0.277, thermophilic). Therefore, methanogenic AD performance was not inhibited by a 1.5:1 magnesium:phosphorus dosage used for heterogeneous struvite precipitation in the acid-phase AD centrate. It is possible that microorganisms in the gas-phase AD would be more sensitive to magnesium inhibition at higher OLR, which is outside the typical range of AD utilizing wastewater sludge.    Figure 4.10 - Anaerobic Digestion Volatile Solids Removal (n = 6) 4.2.3 Methane Production    Similar to VS removal, gas production may be an indication of AD health and is an important performance parameter as methane gas may be beneficially reused. The methane production results plotted in Figure 4.11 show that the acid/gas-phase AD with struvite system did not experience reduced methane production as a result of magnesium inhibition. In fact, the systems performed significantly better than the conventional single-phase AD systems for mesophilic (P = 0.000 < 0.05) and thermophilic (P = 0.000 < 0.05) temperatures. In a comparison to the acid/gas-phase AD systems, the difference was insignificant for mesophilic (P = 0.478 > 0.05) and significant for thermophilic (P = 0.000 < 0.05) temperatures. The stability of thermophilic AD has been shown to be more impacted by influent composition (Labatut et al., 2014) which may be an explanation for the discrepancy in methane production between the mesophilic and thermophilic acid/gas-phase with struvite systems.  404550556065Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteVS Removal (%) MesophilicThermophilic78   Figure 4.11 - Anaerobic Digestion Specific Daily Methane Production (n = 60) 4.2.4 Anaerobic Digestion Effluent Dewaterability    An interest is sparked in the dewaterability data because there is a deviation in the trend between mesophilic and thermophilic acid/gas-phase AD with struvite systems, as shown in Figure 4.12. It is noteworthy that in Figure 4.12, lower CST results correspond to faster dewaterability of digestate. The dewaterability enhancement experienced by the mesophilic acid/gas-phase AD with struvite system has potential to reduce operating cost for a WWTF and make the struvite precipitation scheme feasible from a financial point of view. However, thermophilic acid/gas-phase AD with struvite system did not experience a dewaterability enhancement. To discuss this, Figure 4.2a is referenced in which over 50% of the TSS is removed in the mesophilic acid-phase AD centrate and negligible amount of TSS is removed through the thermophilic acid-phase AD centrate. Therefore, the enhancement in dewaterability is expected to be a result of the removal of colloidal materials in the upstream struvite ballasted coagulation process.  012345Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteSpecific Methane Production at STP (0°C , 1atm)                 (L-CH4/g-VS added) MesophilicThermophilic79   Figure 4.12 - Anaerobic Digestion Effluent Dewaterability (n = 3)    An alternative explanation to enhanced dewaterability is the addition of divalent cations that are reported to act as a bridging agent between two negatively charged particles. This can improve the dewaterability of a sludge (Higgins et al., 2004). Dewaterability improvements have also been associated with struvite precipitation in AD effluent (Bergmans et al., 2014; Brian et al., 2017) which may be associated with the removal of colloidal material. Considering the presented data, the effect of divalent cations may be considered negligible because both thermophilic and mesophilic acid-phase AD centrates were dosed with similar amounts of magnesium and experienced unrelated results. Rather, the removal of colloidal material may be considered to be the cause of dewaterability enhancement. Neyens & Baeyens (2003) reached a similar conclusion in their review article during a discussion on the effect of divalent cation dosing on sludge dewaterability.   4.2.5 Anaerobic Digestion Effluent Nutrient Concentrations     Appling the struvite precipitation methodology to acid-phase AD centrate was effective at removing NRDP, DON and OP. It is unclear as to the fate, production, consumption of these nutrient species through a downstream gas-phase AD vessel. This section documents the nutrient effluent quality of the acid/gas-phase AD with struvite system. The discussion leads to comments on the fate of these nutrient species through the acidogenic and methanogenic steps of AD.   02004006008001000Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteNormalized Capiliary Suction Time                            (sec/% Total Solids) MesophilicThemophilic80  4.2.5.1 Orthophosphate     The OP data plotted in Figure 4.13 shows a significant reduction in effluent OP concentration for the acid/gas-phase with struvite systems. Removal of OP from the AD effluent can result in a significant reduction of scaling in the dewatering centrifuge and downstream processes. For example, Fattah (2012) observed a reduction in phosphorus of 57.2 mg/L between the AD effluent and centrate equalization processes. It should be noted that in the study the AD sludge substrate did not originate from a BNR process. The hydrolysis of non-BNR sludge resulted in an AD effluent OP concentration of 138 mg-P/L and a scaling potential of 57.2 mg-P/L. The hydrolysis of BNR sludge can result in an AD effluent OP concentration of 480 mg-P/L (higher SSR) and a much higher scaling potential. Furthermore, the reduction in OP loading to the mainstream BNR process may reduce the OP concentration in the final effluent of a WWTF (Murakami et al., 1987). No difference in effluent OP concentration was noticed for the conventional digestion systems.   Figure 4.13 - Orthophosphate Concentration in Anaerobic Digestion Effluents (n = 7) 4.2.5.2 Anaerobic Digestion Non-reactive Dissolved Phosphorus Concentrations     Similar to OP, the objective of struvite precipitation on the acid-phase AD centrate was to remove NRDP from the sidestream recycle flows in a WWTF. As shown in Figure 4.14, the acid/gas-phase AD 0100200300400500600Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteOrthophosphate (mg-P/L) MesophilicThermophilic81  with struvite system showed a significant reduction in AD effluent NRDP concentration.  In comparison to single-phase AD, a reduction of 88% and 78% was observed for mesophilic and thermophilic vessels, respectively. No noticeable difference between conventional single-phase AD and acid/gas-phase AD was observed.   Figure 4.14 - Non-reactive Dissolved Phosphorus Concentration in Anaerobic Digestion Effluent   (n = 7) 4.2.5.3 Anaerobic Digestion Total Phosphorus Concentrations     Total phosphorus (TP) representing the particulate and dissolved species of phosphorus was monitored through the AD systems for observational purposes. As expected, acid/gas-phase AD with struvite has a noticeable reduction in effluent TP concentration (Figure 4.15) which is due to the removal of phosphorus as struvite from the system prior to the effluent sampling point (Figure 3.3). This may be seen as a benefit as some biosolids applications are concerned with phosphorus overloading the receiving land (Penn et al., 2002).     TP concentrations in thermophilic AD is consistently higher than the respective mesophilic AD systems. This observation is relevant because it reveals a clue for management of struvite scaling. Considering the OP concentrations were similar (Figure 4.13), the difference in mesophilic and thermophilic TP concentration must be made up of particulate phosphorus. The raw substrate TP concentration is 1100 (± 73) mg-P/L, so there is no overall system removal of TP through the 05101520Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteNon-reactive Dissolved Phosphorus (mg-P/L) MesophilicThermophilic82  thermophilic AD systems and a slight removal of TP through the mesophilic AD systems. A plausible explanation is that mesophilic AD systems are more prone to internal scaling. On the other hand thermophilic AD systems have internal precipitation, but the precipitation is likely nucleating onto materials in the fluid effluent. Thermophilic AD has been known to produce finer colloidal solids and larger colloidal charges (Figure 4.3). These conditions may favour nucleation onto the particulate materials in the fluid substrate rather than external tank surfaces, but this explanation will need to be researched with further detail before being confirmed.   Figure 4.15 - Total Phosphorus Concentration in Anaerobic Digestion Effluents (n = 4) 4.2.5.4 Anaerobic Digestion Ammonia Concentrations     Ammonia forms part of the struvite crystal and is expected to be affected by the struvite precipitation process. When comparing single-phase AD ammonia concentration to the acid/gas-phase AD with struvite system, no significant change in ammonia concentration is observed as a result of struvite precipitation (Figure 4.16). However, ammonia is a by-product of AD, and the extent of ammonia production is related to the extent of organic degradation. Acid/gas-phase AD is expected to produce more ammonia than single-phase AD because of the enhanced AD performance observed in the VS removal data (Figure 4.10). However, the acid/gas-phase AD with struvite system contains less ammonia than the respective acid/gas-phase AD system. Lower ammonia concentration is likely because of ammonia removal from struvite precipitation. Regardless, ammonia concentrations are high among all 0200400600800100012001400Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteTotal Phosphorus (mg-P/L) MesophilicThermophilic83  AD systems and may need pre-treatment in the sidestream to reduce aeration demand on the nitrification processes of the mainstream WWTF.   Figure 4.16 - Ammonia Concentration in Anaerobic Digestion Effluents (n = 7) 4.2.5.5 Anaerobic Digestion Dissolved Organic Nitrogen Concentrations     The DON data in Figure 4.17 was not highly reliable as there were significant deviations from the mean in all samples. Statistically significant observations were not obtained from this data; however, it seems reasonable to observe that thermophilic AD systems produce effluents with slightly more DON in comparison to the respective mesophilic AD systems. Considering that the AD effluent concentrations are similar, using struvite precipitation on acid-phase AD centrates to control sidestream DON may not be the most effective approach to DON treatment.  Figure 4.17 - Dissolved Organic Nitrogen Concentration in Anaerobic Digestion Effluents (n = 7) 8001000120014001600Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteAmmonia (mg-N/L) MesophilicThermophilic010203040506070Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase AD withStruviteDissolved Organic Nitrogen (mg-N/L) MesophilicThermophilic84  4.2.5.6 Anaerobic Digestion Total Nitrogen Concentrations    Similar to the TP data, TN was measured in AD effluent for observational purposes. As expected, acid/gas-phase AD with struvite contained less TN than the other AD systems (Figure 4.18). The removal of ammonia as struvite prior to the effluent sampling location (Figure 3.3) is the likely reason for lower TN concentrations.   Figure 4.18 - Total Nitrogen Concentration in Anaerobic Digestion Effluents (n = 4)    2000210022002300240025002600Single-Phase AD Acid/Gas-Phase AD Acid/Gas-Phase ADwith StruviteTotal Nitrogen (mg-N/L) MesophilicThermophilic85  Chapter 5: Conclusion    The research described in this thesis provides data supporting the use of heterogeneous struvite precipitation as a treatment process to remove NRDP and DON from AD centrate originating from BNR sludge substrate.  1. Heterogeneous struvite precipitation is effective at removing NRDP and DON from AD dewatering centrate. The procedure was successfully demonstrated on: mesophilic acid-phase AD, thermophilic acid-phase AD, mesophilic single-phase AD, and thermophilic single-phase AD. AD operating conditions do not have a significant effect on NRDP removal, which averaged 81.6%. Removal efficiency of DON averaged 65.7% and was significantly (P = 0.001) influenced by the operating temperature of AD reactors. In both cases, the dose of polyacrylamide polymer had significant effect on the NRDP or DON removal rate during heterogeneous struvite precipitation. 2. Polymer dosing rate during centrifugal dewatering had consistent effect on NRDP and DON across all four AD effluents tested. As the dose of dewatering polymer applied to AD effluent increases, the concentration of NRDP may be expected to decrease in the resultant centrate. On the other hand, DON concentration increased in all four AD centrates as the polymer dose was increased. Therefore, polyacrylamide dewatering polymer dosing rate may be used to remove NRDP from AD centrate at the expense of producing DON. 3. Precipitate quality derived from all four centrates was in good standing despite targeting colloidal material for removal. Heavy metal constituents use less than 3% of the emission capacity outlined in the safety guidelines of the Fertilizer Act of Canada under an aggressive fertilizer loading rate. Similarly, antimicrobial concentrations found in struvite can be 3 orders of magnitude lower than the antimicrobial concentrations found in biosolids being land applied to food crops. Therefore, 86  struvite produced from precipitation methodologies targeting heterogeneous nucleation from AD centrate are at low risk of causing environmental concerns for agricultural reuse.   4. Between mesophilic and thermophilic temperature ranges, an additional OP concentration of 331 mg-P/L occurred in acid-phase AD centrates when compared to the respective single-phase AD processes. Therefore, internal scaling due to struvite formation may be significantly reduced by targeting acid-phase AD centrate for phosphorus treatement. 5. In the assessment of the acid/gas-phase AD with struvite system, conventional performance parameters such as biogas production and volatile solids removal were unhindered by added magnesium. The colloidal removal in acid-phase centrate showed promising benefits in terms of digestate nutrient quality and dewaterability.  5.1 Contribution and Significance    The contribution of this research is introduced by applying the heterogeneous precipitation mechanism for the purpose of coagulation. Struvite precipitation is a well-established process for the treatment and recovery of reactive nutrients, but the literature has not explored the potential to use struvite precipitation to coagulate and remove colloidal materials such as recalcitrant nutrients. Demonstration of this precipitation process assists industry and academic professionals in expanding the struvite precipitation technology.     The significance of this research is realized through the successful demonstration of NRDP and DON removal through the precipitation process. This demonstration offers a treatment solution to the production of recalcitrant colloidal material through AD. In particular, large nutrient-removal WWTFs that want to incorporate AD without compromising treated wastewater nutrient concentrations may consider this process.  87  5.2 Future Work    The methodology to remove colloidal material with heterogeneous struvite precipitation is truly engaging. The methodology may be applied to various other streams within a WWTF. Tertiary filtration backwash wastewater is an ideal candidate as this process stream is riddled with colloidal material that is continuously recirculated through a WWTF. Blending the backwash wastewater with AD centrate and applying the heterogeneous precipitation methodology is a promising research topic that was not explored in this work.    The successful demonstration of the heterogeneous struvite precipitation technology was realized for biological phosphorous removal sludge which produce elevated amounts of phosphate in the AD centrates. The effectiveness of the heterogeneous precipitation process on conventional residual sludges without biological nutrient removal is unclear. There is an opportunity to investigate this topic for future research works.     Finally, the heterogeneous precipitation process was demonstrated in batch process configuration. However, full scale technologies are generally continuous processes. The heterogeneous struvite precipitation process should be demonstrated in a continuous flow-through crystallizer. 5.3 Limitations    The research methodology captured the heterogeneous struvite precipitation process under batch process conditions. At full scale this procedure may be realized in large centrate equalization tanks where dewatering equipment is operated semi-continuously. However, as the scale of a WWTF increases, dewatering processes tend to operate continuously and centrate equalization is not required. At this point, a heterogeneous struvite precipitation process that operates continuously is desirable. Developing a continuous process requires laboratory infrastructure such as pumps, in-line controls, and pilot-scale AD vessels. At the time of research, the infrastructure available at the research laboratory was not sufficient to develop a continuous heterogeneous struvite precipitation process.    88  References Aage, H., Andersen, B., Blom, A., & Jensen, I. (1997). 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Water Research, 45(16), 4758-4768.  98  Appendices Appendix A - Preliminary Fermentation Experimental Results   Figure A.1 – Ammonia and Phosphate Release during a) 35°C and b) 55°C Batch Fermentation 00.010.020.030.040.050.060.070.080.090 20 40 60 80 100 120Molar Concentration (mol/L) 35°C Fermentation Time (hours) Ammonia Phosphatea) 0 50 10055°C Fermentation Time (hours) b) 99  Appendix B - Preliminary Polymer Dose Optimization Results    Figure B.1 – Polymer Dose Optimization for Dewatering of a) Mesophilic Acid-Phase AD, b) Thermophilic Acid-Phase AD, c) Mesophilic Acid/Gas-Phase AD, and d) Thermophilic Acid/Gas-Phase AD   00.20.40.61010.51111.51212.50 20 40 60 80 100 120 140 160Centrate TSS (g/L) Cake Solids (%TS) Dose (g-polymer/kg-dry-TS) Cake Total Solids Centrate Total Suspended Solidsa) 00.10.20.30.40.50.61010.51111.51212.50 50 100 150 200Centrate TSS (g/L) Cake TS (%TS) Dose (g-polymer/kg-dry-TS)  Cake Total Solids Centrate Total Suspended Solidsb) 00.20.40.6567890 50 100 150 200Centrate TSS (g/L) Cake Solids (%TS) Dose (g-polymer/kg-dry-TS)  Cake Total Solids Centrate Total Suspended Solidsc) 00.20.40.6567890 50 100 150 200 250 Centrate TSS (g/L) Cake TS (%TS) Dose (g-polymer/kg-dry-TS)  Cake Total Solids Centrate Total Suspended Solidsd) 100  Appendix C - Sample Calibration Curves       y = 9.6951x - 0.0441 R² = 0.9995 01234560 0.2 0.4 0.6Standard Concnetration (mg-P/L) Absorbance (nm) Nov 25, 2016 Total Phosphorus Calibration Curve y = 11.432x - 0.0206 R² = 0.9999 01234560 0.2 0.4 0.6Standard Concentration (mg-P/L) Absorbance (nm) Nov 25, 2016 Orthophosphate Calibration Curve y = 9.1667x - 0.0011 R² = 0.9998 00.511.522.533.544.50 0.2 0.4 0.6Standard Concnetration (mg-P/L) Absorbance (nm) October 2, 2016 Acid-Hydrolysable Phosphorus Calibration Curve y = 1.7541x - 0.7589 R² = 0.9997 0246810120 2 4 6 8Standard Concentration (mg-N/L) Area (microSiemens/minute) Dec 7, 2016 Total Nitrogen Calibration Curve 101    September 14, 2016 Triclocarban Calibration Curve  y = 1.0154x2 + 5.3721x - 0.0043 R² = 0.9999 0123456780 0.5 1 1.5Standard Concentration (mg-N/L) Absorbance (nm) November 14, 2016 Ammonia Calibration Curve y = 2357.7x - 4.8325 R² = 0.9998 01002003004005006007008000 0.1 0.2 0.3 0.4Standard Concentration (mg/L) Absorbance (nm) December 8, 2016 Chemical Oxygen Demand Calibration Curve 102  September 14, 2016 Triclosan Calibration Curve     103  Appendix D - Residual Plots for Methane Production       Figure D.1 – Methane Production Residual Plots for a) Mesophilic Single -Phase AD; b) Thermophilic Single-Phase AD; c) Mesophilic Acid/Gas-Phase AD; d) Thermophilic Acid/Gas-Phase AD; e) Mesophilic Acid/Gas-Phase AD with Struvite; and f) Thermophilic Acid/Gas-Phase AD with Struvite -1.00-0.500.000.501.001.500 50Deviation from Average Value (L-CH4/g-VS added) Time (day) a) -0.60-0.40-0.200.000.200.400.600.801.000 50Deviation from Average Value (L-CH4/g-VS added)  Time (day) b) -0.60-0.40-0.200.000.200.400.600.800 50Deviation from Average Value (L-CH4/g-VS added) Time (day) c) -1.00-0.500.000.501.000 50Deviation from Average Value (L-CH4/g-VS added) Time (day) d) -1.50-1.00-0.500.000.501.001.500 50Deviation from Average Value (L-CH4/g-VS added) Time (day) e) -1.50-1.00-0.500.000.501.001.500 50Deviation from Average  Value (L-CH4/g-VS added) Time (day) f) 

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