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Biological nitrification and dentrification of high ammonia landfill leachate using pre- and post-dentrification… Ilies, Poesis 2000

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BIOLOGICAL NITRIFICATION AND DENITRIFICATION OF HIGH AMMONIA LANDFILL L E A C H A T E USING PRE- AND POST-DENITRJFICATION SYSTEMS AND METHANOL AS SUPPLEMENTARY SOURCE OF ORGANIC CARBON by Poesis Ilies Diploma, Faculty of Civil Engineering "Traian Vuia" Polytechnic Institute, Timisoara, Romania, 1984 A THESIS SUBMITTED PN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF APPLIED SCIENCE in THE FACULTY OF GRADUATE STUDIES DEPARTMENT OF CIVIL ENGINEERING We accept this thesis as conforming to the required standard T H E U M V e W t Y O F B R I T I S H C O L U M B I A December, 1999 © Poesis Hies, 1999 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department of C l \ / l L £K\G I M E£Bl V^Q The University of British Columbia Vancouver, Canada Date b E C . 2 . 0 , DE-6 (2/88) 11 ABSTRACT Landfill leachate discharges characterized by high nitrogen concentrations are detrimental to the environment. This research investigated nitrogen removal capabilities of a nitrification with pre- and post denitrification biological process, when treating a landfill leachate characterized by high ammonia concentrations and low levels of biodegradable organics. The process configuration is generally referred to as 4-Stage Bardenpho process, and consists of a sequence of anoxic and aerobic zones. Two 4-Stage Bardenpho systems with first anoxic reactor actual hydraulic retention times of 1.5 and 1.7 hours, respectively, were operated in parallel. The first system had a first aerobic reactor actual hydraulic retention time of 3 hours, while the second system had one of 3.4 hours. Complete nitrification and denitrification of the "base" landfill leachate, with an average ammonia concentration of 200 mg N/L and organic carbon levels of less than 50 mg N/L, was achieved during the Base Leachate Phase. Higher concentrations were simulated by artificially increasing influent ammonia concentration to about 2200 mg N/L during the Ammonia Loading phase. The pH Phase examined the acclimatization of the bacterial populations to decreased reactor pH levels, as a means to improve system overall denitrification performance and decrease effluent residual NOx concentrations. The overall process performance, at progressively decreased ambient temperature from 20 °C to 10 °C, was investigated during the Temperature Phase. Both systems experienced nitrification inhibition during the first attempt at incrementally increasing influent ammonia concentrations to over 2200 mg N/L. Methanol loadings increased concomitantly with ammonia loadings, to match expected aerobic NOx production and using CH 3OH:NOx ratios of about 20:1; this resulted in methanol bleeding into the first aerobic zone, enhanced aerobic heterotrophic growth, and further inhibition of the mtrifying population, already inhibited by the recycling through the elevated "free" ammonia levels of the first anoxic zone. When the systems were allowed to acclimatize to each incremental ammonia increase, and with methanol loadings changed to yield CH3OH:NOx removed ratios of only 5:1, the desired influent ammonia concentration was reached within 88 days from the start of the Ill second attempt. The systems generated ammonia free effluents with NOx concentrations of 78 mg N/L and 250 mg N/L, respectively, treating simulated landfill leachate containing over 2200 mg N/L of ammonia. The systems started to respond positively to decreased reactor pH values by producing effluents with lower NOx concentrations; However, subsequently, due to some inhibitory constituents in the natural landfill leachate, percentage denitrification rates decreased to about 50% and the systems produced effluent residual NOx concentrations of about 170 mg N/L, When ambient temperature decreased from 20 °G to 17 °C, and, subsequently, to 14 °C, there seemed to be no negative influence on the nitrification processes of the systems, while denitrification inhibition was observed starting with the ambient temperature of 17 °C. Nevertheless, at ambient temperatures of 10 °C, the percentage nitrification rates of the systems decreased from about 100% to 10% and 30%, respectively, while percentage denitrification rates decreased to less than 5%. Decreased bacterial growth rates and nitrifiers inhibition at low temperatures were suspected to be the principal factors that determined the failure of the processes. However, percentage total ammonia removal of about 50% was maintained by both systems. Except for the decreased ambient temperature and system failure periods, the overall performance of the system with first anoxic reactor actual hydraulic retention time of 1.5 hours, and first aerobic reactor actual hydraulic retention time of 3 hours, was more effective than the performance of the system with actual hydraulic retention times of 1.7 hours, and 3.4 hours, respectively, as the system usually generated ammonia free effluents with lower NOx concentrations. T A B L E OF CONTENTS iv ABSTRACT ii TABLE OF CONTENTS iv LIST OF FIGURES viii LIST OF PLATES xiv LIST OF TABLES xiv ACKNOWLEDGMENTS xv INTRODUCTION 1 1.1 Leachate Generation and Quality 1 1.2 Effects of Nitrogenous Discharges on the Environment 4 1.3 Nitrogen Removal Options 5 1.3.1 Physical-Chemical Methods 6 1.3.2 Bacterial Assimilation as Ammonia Removal Method 6 1.3.3 Biological Nitrification and Denitrification 7 (a) Nitrification 7 (b) Denitrification 8 1.4 Treatment Option and Process Configuration 9 1.5 Modified Ludzack-Etringer Process versus Four-Stage Bardenpho Process 11 1.6 Study Rationale and Objectives 16 LITERATURE REVIEW 18 2.1 UBC Research Experience on Leachate Treatment 18 2.1.1 Biological Nitrification and Denitrification of Landfill Leachate 18 2.1.2 Effect of Increased Ammonia Levels on Landfill Leachate Biological Treatment 18 2.1.3 Effect of External Organic Carbon Source and Loading on Denitrification 19 2.1.4 Effect of Low Temperature on Landfill Leachate Biological Treatment 20 2.1.5 Effect of Heavy Metals on Landfill Leachate Biological Treatment 20 2.1.6 Effect of Sludge Recycle Ratio on Landfill Leachate Biological Treatment 22 2.2 Research Experience on Nitrification Performance and Inhibition 22 2.3 Research Experience on Denitrification Performance and Inhibition 24 EXPERIMENTAL SETUP AND OPERATION 26 3.1 Treatment System Design and Operating Parameters 26 3.2 Leachate Feed 31 3.3 Chemical Feed 33 3.4 System Startup and Operation 34 ANALYTICAL METHODS 37 4.1 Total Ammonia ( N H / + NH 3) 37 4.2 Nitrite Plus Nitrate (NOx) 38 4.3 Nitrite(N0 2) 38 4.4 Orthophosphate 38 4.5 Suspended Solids (TSS and VSS) 39 4.6 Chemical Oxygen Demand (COD) 39 4.7 Biochemical Oxygne Demand (BOD5) 39 4.8 Alkalinity 4.9 pH 4.10 Oxidation-Reduction Potential (ORP) 4.11 Dissolved Oxygen (DO) 4.12 Temperature RESULTS AND DISCUSSION 42 5.1 Base Leachate Phase 43 5.1.1 Ammonia Levels and Removal 44 5.1.2 NOx Levels, Nitrification and Denitrification 47 5.1.3 pH levels, Alkalinity and Methanol Loadings 52 5.1.4 VSS Levels 55 5.1.5 COD Loading versus COD Removal 59 5.2 Ammonia Loading Phase 62 5.2.1 Ammonia and NOx Levels 64 5.2.2 Alkalinity and Methanol Loadings and Their Effects on Process Performance 70 5.2.3 VSS Levels 91 5.2.4 System Performance and Sequenced Reactor Performance 96 (a) Anoxic #1 Reactor 101 (b) Aerobic #1 Reactor 106 (c) Anoxic #2 Reactor 106 (d) Aerobic #2 Reactor 111 5.3 pH Phase 111 5.3.1 Ammonia Levels and Nitrification Rate 113 5.3.2 NOx Levels and Denitrification Rate 118 5.3.3 Specific Denitrification Rates and VSS Levels 125 40 40 41 41 41 vn 5.3.4 System performance 130 5.4 Temperature Phase 133 5.4.1 Ammonia Levels and Aerobic #1 Reactor %Nitrification 135 5.4.2 Methanol Loading Effects on Percentage Nitrification and Denitrification 140 5.4.3 VSS Levels and Specific Nitrification and Denitrification Rates 144 5.4.4 Effluent Total Ammonia and NOx Levels 151 SUMMARY, CONCLUSIONS AND RECOMMENDATIONS 155 6.1 Summary 155 6.2 Conclusions 159 6.3 Recommendations 162 REFERENCES 164 APPENDICES APPENDIX A: Formulas and Calculation Definitions & Explanatory Notes 168 APPENDIX B: Operating Temperature and Leachate Characteristics & Raw and Calculated Data 175 LIST OF FIGURES viii Figure 1.1: Pre/Post Denitrification System Configuration 12 Figure 1.2: Combined Nitrification/Denitrification Systems 14 Figure 5.1: System # 1 - Base Leachate Phase Anoxic and Aerobic Total Ammonia ( N H / + NH 3) 45 Figure 5.2: System #2 - Base Leachate Phase Anoxic and Aerobic Total Ammonia ( N H / + NH 3) 46 Figure 5.3: System #1 - Base Leachate Phase Anoxic and Aerobic NOx 48 Figure 5.4: System #2 - Base Leachate Phase Anoxic and Aerobic NOx 49 Figure 5.5: System # 1 - Base Leachate Phase Aerobic #1 & Anoxic NOx and Methanol Loading 50 Figure 5.6: System #2 - Base Leachate Phase Aerobic #1 & Anoxic NOx and Methanol Loading 51 Figure 5.7: System # 1 - Base Leachate Phase Anoxic & Aerobic pH and Alkalinity & Methanol Loading 53 Figure 5.8: System #2 - Base Leachate Phase Anoxic & Aerobic pH and Alkalinity & Methanol Loading 54 Figure 5.9: System #1 - Base Leachate Phase Anoxic & Aerobic VSS and Methanol Loading 56 Figure 5.10: System #2 - Base Leachate Phase Anoxic & Aerobic VSS and Methanol Loading 57 Figure 5.11: System # 1 - Base Leachate Phase Anoxic #1 VSS and Ammonia Concentrations 58 Figure 5.12: System # 1 - Base Leachate Phase Anoxic #1 & Aerobic #1 COD Removal Rate and COD Loading Figure 5.13: System #2 - Base Leachate Phase Anoxic #1 & Aerobic #1 COD Removal Rate and COD Loading Figure 5.14: System # 1 - Ammonia Loading Phase Anoxic and Aerobic Total Ammonia [ N H / + NH 3] Figure 5.15: System #2 - Ammonia Loading Phase Anoxic and Aerobic Total Ammonia [ N H / + NH 3] Figure 5.16: System # 1 - Ammonia Loading Phase Anoxic and Aerobic NOx Figure 5.17: System #2 - Ammonia Loading Phase Anoxic and Aerobic NOx Figure 5.18: System # 1 - Ammonia Loading Phase Aerobic Nitrification Rate and Alkalinity Loading Figure 5.19: System #2 - Ammonia Loading Phase Aerobic Nitrification Rate and Alkalinity Loading Figure 5.20: System #1-Ammonia Loading Phase Aerobic NOx and Alkalinity Consumption Figure 5.21: System #2 - Ammonia Loading Phase Aerobic NOx and Alkalinity Consumption Figure 5.22: System # 1 - Ammonia Loading Phase Anoxic Denitrification Rate and Methanol Loading Figure 5.23: System #2 - Ammonia Loading Phase Anoxic Denitrification Rate and Methanol Loading Figure 5.24: System #1 - Ammonia Loading Phase Anoxic NOx and Methanol Addition (CH 3OH : NOx Removed) Figure 5.25: System #2 - Ammonia Loading Phase Anoxic NOx and Methanol Addition (CH 3OH : NOx Removed) Figure 5.26: System #1 - Ammonia Loading Phase Anoxic and Aerobic "Free" Ammonia [NH3] Figure 5.27: System #2 - Ammonia Loading Phase Anoxic and Aerobic "Free" Ammonia [NH3] Figure 5.28: System #1 - Ammonia Loading Phase Anoxic "Free" Ammonia and pH Figure 5.29: System #2 - Ammonia Loading Phase Anoxic "Free" Ammonia and pH Figure 5.30: System #1 - Ammonia Loading Phase Anoxic and Aerobic Nitrite [N02] Figure 5.31: System #2 - Ammonia Loading Phase Anoxic and Aerobic Nitrite [ N O 2 ] Figure 5.32: System #1 - Ammonia Loading Phase Anoxic and Aerobic NO2 /NOX Ratio Figure 5.33: System #2 - Ammonia Loading Phase Anoxic and Aerobic NO2 /NOX Ratio Figure 5.34: System #1 - Ammonia Loading Phase Anoxic and Aerobic VSS Figure 5.35: System #2 - Ammonia Loading Phase Anoxic and Aerobic VSS Figure 5.36: System #1 - Ammonia Loading Phase Anoxic VSS Figure 5.37: System #2 - Ammonia Loading Phase Anoxic VSS Figure 5.38: System #1 -Ammonia Loading Phase Effluent Total Ammonia and NOx Figure 5.39: System #2 - Ammonia Loading Phase Effluent Total Ammonia and NOx Figure 5.40: System #1 - Ammonia Loading Phase System %Ammonia and %Total Inorganic Nitrogen Removal Figure 5.41: System #2 - Ammonia Loading Phase System %Ammonia and %Total Inorganic Nitrogen Removal Figure 5.42: System #1 - Ammonia Loading Phase Anoxic #1 %Ammonia Removal and %Demtrification Figure 5.43: System #2 - Ammonia Loading Phase Anoxic #1 %Ammonia Removal and %Denitrification Figure 5.44: System #1 - Ammonia Loading Phase Aerobic #1 %Ammonia Removal and %Nitrification Figure 5.45: System #2-Ammonia Loading Phase Aerobic #1 %Ammonia Removal and %Nitrification Figure 5.46: System #1 - Ammonia Loading Phase Anoxic #2 %Ammonia Removal and %Denitrification Figure 5.47: System #2 - Ammonia Loading Phase Anoxic #2 %Ammonia Removal and %Denitrification Figure 5.48: System #1 - Ammonia Loading Phase Aerobic #2 %Ammonia Removal and %Nitrification Figure 5.49: System #2 - Ammonia Loading Phase Aerobic #2 %Ammonia Removal and %Nitrification Figure 5.50: System #1 - pH Phase Anoxic and Aerobic Total Ammonia [ N H / + NH3] Figure 5.51: System #2 - pH Phase Anoxic and Aerobic Total Ammonia [NFL/ + NH3] Figure 5.52: System #1 - pH Phase Nitrification Rate and Aerobic pH Figure 5.53: System #2 - pH Phase Nitrification Rate and Aerobic pH Figure 5.54: System #1 - pH Phase Anoxic and Aerobic NOx Figure 5.55: System #2 - pH Phase Anoxic and Aerobic NOx Figure 5.56: System #1 - pH Phase Denitrification Rate and Anoxic pH Figure 5.57: System #2 - pH Phase Denitrification Rate and Anoxic pH Figure 5.58: System #1 - pH Phase Specific Denitrification Rate and Anoxic pH Figure 5.59: System #2 - pH Phase Specific Denitrification Rate and Anoxic pH Figure 5.60: System #1 - pH Phase Specific Denitrification Rate and Anoxic VSS Figure 5.61: System #2 - pH Phase Specific Denitrification Rate and Anoxic VSS Figure 5.62: System #1 - pH Phase Effluent Nitrogen Levels and System %Removals Figure 5.63: System #1 - pH Phase Effluent Nitrogen Levels and System %Removals Figure 5.64: System # 1 - Temperature Phase System Total Ammonia Levels [NH/ + NH3] Figure 5.65: System #2 - Temperature Phase System Total Ammonia Levels [NFl/ + NH 3] Figure 5.66: System #1 - Temperature Phase Aerobic #1 %Nitrification and Anoxic #1 %Denitrification Figure 5.67: System #2 - Temperature Phase Aerobic #1 %Nitrification and Anoxic #1 %Denitrification Figure 5.68: System #1 - Temperature Phase Aerobic #1 COD Levels versus %Nitrification Figure 5.69: System #2 - Temperature Phase Aerobic #1 COD Levels versus %Nitrification Figure 5.70: System #1 - Temperature Phase Anoxic #1 CH30H:NOxE n t ering Ratio and Aerobic #1 %Nitrification & Anoxic #1 %Denitrification Figure 5.71: System #2 - Temperature Phase Anoxic #1 CH30H:NOxEntCTing Ratio and Aerobic #1 %Nitrification & Anoxic #1 %Denitrification Figure 5.72: System #1 - Temperature Phase Aerobic Specific Nitrification and VSS Levels Figure 5.73: System #2 - Temperature Phase Aerobic Specific Nitrification and VSS Levels Figure 5.74: System # 1 - Temperature Phase Anoxic Specific Denitrification and VSS Levels Figure 5.75 System #2 - Temperature Phase Anoxic Specific Denitrification and VSS Levels XIV Figure 5.76: System #1 - Temperature Phase Anoxic NCVNOx Ratio and pH Levels Figure 5.77: System #2 - Temperature Phase Anoxic N0 2/NOx Ratio and pH Levels Figure 5.78: System #1 - Temperature Phase System %Total Ammonia and %Total Inorganic Nitrogen Removal Figure 5.79: System #2 - Temperature Phase System %Total Ammonia and %Total Inorganic Nitrogen Removal 149 150 152 153 LIST OF PLATES Plate 3.1: System #1 Setup in the Environmental Room 30 LIST OF TABLES Table 3.1: System Design and Operating Parameters Table 3.2: Burns Bog Leachate Composition Table 5.1: System #1 Anoxic Reactor pH, NOx and %Denitrification Table 5.2: System #2 Anoxic Reactor pH, NOx and %Denitrification Table 5.3: System Nitrogen Removal Overall Performance 27 32 123 124 130 ACKNOWLEDGMENTS xv I would like to thank Professor Dr. D.S. Mavinic for his guidance, support, and personal encouragement, throughout the course of this research, and especially, for listening with patience and understanding at my endless "bug stories". I would also like to thank Paula Parkinson and Susan Harper of the UBC Environmental Engineering Laboratory for their invaluable technical assistance, for analyzing hundreds of samples and, sometimes, reanalyzing them because "they didn't look quite right", for their good humor and support that really made a difference. In addition, I would like to thank the technicians of the UBC Civil Engineering Workshop for their assistance related to experimental setup and electrical equipment safety. I am extremely grateful to Dean Shiskowski, P.Eng., consultant engineer and former environmental engineering graduate student, for his recommendations regarding my research work, and his helpful tips regarding rising generation after generation of "happy bugs". I am also grateful to Paul Henderson, Manager of the Landfill Operation Branch of the City of Vancouver for facilitating the collection of the leachate used in this study. Also, I wish to thank George Twarog, Environmental Technician at the Burns Bog Landfill, for helping me with the actual leachate collection and for making my "leachate collection days" easy. I would like to acknowledge the tremendous contribution of previous research works and findings to the completion of this study. Accomplished dreams and remarkable achievements of numerous known and unknown researchers made this research possible. I hope my work will also be a source of inspiration for further endeavors and who better to say it in my place than John Lennon (1940-1980) with his Imagine (1971): "You may say I'm a dreamer But I'm not the only one I hope someday you '11 join us..." r Chapter 1 INTRODUCTION Disposal of solid wastes in soils, or landfilling of solid wastes, is a traditional method of waste disposal and a fundamental part of solid waste management. Modern landfill designs, advanced landfill technologies, and new landfill operation strategies have been used to prevent and control potential impacts of this disposal method on the environment. The most significant environmental concern associated with sanitary landfilling is landfill leachate production. Landfill leachate is generated mainly by the infiltrating water that passes through the solid waste fill and facilitates transfer of contaminants from a solid phase (waste) into a liquid phase (infiltealing water). Leachate generation and composition are determined by the quality of solid waste, the biological and chemical processes occurring in the waste fill, and the amount of precipitation and percolation. Landfill leachates usually contain high levels of contaminants and can not be discharged into environment without prior treatment. Landfill leachate production starts at the early stages of the landfill and continues several decades after landfill closure. Therefore, landfill leachate management facilities have to remain operational over long periods of time. Leachate control and treatment are important parts of if the new strategies for integrated solid waste management and environmental protection. A thorough understanding of leachate physical and chemical characteristics and the effects of these characteristic on treatment strategies is required before the selection of the most appropriate treatment schemes. There is no unique solution to leachate treatment as landfill leachate composition and treatability vary among landfills worldwide (Mavinic, 1998). 1.1 Leachate Generation and Quality Leachate generation is determined by the infiltrating rainfall, snow melt, or groundwater that provides the major transport phase for the dissolution and migration of the contaminants from the solid waste. The mass transfer of the contaminants from the waste into the percolating liquid and, therefore, landfill leachate quality, is controlled by the biological activity occurring within the fill. The microbial activity within landfill is regulated by the fill "age" and eventually leads to the complete stabilization of the waste. The stabilization of municipal sanitary landfills has been described by Pohland et al. (1985) as going through five phases, distinguishable one from another by their leachate and gas characteristics. The initial phase, or initial adjustment phase, is the time between the deposition of the waste and the production of measurable amounts of gas and leachate. During this phase the fill settles and accumulates moisture up to a point of significant reaction opportunity. The second phase, or transition phase, is characterized by quick decline of oxygen levels and sequential changes from aerobic(i.e., molecular oxygen is present) conditions to anoxic(i.e., oxygen is replaced by oxygen compounds, such as nitrates and sulfates, as electron acceptors) conditions and, then, to anaerobic (i.e., total absence of oxygen) conditions. The incipient production of intermediate organic volatile fatty acids (VFAs) determines an increase in the biochemical oxygen demand (BOD5) levels of the generated leachate. Once conditions are anaerobic, fermentation occurs, and the production of VFAs is predominant during the third phase, or acid formation phase. The presence of VFAs causes a significant drop in pH and generates the dissolution of metals. Phosphorus and ammonia are released and partially metabolized by bacteria. As most of the phosphorus is utilized by bacteria while ammonia is in excess, the generated leachate is characterized by low levels of phosphorus and high levels of ammonia. BOD 5 and COD (chemical oxygen demand) concentrations of the leachate are also high with typical BOD 5/COD ratios higher than 4 (Ehrig, 1985). A landfill in the acid formation phase is generally referred to as "young" landfill or "acetogenic" landfill (Azevedo, 1993) and its leachate is called "young" leachate or carbon based leachate (i.e., due to high concentration of organic carbon still present in the leachate) (Mavinic, 1998). 3. In the fourth phase, or methane fermentation phase, methanogenic bacteria convert VFAs to methane gas, carbon dioxide, and water. Landfill buffer system, previously controlled by VFAs, changes to a buffer system controlled by bicarbonate alkalinity generating increased pH levels and, consequently, metal precipitation. Leachate metal concentrations significantly decrease while ammonia concentrations remain high. Increased methane production determines considerable decreased BOD 5 and COD levels in the leachate. Typical BOD5 /COD ratios are smaller than 0.1 (Ehrig, 1985). A landfill in a methane fermentation phase is usually called an "old", "older", or "methanogenic" landfill (Azevedo, 1993). The generated leachate is referred to as "old" leachate, or nitrogen based leachate (Mavinic, 1998). As substrates are exhausted, methane and carbon dioxide production decreases indicating the installment of the fourth phase, or final maturation phase. Compounds more resistant to biodegradability are transformed into humus like matter, which may generate some increase in metal concentrations. Ammonia leachate levels may continue to be relatively high before ceasing. The wastes are considered stabilized. Pohland et al. (1985) description of the five phases of a landfill contributes to a better understanding of the biological processes occurring within a solid waste fill. However, in reality, the complexity of landfilling operations in addition to the intricacy of a biological process makes difficultly to prognosticate the relative duration and magnitude of each of these phases and, therefore, the expected composition of a leachate. Furthermore, commonly used terms, such as "young", "old", and "age" when referring to landfills and leachates, are arbitrary and of no consequence. A leachate characterized by high levels of organic carbon and relative low levels of nitrogen should be referred to as a carbon based leachate. Similarly, a leachate with high levels of nitrogen and relative low levels of carbon should be referred to as a nitrogen based leachate. The two landfill leachate constituents considered major potential environmental concerns are organic carbon and nitrogen. Therefore, treatment method priority has to be organic carbon removal from a 4 carbon based leachate and nitrogen removal from a nitrogen based leachate. The present study presents the findings of a research investigating the treatability of a landfill leachate characterized by high ammonia levels and low organic carbon concentrations, hence, it is specifically focused on nitrogen removal processes. Therefore, carbon removal will be discussed only relative to biological nitrogen removal, and nitrification and denitrification inhibition. 1.2 Effects of Nitrogenous Discharges on the Environment Nitrogen exists in many forms in the environment and is essential in maintaining the natural balance of the ecosystems. However, excessive accumulations of nitrogen due to anthropogenic influences may cause significant undesirable effects on the environment. The most significant environmental problems associated with nitrogenous discharges into aquatic ecosystems include biostimulation of plant and algal growth in surface waters, dissolved oxygen depletion caused by nitrification, aquatic life toxicity to ammonia, concerns related to public health (EPA, 1993). Biostimulation of plant and algal growth in surface waters, or cultural eutrophication, is the excessive plant and algae growth as a result of nutrient, primarily nitrogen and/or phosphorus, enrichment, or over-fertilization, of rivers, lakes and estuaries (EPA, 1993). Eutrophication of lakes is of most concern as nutrient inputs into the body of water tend to be recycled within the lake and to accumulate over a time period. Phosphorus-induced eutrophication tends to occur within freshwater systems, usually deficient in phosphorus compounds, while nitrogen- induced eutrophication tends to occur in within marine environments, usually deficient in nitrogen compounds (Azevedo, 1993; EPA, 1993). Eutrophication is typically controlled by limiting nitrogen and phosphorus loads into aquatic environments. Major consequences of eutrophication are deterioration of the appearance of the previously clean waters, unpleasant odors and reduced dissolved oxygen levels as a result of plant and algae decomposition, and respiratory problems for fish and benthic aquatic animals (EPA, 1993). 5 In addition to dissolved oxygen depletion as a result of bacterial decomposition of organic carbon compounds, biological nitrification of ammonia discharges further decreases oxygen concentration within the body of water. Toxicity effects of the aquatic life are generated by high levels of ammonia-nitrogen, specifically by molecular ammonia, or unionized "free" ammonia (NH3). In aquatic environments, dissolved ammonia exists as an equilibrium of molecular ammonia (NH3) and ammonium ion (N0 4) (EPA, 1993). The balance between the two compounds is controlled by pH and temperature. The percentage of available molecular ammonia, relative to ammonia-nitrogen, considerable increases with increases in pH and/or temperature (EPA, 1993; Shiskowski, 1995). Acute toxicity effects for freshwater fish species and saltwater fish species have been observed at concentrations ranging from 0.1 mg N/L of molecular ammonia to 10 mg N/L of molecular ammonia. Public health concerns are associated with the undesirable effects that nitrates (N0 3) and nitrites (N0 2) may have on humans. Methemoglobinemia, or infantile cyanosis, or blue babies syndrome, is an illness that primarily affects infants The disease is a result of nitrate reduction to nitrite in the stomach or saliva, followed by iron oxidation using nitrite ion in the hemoglobin molecules (EPA, 1993). Anoxia, suffocation, or death may occur as oxygen transfer can not be accomplished by the resulting methemoglobin. Therefore, drinking water guidelines recommend that the nitrates levels of the drinking water should be less than 10 mg N/L. Furthermore, high levels of nitrates in the drinking water result in increased production of nitrites that can react with amines and amides. The reaction of nitrite with amines and amides results in nitroso-compounds that may facilitate carcinogenic diseases. 1.3 Nitrogen Removal Options Physical-chemical treatment, bacterial assimilation, and aerobic biological nitrification with discretional subsequent denitrification are the principal nitrogen removal methods from liquid wastes. The selection of the treatment method, treatment process configuration, and treatment plant design and operating parameters for a landfill leachate is a laborious process that has to take in consideration not only the general characteristics of the leachate, but also seasonal surges in hydraulic loadings and leachate composition (EPA, 1995; Mavinic, 1998). Furthermore, landfill leachate treatment schemes have to be flexible in order to accommodate expected time changes of landfill leachate characteristics. A "young" landfill, that generates a carbon based leachate, will eventually become an "old" landfill, that generates a nitrogen based leachate (Mavinic, 1998). 1.3.1 Physical-Chemical Methods Air stripping and ion exchange, using clinoptilolite (i.e., zeolite with high selectivity for ammonium and calcium ions) columns are two available methods viewed by Forgie (1998b)as potential landfill leachate ammonia removal options. However, due to high related costs, none of these two methods can be alone considered for ammonia removal from a landfill leachate. The need of considerable amounts of basic solution additions for increasing the leachate pH and large aeration volumes for air stripping limit the employment of air stripping to situations requiring only residual ammonia removal (Ehrig, 1998). Similarly, the solely use of the ion exchange method as landfill leachate ammonia removal option is not feasible due to prohibitive costs of the media and media subsequent regeneration (EPA, 1995; Forgie, 1998b). Ammonia removal using a reverse osmoses process, that returned the resulting liquid concentrate back into the landfill, was considered by Ehrig (1991) only as a provisory treatment solution for a landfill leachate. However, in combination with other physical-chemical or biological ammonia removal methods, reverse osmosis is a reliable and feasible landfill leachate treatment option (Albertsen et al., 1998; EPA, 1995). 1.3.2 Bacterial Assimilation as Ammonia Removal Method Acknowledging that, in the process of aerobic oxidation of the biodegradable organic mater, nitrogen is required for bacterial growth, any conventional biological treatment scheme is a potential landfill leachate ammonia removal method, providing that a BOD 5:N ratio of about 100:5 exists in the influent leachate 7 (Mavinic, 1997 and 1998). Based on this concept, Robinson and Grantham (1988) used bacterial assimilation to remove ammonia from a landfill leachate with concentration of about 800 mg N/L of ammonia and low concentration of BOD5. The supplementary organic carbon required for a complete assimilation of the available ammonia was provided using jam waste from a nearby factory. The feasibility of bacterial assimilation as landfill leachate ammonia removal method is limited by the existence of a cheap and reliable source of supplementary organic carbon, and the availability of a disposal option for large quantities of nitrogenous sludge (Mavinic, 1997). 1.3.3 Biological Nitrification and Denitrification (a) Nitrification Nitrification is the sequential oxidation of ammonium ion (NH/) first to nitrite ( N O 2 " ) and than to nitrate (NO3"). The process is carried out by autotrophic bacteria that derive energy for growth from the oxidation of inorganic nitrogen compounds (i.e., ammonia and nitrite) and use inorganic carbon (i.e., carbon dioxide) for cell synthesis. Nitrosomonas and Nitrobacter are the two principal genera of bacteria responsible for nitrification. Characteristic for these two genera of bacteria is their ability to oxidize only one specific species of inorganic nitrogen compounds. Nitrosomonas converts ammonium ion to nitrite while Nitrobacter converts nitrite to nitrate. Assuming growth yields of 0.08 g VSS/g N H / - N and of 0.05 g VSS/g N0 2"-N, and empirical formulation of bacterial cells to be C5H7NO2, stoichiometric equations for oxidation and cell synthesis for Nitrosomonas and Nitrobacter respectively, are as follows (EPA, 1993): Nitrosomonas: 1.00 N H / + 1.44 0 2 + 0.0496 C 0 2 -» 0.01 C 5 H 7 N 0 2 + 0.990 N0 2" + 0.970 H20 + 1.99 FT Nitrobacter. 1.00 N0 2" + 0.00619 N H / + 0.031 C 0 2 + 0.0124 H 20 + 0.50 0 2 -> 0.00619 C 5 H 7 N 0 2 + 1.00 N0 3" +0.00619 H + The overall reaction describing complete nitrification can be obtain combining the above equations: 1.00 NFL,+ + 1.89 0 2 +0.0805 C 0 2 -> 0.00161 C 5 H 7 N 0 2 + 0.952 H 20 + 0.984 NO3" + 1.98 H + 8 The above stoichiometric equations demonstrate that nitrification requires a significant amount of oxygen (i.e., 4.6 g 02required/g N H / - N , design recommendation), generates a small amount of biomass (i.e., 0.1 g VSS produced as nitrifiers/g N H / - N , design recommendation), and destroys alkalinity through hydrogen ions production (i.e., 7. lg alkalinity as CaC03/g N H / - N , design recommendation) (EPA, 1993). EPA (1993) recommends a pH range of 6.5 to 8.0 for a stable and reliable nitrification performance. Other authors are more conservative (Metcalf & Eddy, 1991; Robinson et al., 1998). (b) Denitrification Denitrification is the anoxic two-step conversion of nitrate first to nitrite, and than to gaseous nitrogen compounds, such as nitrogen gas (N2), that are released into the atmosphere. Denitrification is carried out by heterotrophic bacteria that use organic carbon for cell synthesis as well as for deriving energy for growth. Denitrifiers are facultative aerobes and, therefore, they have the ability to use both oxygen or nitrate and nitrite as electron acceptors. In the presence of oxygen, denitrifiers preferentially use molecular oxygen as electron acceptor for generating more energy. Nitrate and nitrite are used as electron acceptor only in the absence of oxygen and as a substitute for oxygen (EPA 1993). Therefore, anoxic conditions (i.e., absence of molecular oxygen) are essential for denitrification. A wide range of bacteria can accomplish denitrification mcluding Achromobacter, Aerobacter, Alcaligenes, Bacillus, Brevbacterium, Flavobacterium, Lactobacillus, Micrococcus, Proteus, Pseudomonas, Spirillum, etc. (Metcalf & Eddy, 1991). Stoichiometric equations for the complete conversion of nitrate and nitrite to nitrogen gas, using methanol (CH3OH) as organic carbon source and nitrate and nitrite as electron acceptor, are as follow (EPA 1993): Nitrate Reduction: N0 3" + 1.08 CH 3 OH + 0.24 H 2 C 0 3 -> 0056 C 5 H 7 N 0 2 + 0.47 N 2 + 1.68 H20 + HCO3" Nitrite Reduction: N0 2 _ + 0.67 CH 3 OH + 0.53 H 2 C 0 3 -> 0.04 G 5 H 7 N0 2 + 0.48 N 2 + 1.23 H 20 + HCO3-9 For a complete nitrate reduction the methanol requirements are 2.47 g CH 3OH/g NO3-N or 3.7 g COD/g N 0 3 - N while for a complete nitrite reduction the methanol requirements are 1.53 g CH 3OH/g N0 2"-N or 2.3 g COD/g N0 2"-N. Methanol is also required to reduce any dissolved oxygen (DO) present and, therefore, the overall methanol requirements for denitrification are described by the following equation (EPA, 1993): M = 2.47 (N03~-N) + 1.53 (N02~-N) + 0.87 DO Bicarbonate alkalinity is generated during denitrification at a ratio of 3.57 g CaC03/g N0 2"-N denitrified. The hydroxide ions released during nitrite conversion react with the carbon dioxide from water and produce bicarbonate ions (EPA, 1993), hence, bicarbonate alkalinity. EPA (1993) recommends a pH range between 6 and 8, as below these values denitrification rates may be considerable low. 1.4 Treatment Option and Process Configuration Biological nitrification and denitrification is a reliable, effective and relatively low cost method of removing nitrogen from municipal and industrial discharges (Metcalf &Eddy, 1991; EPA, 1995). In the most recent years, landfill leachates characterized by low organic carbon concentration, high nitrogen levels, and a large spectrum of metals have become a major concern all over the world (Mavinic, 1998; Robinson and Gronow, 1998). Various biological systems have been widely investigated, proven to efficiently remove nitrogen, and successfully used for treating high ammonia strength landfill leachates. The rotating biological contactor (RBC) technology, an attached-growth biological treatment process, has been found capable of nitrifying and effectively removing ammonia from landfill leachate (Opatken and Bond, 1991; Peddie and Atwater, 1985). In 1993, Henderson investigated the capability of a RBC system to treat leachate with an ammonia concentration of 2000 mg N/L. The system percentage ammonia removal was 97% at an ammonia loading of 1.3 g/m2/day. Nevertheless, at higher loading rates percentage ammonia removal was lower and complete nitrification could not be achieved. A pilot plant 10 study, conducted by Knox (1985), compared the performance of an attached-growth trickling filter process with the performance of a suspended-growth, activated sludge process, when treating leachate with a concentration of up to 500 mg/L ammonia. The attached-growth process was found to be more resistant to fluctuations in hydraulic and organic loading than the suspended-growth process. However, due to the potential of inorganic precipitate formation, that could interfere with substrate transfer to the biomass, the attached-growth system may not always be the most appropriate method of treatment (Forgie, 1988a; Forgie, 1988c). Several different suspended-growth process configurations have been tested for their nitrogen removal potential and used to treat high ammonia strength landfill leachate. Robinson and Barr (1998) utilized extended aerated lagoons, operated in sequencing mode, to nitrify leachate with an ammonia concentration in excess of 1000 mg N/L. A suspended-growth, activated sludge, sequencing batch reactor (SBR) system was employed by Hosomi et al. (1989) to remove nitrogen from leachate containing 200 mg N/L. Robinson and Luo (1991), using the same technology, successfully nitrified ammonia from a leachate with a concentration of up to 2000 mg N/L. Pilot plant trials using nitrification with pre-denitrification process configuration, and nitrification with post-denitrification process configuration, or a combination of the two systems demonstrated that effluent with total nitrogen levels of less than 200 mg N/L may be consistently achieved, when treating leachate with ammonia concentrations of up to 1500 mg N/L (Robinson et al., 1998). The performance of two parallel, continuous flow, complete mix, suspended-growth, single sludge, activated sludge, nitrification with pre-denitrification processes, when treating an influent leachate with an ammonia concentration of up to 2000 mg N/L, were investigated by Azevedo (1993). The processes, generally referred to as Modified Ludzack-Ettinger (MLE) processes (Azevedo, 1993; Guo, 1992; Shiskowski, 1995), relied on a separate reactor for denitrification. At an influent leachate ammonia level of 1500 mg N/L, both systems produced an effluent of approximately 170 mg N/L total nitrogen. An increase from 1500 to 2000 mg N/L in the influent ammonia concentration generated a 80% nitrification decrease in both systems. 11 The biological process configuration selected for this study was the pre- and post-denitrification system (Figure 1.1) generally known as the 4-Stage Bardenpho process, or the 4-Stage Nitrogen Removal process (EPA, 1993). A typical 4-Stage Bardenpho process uses a return sludge recycle rate of 1:1 (EPA, 1993). Based on recent research (Shiskowski, 1995), the return sludge recycle rate chosen for the process investigated in this study was 3:1 (i.e., Clarifier to Anoxic #1 recycle or external recycle). In a 1995 comparative study, Shiskowski found the Bardenpho process capable of producing an effluent containing less than 1 mg N/L of ammonia and about 15 mg N/L of NOx, when treating 1100 mg N/L ammonia leachate. The MLE system, mvestigated in the same study, produced an effluent of 160 mg N/L total inorganic nitrogen concentration, when treating a leachate with an ammonia concentration of up to 1200 mg N/L. Even though the "pre-denitrification" part of the Bardenpho processes is similar to that of the MLE process, the existence of a second anoxic reactor (Figure 1.1) considerably improves its performance (EPA 1993; Shiskowski, 1995). The second anoxic reactor assumes the denitrification of the NOx that is not returned to the first anoxic reactor so that the produced effluent may theoretically be free of NOx. Thorough research studies, conducted at the University of British Columbia, have proven that both these systems are particularly suitable for treating high ammonia concentration landfill leachates and other nitrogenous liquid wastes (Azevedo, 1993; Guo, 1992; Pers. Comm., Mavinic, 1999; Shiskowski, 1995). Therefore, a comparative brief description of their configuration and treatment method, and their advantages and disadvantages over other treatment methods and over each other, seems essential. 1.5 Modified Ludzack-Ettinger (MLE) Process versus Four-Stage Bardenpho Process MLE and Bardenpho systems are combined stage biological nitrification/denitrification processes that rely on separate tanks or reactors for denitrification (EPA, 1993; Metcalf & Eddy, 1991). Both systems are continuous flow, complete mix, suspended-growth, single sludge, activated sludge processes, that consist of a sequence of anoxic and aerobic zones or reactors. The biological nitrification of the influent ammonia that occurs in the aerobic zone is followed by the biological denitrification that occurs in the 12 2 3 0) e o o E © ** (A >i (0 c o n o B e 0 0. 2 fi. 2 3 apuoiqo uinmouiuie 13 anoxic zone. In addition to nitrates (NO3) and nitrites (N02), an organic carbon source has to be present in the anoxic zone of the system for denitrification to take place (EPA, 1993). The MLE system consists of a single anoxic reactor that precedes the aerobic reactor (Figure 1.2a). When entering the anoxic reactor, the influent leachate is diluted by sludge returned from the clarifier. The mixed liquor, containing high levels of ammonia, flows from the anoxic reactor into the aerobic reactor where nitrifying bacteria, through oxidation, convert ammonia to nitrates and nitrites (NOx). Settled sludge, rich in oxidized nitrogen in the form of nitrates and nitrites, is returned from the clarifier to the anoxic zone where denitrifying bacteria reduce NOx to gaseous nitrogen compounds such as nitrogen gas (N2). The Four-Stage Bardenpho system uses a dual anoxic zone that consists of a first anoxic reactor staged before the first aerobic reactor, and a second anoxic reactor staged after the first aerobic reactor (Figure 1.2b). Mixed liquor, rich in nitrates and nitrites, is returned from the first aerobic reactor to the first anoxic reactor where denitrification takes place. When complete nitrification occurs in the first aerobic reactor, the first anoxic reactor NOx source for denitrification is mainly the returned nitrified mixed liquor and not the returned clarifier settled sludge. The second anoxic reactor denitrifies the NOx that is not returned to the first aerobic reactor (Shiskowski, 1995). When complete nitrification of the incoming ammonia takes place in the first aerobic reactor, the function of the second aerobic reactor is mainly to remove the residual organic carbon, and to aerate and strip nitrogen gas from the mixed liquor prior to clarification. When only partial nitrification takes place in the first aerobic reactor, the role of the second aerobic reactor is also to remove the remaining ammonia. Both, MLE and Bardenpho process configurations, offer several advantages over otheT nitrification/denitrification systems (Argaman, 1982; Azevedo, 1993; EPA, 1993; Shiskowski, 1995). The dilution of the influent leachate by the returned sludge (i.e., MLE and Bardenpho) and by the returned mixed liquor (i.e., Bardenpho) reduces the incoming ammonia levels and, consequently, the risk of 14 ANOXIC REACTOR AEROBIC REACTOR CLARIFIER influent /FT O effluent RETURN SLUDGE (External Recycle) (a) Modified Ludzack-Ettinger System Schematic MIXED LIQUOR RETURN (Internal Recycle) < ANOXIC JREACTOR AEROBIC REACTOR influent ANOXIC REACTOR AEROBIC REACTOR RETURN SLUDGE (External Recycle) CLARIFIER effluent (b) Four-Stage Bardenpho System Schematic Figure 1.2: C o m b i n e d Ni t r i f i ca t ion /Den i t r i f i ca t ion S y s t e m s 15 denitrification inhibition. The anoxic reactor staged before the aerobic reactor minimizes the requirements for additional carbon by favoring the usage of the influent organic carbon for denitrification. The removal of organic carbon in the preceding anoxic reactor reduces oxidation in the subsequent aerobic reactor, and therefore minimizes oxygen requirements. The alkalinity addition requirements are considerable reduced since, theoretically, 50% of the alkalinity consumed by nitrification is provided by denitrification. The aerobic reactor staged before the clarifier generates a less noxious effluent and minimizes the risk of rising sludge in the clarifier. The single sludge, combined stage nitrification/denitrification configuration requires only one clarifier. A multiple sludge, separate stage nitrification and denitrification process needs multiple clarifiers. Both, MLE and Bardenpho systems, have flexible designs and may overcome seasonal surges of the leachate hydraulic loading. Seasonal fluctuations of the leachate strength concomitantly with changing characteristics of the leachate with time (i.e. a carbon based leachate or a" young" leachate will eventually become a nitrogen based leachate or an "old" leachate) (Mavinic, 1998) may be also easily accommodated by these two particular type of processes. The major disadvantage of the MLE system is that the NOx effluent concentration may remain unacceptable high, as not all of the nitrified sludge is returned to the anoxic reactor for denitrification. The expected NOx effluent levels can be roughly approximated (i.e., ignoring bacterial assimilation and air stripping) by dividing the influent ammonia concentrations by the sum of 1 and sludge recycle ratio (i.e., external recycle). Therefore, increases in the sludge recycle ratio should theoretically generate decreases in the effluent NOx concentrations. However, Shiskowski (1995) found that sludge recycle ratios higher than 6:1 resulted in higher effluent inorganic nitrogen concentration and decreased process stability. High power costs due to large recycle flows (i.e., when treating high ammonia strength liquid wastes) and high solids loading on the clarifier are other disadvantages associated with the MLE process configuration (Pers. Comm., Mavinic, 1999; Shiskowski, 1995). Shiskowski's study (1995) found that the overall performance of the Bardenpho process is superior to the performance of the MLE process, when treating high ammonia strength leachate. Complete nitrification 16 of the influent ammonia in the first aerobic reactor, followed by the denitrification of the produced NOx in the preceding and subsequent anoxic reactor respectively, may result in an effluent free of nitrogen. The most important disadvantages of the Bardenpho process are the higher capital and operating costs associated with the extra reactors, and the additional necessary pumps and piping. When treating high ammonia landfill leachates characterized by low biodegradable organic levels and low orthophosphate concentrations, both systems, MLE and Bardenpho, require a supplementary source of organic carbon for denitrification and orthophosphate addition for bacterial growth. 1.6 Study Rationale and Objectives Rapidly increasing ammonia concentration in landfill leachates from all over the world prompts further investigation of the treatment methods capable to achieve full ammonia removal. High ammonia leachates such as those from the existing landfills in Hong Kong, where wastes reach methanogenic stage and generate leachate ammonia concentrations in excess of 2000 mg N/L within months of their disposal (Robinson et al, 1998), or those from Germany with concentrations over 1500 mg N/L of ammonia (Rettenberger, 1998), are notorious. Thorough research studies and successful applications (Azevedo and Mavinic, 1993; Robinson andBarr, 1998; Robinson et al., 1998; Shiskowski and Mavinic, 1998a; etc.) have demonstrated that biological nitrification with denitrification process is one of the most effective and economical removal methods of ammonia from high strength landfill leachates. In his 1995 comparative study, Shiskowski found that, under similar operating parameters (i.e., temperature, anoxic/first anoxic actual hydraulic retention time, ammonia loading, etc.), the 4-Stage Bardenpho system investigated produced an effluent of about 16 mg N/L total inorganic nitrogen while the MLE system produced, at best, an effluent of 160 mg N/L nitrogen. The study concluded that the Bardenpho process demonstrated strong capability of ammonia removal from high strength leachates and that the process overall performance should be further investigated under consistently rising influent ammonia concentrations. 17 Based on Shiskowski's research (Shiskowski, 1995; Shiskowski and Mavinic, 1998a; Shiskowski and Mavinic, 1998b), the purpose of the present study was to further explore and evaluate this particular type of pre- and post-denitrification biological process under increasing influent leachate ammonia concentrations. The objectives of the study were: 1. Investigate and evaluate process performance under increasing influent ammonia concentration (i.e., up to a limit of about 2000 mg N/L mcoming ammonia); 2. Improve process stability and performance through operating parameters changes (i.e., required parameters determined by Objective 1); 3. Investigate and evaluate process performance under decreasing ambient temperature (i.e., from an operating temperature of 20 °C to a miriimum of 10 °C). 18 Chapter 2 LITERATURE REVIEW A short overview of the studies and published work related to biological nitrification and denitrification is presented in this literature review chapter. Special consideration is given to research investigating continuous flow, suspended growth activated sludge systems treating high ammonia landfill leachates and other high strength nitrogenous liquid wastes. Emphasis is placed on literature relevant to the present study. 2.1 UBC Research Experience on Leachate Treatment The Environmental Engineering Group of the Department of Civil Engineering at the University of British Columbia (UBC) has more than 25 years of extensive research experience, with over 35 studies, on leachate treatability and management (Pers. Comm., Mavinic, 1999). The leachate used in most of these studies had ammonia concentrations of about 200 mg N/L and biodegradable organic concentrations lower than 50 mg 0 2 /L, characteristics typical for a nitrogen based leachate. The present section is a brief summary of the UBC research on leachate treatment, using continuous flow, combined stage nitrification/denitrification systems, mainly the Modified Ludzack-Ettinger (MLE) process configuration. 2.1.1 Biological Nitrification and Denitrification of Landfill Leachate Landfill leachate containing up to 288 mg N/L of ammonia was successfully nitrified to less than 1 mg N/L by Dedhar and Mavinic (1985), using a MLE system. Reliable ammonia removal rates of 100% were consistently achieved. Using glucose as an external source of organic carbon, denitrification was only occasionally achieved. 2.1.2 Effect of Increased Ammonia Levels on Landfill Leachate Biological Treatment Two continuous flow, complete mix, suspended-growth, single sludge, activated sludge, nitrification with pre-denitrification processes (i.e., MLE processes) were investigated by Azevedo (1993). The leachate 19 ammonia concentrations were artificially increased up to 2000 mg N/L, using ammonium chloride solution. The systems operated with 10 and 20 days, respectively, aerobic solids retention times (SRT). At an influent leachate ammonia level of 1500 mg N/L, both systems produced an effluent of approximately 170 mg N/L total nitrogen. Increases in the influent ammonia concentrations from 1500 to 2000 mg N/L generated a 80% nitrification decrease in both systems. Nitrification failure was attributed to: toxicity of ammonia oxidizing bacteria {Nitrosomonas) to "free" ammonia (NH3), insufficient dissolved oxygen concentration, and sludge foaming problems. Shiskowski (1995) also artificially increased the ammonia concentrations in the leachate being treated by two parallel MLE systems and, subsequentiy, by a 4-Stage Bardenpho system. The MLE systems generated an effluent of 160 mg N/L total inorganic nitrogen concentration, when treating a leachate with a "simulated" ammonia concentration of 1200 mg N/L. The Bardenpho process was capable of producing an effluent containing less than 1 mg N/L of ammonia and about 15 mg N/L of NOx (i.e., nitrite + nitrate), when treating 1100 mg N/L "simulated" ammonia leachate. 2.1.3 Effect of External Organic Carbon Source and Loading on Denitrification Manoharan et al. (1989) explored the suitability of methanol and glucose as supplementary organic carbon sources using the MLE process treating leachate with an average ammonia concentration of 188 mg N/L and a BOD 5 of 25 mg/L. The use of glucose as an additional source of carbon determined an erratic denitrification, with percentage denitrification fluctuations between 10% and 100% as a result of suspected fermentative conditions in the anoxic reactor. Furthermore, the unused glucose flowed from the anoxic zone into the aerobic zone causing nitrification inhibition Stable and reliable nitrification, and consistent complete denitrification where achieved when methanol was used as a supplementary source of carbon. In a similar study conducted by Carley and Mavinic (1991), methanol, acetate, glucose, and brewery yeast were tested as external carbon sources for the denitrification of a nitrogen based landfill leachate. 20 Methanol and acetate were proved to be the most effective carbon sources. For complete denitrification, the required COD to NOx ratios were found to be approximately 6.2 to 1 (mg COD/mg NOx-N) for methanol and 5.9 to 1 for acetate. Good system performance was also achieved by using brewery yeast as external carbon source, but at a much higher COD to NOx ratio (i.e., 8.5 to 1). The usage of glucose resulted in an unstable, unreliable denitrification. 2.1.4 Effect of Low Temperature on Landfill Leachate Biological Treatment The effects of decreasing temperature on biological nitrification and denitrification of a high ammonia landfill leachate, using the MLE processes, were investigated by Guo (1992). The systems were operated with aerobic solids retention times (SRT) ranging from 20 to 60 days, at ambient temperatures of 20 °C, 12 °C, and 4 °C. At a temperature of 4 °C and an operating aerobic SRT of 60 days, the average effluent ammonia concentrations were about 2 mg N/L, when treating incoming ammonia leachate of up to 210 mg N/L. An average effluent ammonia concentration of 9 mg N/L resulted at an aerobic SRT of 20 days. The decrease in temperature prompted the necessity of methanol loading reductions to prevent bleeding of the unused methanol from the anoxic reactor into the aerobic reactors and nitrification inhibition. The methanol loadings were increased, once nitrifying bacteria acclimatized to the decreased temperature. Major nitrification inhibition (i.e., 80% nitrification decrease) as a result of decreased ambient temperature (i.e., from 20 °C to 10 °C) was experienced by Azevedo et al. (1995) when treating "simulated" high ammonia landfill leachate with a concentration of 1500 mg N/L. A 99% to 30% denitrification decrease was observed in the MLE system operated with a 10 day aerobic SRT and a 99% to 82% denitrification decrease in the MLE system operated with a 20 day aerobic SRT. Ceasing anoxic methanol addition and aerobic wasting generated complete nitrification recovery. 2.1.5 Effect of Heavy Metals on Landfill Leachate Biological Treatment Inconsistencies between the results of two similar research works, conducted by Jasper et al. (1985) and Dedhar and Mavinic (1985) respectively, prompted a detailed investigation of the characteristics of the 21 leachate used in these studies. Jasper et al. (1985) observed erratic ammonia removal (i.e., percentage nitrification fluctuations from 7% to 45%) when treating leachate with an average ammonia concentration of 160 mg N/L, using MLE processes with aerobic SRTs of up to 20 days. Dedhar and Mavinic (1985), using identical systems and the same landfill leachate source, achieved 100% ammonia removal from leachate containing about 288 mg N/L of ammonia A thorough investigation of the two leachates compositions found that the leachate used in the first study contained higher concentrations of zinc and manganese than the leachate used in the second study. It was concluded that the elevated levels of zinc and manganese compromised the ammonia removal process in the first study. The heavy metal inhibition of biological nitrification and denitrification of nitrogenous landfill leachates (i.e., concentrations of about 188 mg N/L ammonia) was further explored by Mavinic and Randall (1990). It was observed that an addition of phosphorus in excess determined metal precipitation and that bacterial inhibition occurred only at fairly high mcoming metal concentrations. Operated at 20 °C ambient temperature, with an aerobic SRT of 10 days, the MLE system was able to handle zinc concentrations up to 130 mg/L, but only much lower levels of chromium and nickel. A similar study, conducted by Manoharan et al. (1992), investigated bacterial inhibition to zinc, chromium nickel, and manganese when using nitrification and denitrification as treatment method for a landfill leachate. Even though increased zinc concentrations in the incoming leachate generated nitrification inhibition, similarly to the previous research, increased phosphorus addition determined zinc phosphate precipitation and nitrification recovery. However, it was concluded that the deficiency in biologicaly available phosphorus, due to precipitate formation, was responsible for nitrification inhibition, rather than the exposure of the bacterial population to elevated levels of zinc. Again, similarly to the previous research, when phosphorus was available in excess, the leachate treatability was not affected by up to 130 mg /L of incoming zinc concentrations. Nitrification inhibition occurred at much lower additions of chromium (i.e., 50 mg/L) and nickel (i.e., 2 mg/L) mainly due to metal toxicity. No bacterial inhibition was observed when manganese was added to the mcoming leachate. 22 2.1.6 Effect of Sludge Recycle Ratio on Landfill Leachate Biological Treatment Elefsiniotis et al. (1989) attempted to reduce the NOx effluent concentration of an MLE process, treating incoming leachate ammonia of 240 mg N/L, by increasing the system sludge recycle ratio. At higher than 6:1 recycle ratios (i.e., anoxic actual hydraulic retention time lower than 1.71 hours, and aerobic actual hydraulic retention time lower than 3.42 hours) the system performance was unstable and an effluent of increased NOx concentrations was generated. Larger recycle ratios determined decrease in the reactor actual hydraulic retention times, hence lower contact times. It was concluded that the reduced contact time was not sufficient for complete nitrification and denitrification to occur. Shiskowski (1995) also found that sludge recycle ratios higher than 6:1 resulted in higher effluent inorganic nitrogen concentration and decreased process stability. At a 6:1 sludge recycle rate, an effluent of about 160 mg N/L total inorganic nitrogen concentration was generated by MLE systems treating leachate with a "simulated" ammonia concentration of 1200 mg N/L. An increase from 6:1 to 7:1 in the sludge recycle ratio of one system, and from 6:1 to 8:1 in the sludge recycle ratio of the other system determined effluent nitrogen concentrations of 190 mg N/L and 170 mg N/L respectively. 2.2 Research Experience on Nitrification Performance and Inhibition In a study conducted by Anthonisen et al. (1976), increased levels of nitrous acid ( H N O 2 ) at low pHs and of "free" ammonia (NH3) at high pHs were shown to strongly affect nitrification performance. "Free" ammonia was found to initiate the inhibition of Nitrobacter organisms at much lower concentrations (i.e., 0.1 to 1 mg N/L) than the inhibition of Nitrosomonas organisms (i.e., 10 to 150 mg N/L), so that the overall result was nitrite ( N O 2 ) accumulation. For both organisms, nitrous acid inhibition started at concentrations between 0.22 and 2.8 mg N/L. Temperature, the number of existent nitrifying organisms, and acclimatization were acknowledged to also influence nitrous acid and free ammonia nitrification inhibition. 23 The effect of temperature on nitrification was investigated by Randall and Buth (1984) using a synthetic liquid waste. The results showed that biological nitrification was very sensitive to even small temperature decrease steps between 17 °C and 10 °C. Nitrite accumulation within the process led to the conclusion that temperature inhibition of nitrate (NO3) former organisms (i.e., Nitrobacter, nitrite oxidizers) was stronger than that of nitrite (N0 2) former organisms (i.e., Nitrosomonas, ammonia oxidizers). Furthermore, nitrifiers were found to be more temperature sensitive than heterotrophs. Turk and Mavinic (1989) explored process operating parameters changes to maintain nitrite accumulation and prevent bacterial automatization to "free" ammonia while nitrifying and denitrifying a landfill leachate and a synthetic waste. Consistent nitrite accumulation was thought as means of reducing organic carbon denitrification requirements when treating high strength nitrogenous liquid wastes. Investigated parameters included "free" ammonia, nitrous acid, and dissolved oxygen. Unacclimatized nitrite oxidizing bacteria were inhibited by "free" ammonia concentrations of 5 to 10 mg N/L. Changes in nitrous acid and dissolved oxygen concentrations did not seem to stimulate nitrite accumulation. Furthermore, acclimatized ammonia and nitrite oxidizing bacterial population were able to handle up to 40 mg N/L of "free" ammonia. Therefore, it was suggested that the most efficient way of delaying acclimatization, thus encouraging nitrite accumulation, would be the continuos nitrifiers recycle through elevated "free" ammonia concentrations such as those of the anoxic reactor of a pre-denitrification process. In his study, using a pre-denitrification activated sludge system, with a 4:1 solids recycle ratio, treating landfill leachate, Carley (1988) found that a COD:NOx ratio of 20:1 resulted in carbon bleeding into the aerobic reactor and generated nitrification decrease of up to 40%. It was speculated that, in the presence of elevated organic carbon levels, faster-growing heterotrophic bacteria outcompeted slow-growing autotrophic nitrifiers for dissolved oxygen, hence, detennined nitrification inhibition. 24 Hanaki et al. (1990a) investigated nitrification inhibition due to enhanced heterotrophic activity within a complete mix aerobic reactor, without solids recycle. Increased influent organic carbon loading (i.e., . glucose) resulted in decreased ammonia oxidation. Ammonia oxidation was further inhibited by decreasing hydraulic retention times at given organic carbon loads. Higher organic carbon levels stimulated heterotrophic growth while lower hydraulic retention times hindered nitrifiers growth. It was assumed that growing heterotrophic population "crowded" nitrifying bacteria and obstructed ammonia and oxygen transfer to the ammonia oxidizers organisms within the floes. The assumption was confirmed (Hanaki et al., 1990b) when decreased dissolved oxygen levels were found to ftirfher enhance the inhibitory effect of increased carbon loading on ammonia oxidation. 2.3 Research Experience on Denitrification Performance and Inhibition Treating tannery waste using an MLE system with a four-stage anoxic reactor (i.e., to simulate plug flow process conditions), Panzer et al. (1981) demonstrated the importance of the anoxic carbon levels on denitrification performance. The first stage of the anoxic zone achieved specific denitrification rates up to four times higher than the third and the fourth anoxic stages, due to higher organic carbon availability. Large anoxic removal rates when using an aerobic mixed liquor recycle ratio of 30:1 were found by Argaman (1982). The calculated anoxic BOD 5 to NOx removal rate ratio was 29.2 to 1. It was speculated that, even though the anoxic dissolved oxygen levels were about 0.5 mg/L, aerobic oxidation of organic carbon occurred concurrently with nitrate and nitrite reduction within the anoxic reactor. Elevated nitrite (N0 2) levels were suspected by Beccari et al. (1993) to have caused the denitrification inhibition in a suspended-growth process. Jain et al. (1992) in a pre-denitrification modeling study found that the performance of anoxic denitrification would be negatively affected by the molecular oxygen introduced with the aerobic mixed 25 liquor recycle. The negative effect of oxygen on denitrification would be enhanced by higher mixed liquor recycle rates, not only due to the introduction of larger oxygen amounts into the anoxic zone, but also due to decreased anoxic hydraulic retention times. 26 Chapter 3 EXPERIMENTAL SETUP AND OPERATION A nitrification with pre- and post-denitrification process configuration was used in the present study. The process configuration, generally known as 4-Stage Bardenpho process, is a continuous flow, complete mix, suspended-growth, single sludge, activated sludge processes, that relies on a sequence of anoxic and aerobic zones for biological nitrification and denitrification (Figure 1.1). Two identical, parallel, laboratory scale systems were run during this study. Based on previous results (Shiskowski, 1995), both systems were operated at an aerobic solids retention time (SRT) of 20 days and a sludge recycle ratio (herein, external recycle or Clarifier to Anoxic #1 recycle, interchangeable) of about 3:1. Throughout the study, one system (System #1) used a mixed liquor recycle ratio (herein, internal recycle or Aerobic #1 to Anoxic #1 recycle, interchangeable) of about 4:1, while the other system (System #2) used a recycle ratio of 3:1. The recycle ratios were set so that an average first anoxic (Anoxic #1) actual hydraulic retention time (AHRT) of 1.5 hours for System #1, and of 1.7 hours for System #2 could be maintained during the entire study period. The first aerobic (Aerobic #1) AHRT was of about 3 hours for System #1, and of 3.4 hours for System #2. The differences of 0.4 hours between the Aerobic #1 AHRTs of the systems and of 0.2 hours between the Anoxic #1 AHRTs of the systems, were expected to determine significant different performances of the systems, while mamtaining sufficient contact times for adequate nitrification and denitrification. Previous results (Shiskowski, 1995), showed that a decrease of about only 0.2 hours in the aerobic AHRT of an MLE system (i.e., from 3.42 hours to 3.18 hours by changing the sludge recycle ratio from 6:1 to 7:1) generated decreased nitrification stability and increased effluent ammonia concentration. Anoxic #1 AHRT of 1.64 hours and Aerobic #1 AHRTs of 3.17 hours, were found sufficient for adequate nitrification and denitrification, when a 4-Stage Bardenpho system was investigated (Shiskowski, 1995). 3.1 Treatment System Design and Operating Parameters The Environmental Room (i.e., controlled-temperature room) of the Environmental Laboratory of the Department of Civil Engineering at UBC was used for the setup of the systems (Plate 3.1). The 27 temperature was maintained at 20 °C during the most part of the study (i.e., study Objective 1 and Objective 2) and was progressively decreased from 20 °C to 10 °C at the end of the study period (i.e., study Objective 3). Both systems consisted of an anoxic reactor (Anoxic #1), an aerobic reactor (Aerobic #1), a second anoxic reactor (Anoxic #2), followed by a second aerobic reactor (Aerobic #2) and then a clarifier (Figure 1.1). The average design operating parameter values of the systems are summarized in Table 3.1. T a b l e 3 .1 : S y s t e m D e s i g n a n d Ope ra t i ng P a r a m e t e r s OPERATING PARAMETER SYSTEM #1 SYSTEM #2 (Mean Value) (Mean Value) Anoxic #1 Volume (L) 5 5 Aerobic #1 Volume (L) 10 10 Anoxic #2 Volume (L) 5 5 Aerobic #1 Volume (L) 10 10 Clarifier Volume (L) 4 4 System Volume (L) 35* 35* Influent Flow (L)** 10 10 External Recycle Ratio 3:1 3:1 Internal Recycle Ratio 4:1 3:1 Aerobic Wasting (L/day) 0.5 0.5 Aerobic SRT (days) 20 20 Anoxic #1 Nominal HRT (hours) 12 12 Aerobic #1 Nominal F1RT (hours) 24 24 Anoxic #2 Nominal HRT (hours) 10 10 Aerobic #1 Nominal HRT (hours) 24 24 Clarifier Nominal HRT (hours) 9.6 9.6 System Nominal HRT (hours) 84 84 Anoxic #1 Actual HRT (hours) 1.5 1.7 Aerobic #1 Actual HRT (hours) 3 3.4 Anoxic #2 Actual HRT (hours) : 3 3 Aerobic #2 Actual HRT (hours) 6 6 Clarifier Actual HRT (hours) 2.4 2.4 *1 L is estimated for pumps and tubing in each system. **Refers to Anoxic #1 Total Flow, including chemical feed flows into Anoxic #1 (see Appendix A). The systems were constructed so that mixed liquor from each antecedent reactor flowed by gravity into each subsequent reactor (Plate 3.1). 28 The first anoxic reactor (Anoxic #1) function was to denitrify the mixed liquor, rich in nitrite and nitrate(NOx), returned from the first aerobic reactor (Aerobic #1) through the internal recycle line. The dilution of the influent leachate, mainly by the internal recycle and external recycle flows, occurred also in the Anoxic #1 reactor. The reactor, with a volume of 5 L, was constructed from a plastic container with a cylindrical shape. An electrical motor with a stirring rod was used to maintain the mixed liquor suspended and completely mixed. The provided mixing speed of 60 rotations per minute (rpm) limited the entrainment of molecular oxygen while supplying adequate mixing within the reactor. The oxidation-reduction potential was continuously monitored by a submerged ORP probe connected to a millivolt meter. The nitrification of the mixed liquor from the first anoxic reactor, containing elevated ammonia levels, occurred in the first aerobic reactor (Aerobic #1). The reactor had a volume of 10 L and was constructed also from a plastic container with cylindrical shape. The air required for nitrification was supplied using a couple of small porous stone diffusers, placed at the bottom of the reactor. The diffusers were connected the laboratory compressed air supply line and the air flow was adjusted using intermediary flow meters. Dissolved oxygen concentrations of minimum 2 mg/L required for adequate nitrification performance (Mavinic, 1997) were consistently maintain. A submerged DO probe connected to a dissolved oxygen meter continuously monitored reactor oxygen levels. Mixing was provided using an electrical motor and stirring rod. A mixing speed of 30 rmp provided adequate uniform distribution of the oxygen bubbles within the reactor and kept the mixed liquor in suspension. The second anoxic reactor (Anoxic #2) was identical to the first anoxic reactor (Anoxic #1). The Anoxic #2 reactor function was to perform additional denitrification on the nitrified mixed liquor passed from the Aerobic #1 reactor. P l a t e 3 .1 : S y s t e m #1 S e t u p in the E n v i r o n m e n t a l R o o m 30 31 The role of the second aerobic reactor (Aerobic #2) was mainly to reaerate and strip nitrogen gas from the mixed liquor prior to clarification, hence an actual hydraulic retention time (AHRT) of 0.5 hours would have been sufficient (EPA, 1993). However, speculating that, at high mcoming ammonia concentration, ammonia, that may have escaped treatment in the first aerobic reactor, could have been treated in the second aerobic reactor, providing the existence of a long enough AHRT, Aerobic #2 reactor had a volume of 10 L and an AHRT of about 6 days. Similarly to the Aerobic #1 reactor, the Aerobic #2 was constructed from a plastic container, mixing was provided using an electrical motor and stirring rod, and the air was supplied using two diffusers connected to the laboratory air supply installation. A 4 L volume Plexiglas clarifier, with a conical bottom, was used to settle solids from the final effluent and produce a clear effluent supernatant. Aerobic #2 mixed liquor flowed into the clarifier through an inner, cylindrical baffle used to prevent short-ckcuiting. The thickened sludge from the bottom of the clarifier was returned to the Anoxic #1 reactor using a peristaltic pump. The pump operated intermittently at 1 minute on and 9 minutes off in order to prevent recycle line blockages (Shiskowski, 1995). The sides and the bottom of the clarifier were continuously swept using a stainless steel rod attached to a 1 rpm electric motor (herein, scraper). A peristaltic pump operating at 1 minute on and 9 minutes off was used to return the nitrified mixed liquor from Aerobic #1 to Anoxic #1 for denitrification. 3.2 Leachate Feed In order to maintain consistency with previous studies (Guo, 1992; Azevedo, 1993; Shiskowski, 1995), the leachate used in this study was from Burns Bog Landfill located in Delta, British Columbia. The leachate was collected from a wet well situated at the southwest corner of the landfill, using a submergible pump. Leachate collection took place every second week, using 15 plastic containers, each with a volume of 20 L. Collected leachate was stored at 4 °C to minimize changes in its characteristics. 32 The landfill, still in use to day, began operation in 1966 (Guo, 1992). The leachate generated at present has the typical characteristics of an "old" leachate or a nitrogen based leachate (i.e., low levels of readily biodegradable organic matter and high concentrations of ammonia). A summary of the basic composition of the leachate used in the present study is shown in Table 3.2. T a b l e 3 .2: B u r n s B o g L e a c h a t e C o m p o s i t i o n PARAMETER* CONCENTRATION RANGE MEAN CONCENTRATION COD (mg/L) 207-440 311 BOD 5 (mg/L) 25-84 46 Ammonia** (mg N/L) 50-310 149 NOx** (mg N/L) 0.00-1.71 0.53 N0 2 ** (mg N/L) 0.00-0.58 0.20 O-PO4** (mg P/L) 0.00 - 3.40 0.32 Suspended Solids (mg/L) 49 - 150 81 Alkalinity** (mg CaC0 3/L) 1020-2290 1513 pH (pH units) 7.30 - 7.81 7.53 Total Aluminum (mg/L) <0.2 - 0.4 <0.2 Total Cadmium (mg/L) <0.025 <0.025 Total Chromium (mg/L) <0.03 - 0.05 O.03 Total Cobalt (mg/L) <0.02 <0.02 Total Cooper (mg/L) <0.02 <0.02 Total Iron (mg/L) 7.62 - 34.2 16 Total Lead (mg/L) <0.08 <0.08 Total Manganese (mg/L) 0.79-1.41 1.1 Total Molybdenum (mg/L) <0.04 <0.04 Total Nickel (mg/L) <0.03 - 0.06 <0.03 Total Zinc (mg/L) 0.04 - 0.27 0.1 Total Sulfides (mg/L) <0.05 -0.31 0.09 Total Phenol (mg/L) 0.010-0.150 0.043 Total Cyanide (mg/L) <0.01 -0.10 0.05 Calculated Toxicity (%) 9-25 16 *Based on: City of Vancouver (1999), Vancouver Landfill Water Quality, Location 99-Pump Station, Monthly Composite Data, April 1998-March 1999. **Based on. Appendix B, C, D, and E, Influent (Leachate) Characteristics Data, September 1998 - July 1999. A double-headed peristaltic pump continuously pumped leachate from a covered, plastic pail, with a volume of 60 L, into the first anoxic reactors of both systems. An electrical motor and stirring rod was used to slowly mix the leachate without entraining oxygen into the leachate pile (or feed pile, 3 3 interchangeable). Leachate was allowed to came to the room temperature (i.e., from 4 °C in the storage room to ambient temperature in the controlled -temperature room) 24 hours before its discharge into the feed pile. In order to reduce aeration of leachate and minimize possible nitrification occurrence within the feed pile, the leachate was transferred into the pile using a siphon. The leachate flow rate was set to about 9 L/day, so that the total flow into the first anoxic reactor, including chemical addition flows, could be maintained to approximately 10 L/day. 3.3 Chemical Feed Solutions, in tap water, of the various required chemical additions were pumped from graduated cylinders into the systems using peristaltic pumps. In order to minimize the effect of the chemical solution additions on the systems hydraulic retention times (HRTs), the pumps were set at the lowest flow rates that could consistently be maintained (i.e., 10 ml/hour to 20 ml/hour) and the necessary adjustments to chemical loadings were done by changing the concentrations of the feed solutions. Daily mass loadings were calculated based on hourly volumetric changes within the graduated cylinders and solution chemical concentration. Phosphorus solutions were fed from separate, 1000 ml graduate cylinders to the first anoxic reactor of each system. Solution concentrations were adjusted so that a minimum concentration of 2 mg P/L of biological available ortho-phosphate, considered sufficient for nitrification and denitrification (Manoharan et al., 1992), would be ensured within all reactors. Tribasic sodium phosphate (Na3P04*12H20) was used to prepare the phosphorus solution. Methanol solution was added to both anoxic reactors of each system, to provide the organic carbon levels required for denitrification. Methanol, methyl alcohol, or wood alcohol ( C H 3 O H ) is a colorless, flammable liquid and the simplest alcohol. It used to be obtained mainly from wood, but to day is made synthetically from the direct combination of hydrogen and carbon monoxide gases (Nenitescu, 1980 and 34 1985). Methanol is a reliable source of supplementary organic carbon and has been widely used successfully, especially in landfill leachate treatment (Azevedo, 1993, Shiskowski, 1995, Robinson et al., 1998). Similarly to phosphorus solutions, methanol solutions were pumped into the anoxic reactors from separate graduate cylinders, with a volume of 1000 ml each. Increased ammonia loading levels were simulated by adding ammonium chloride (NH4CI) solution to the first aerobic reactors of the systems. Desired increase in the "simulated" leachate ammonia concentration was done by changing the concentration of the ammonium chloride solution, stored separately in 1000 ml graduated cylinders for each system. Sodium bicarbonate (NaHC03) solution was feed into the first aerobic reactor of each system to provide adequate alkalinity for nitrification. Cole Parmer 7142 pH/pump controllers were used to continuously monitor Aerobic #1 reactors pH values (i.e., initial pH set point of 7.5) and to pump bicarbonate solutions from the 2000 ml graduated cylinders into the reactors whenever the pH levels dropped bellow the controllers pH set point. The solutions were maintained near saturation (i.e., 80 g NaHC0 3/L) in order to reduce the effects of the sodium bicarbonate volumetric additions on the Aerobic #1 reactors actual hydrauUc retention time (AHRT). 3.3 System Startup and Operation On September 2, 1998, the anoxic reactors and the clarifiers of both systems were filled with anoxic mixed liquor from the Biological Phosphorus Removal (Bio-P) Pilot Plant of the University of British Columbia. Similarly, the aerobic reactors were filled with aerobic sludge from the same source. Leachate and phosphorus solutions were fed into the Anoxic #1 reactors of the systems. Methanol solutions were added separately to the Anoxic #1 and Anoxic #2 reactors of the systems* as a function of each reactor estimated denitrification requirements. External recycle line ratios were set at 3:1 for both systems while internal recycle ratios were 4:1 for System # 1 and 3:1 for System #2. The ambient temperature was set to 20 °C. 35 The systems produced ammonia free effluents by September 21, 1998 (i.e., study day no. 20). Aerobic #1 wasting started on October 6, 1998 (i.e., study day no. 35) at a rate of 0.5 L/day for both systems. Methanol loadings were adjusted to match nitrites and nitrates (NOx) production, meet denitrification requirements, and produce effluents free of nitrogen. Stable and complete nitrification and denitrification of the "base" leachate (i.e., landfill leachate) and NOx effluent concentrations of under 2 mg N/L were consistently achieved, by both systems, by day 64 of the study (i.e., November 4,1998). Total inorganic nitrogen (UN) removal percentages of about 99% were reached starting with the 52n d day of the study (i.e., October 23, 1998). Influent ammonia levels were artificially increased, using ammonium chloride, from the concentrations contained by the landfill leachate (herein, base leachate, in keeping consistency with 1995 Shiskowski's study) to ammonia concentrations of over 2200 mg N/L (i.e. simulated leachate concentrations). During the first attempt at incrementally increasing the incoming ammonia concentrations (i.e., between November 6, 1998 and December 2,1998), methanol loadings were increased concomitantly with ammonia loading increases. Based on the results from the beginning of the study (i.e., first period or phase 1 of the study) the CH 3OH:NOx ratios were kept at about 20:1. By day 27 of the study phase, strong nitrification inhibition and ammonia accumulation were experienced within both systems and the ammonium chloride addition was stopped. The second attempt to increase the influent ammonia concentrations started on December 7, 1998 (i.e., phase day no. 32). The systems were allowed to adjust to each incremental ammonia concentration increase and only afterwards methanol loadings were increased to meet denitrification requirements. Furthermore, anoxic methanol loadings were increased in a more conservative manner. By day 120 (i.e., March 5, 1999) of the study phase, both systems consistently generated effluents virtually ammonia free when treating simulated ammonia leachate concentrations of about 2200 mg N/L. System #1 effluent had a NOx concentration of about 80 mg N/L while the System #2 effluent had one of about 300 mg N/L. TIN removal percentages were over 95% for System #1 and over 85% for System #2. 36 Hypothesizing that lower pH levels in the anoxic reactors of both systems (i.e., a pH as high as 9.6 was reached in the Anoxic #2 reactor of System #2 during previous study phase) would improve denitrification, an incrementally decrease in the controller pH set point of the Aerobic #1 reactors started on March 6, 1999. After 7 days (i.e., March 12, 1999), at Anoxic #2 pH values of about 9.2, the systems were generating ammonia free effluents with NOx concentrations of about 60 mg N/L and 35 mg N/L respectively. TIN removal percentages of both systems were over 95%. On March 17, 1999 incipient denitrification inhibition was observed as the systems started to generate increased NOx effluent concentrations. By April 12, 1999 both systems experienced massive nitrification failure. The composition of the last couple of landfill leachate batches was one of the factors suspected for the systems failure. On April 13, 1999, the simulated ammonia concentrations were decreased to about 1200 mg N/L and the systems were fed from a new batch of leachate (i.e., collected on the same day). Nitrification recovered, and by April 19, 1999, both systems were treating mcoming ammonia levels of about 2200 mg N/L. However, denitrification did never completely recover and by May 19,1999, both systems were producing ammonia free effluents, but with mean NOx concentrations of about 150 mg N/L. On May 19, 1999, the ambient temperature was decreased from 20 °C to 17 °C. Over 24 hours, an approximate 15% denitrification decrease was experienced in Anoxic #1 reactors of the systems while no nitrification inhibition was observed. The temperature was subsequently decreased from 17 °C to 14 °C, and finally to 10 °C. Starting with the temperature of 14 °C, Aerobic #1 wasting was ceased and methanol loading levels were progressively decreased. By June 9, 1999, both systems suffered massive nitrification and denitrification inhibition. Changes in operating parameters, such as decreased ammonia and methanol loadings, and increased ambient temperatures (i.e., from 10 °C to 15 °C on July 5, 1999), did not considerably improve systems performance. On July 9, 1999, the 51 s t day of the study phase (i.e., study day no. 311) both systems were shutdown. 37 Chapter 4 ANALYTICAL METHODS The analysis of various constituents of the leachate, mixed liquors and effluent supernatants, and the measurements of the parameters of interest were performed in accordance with the Standards Methods for the Examinations of Water and Wastewater (APHA, 1993) and the instrument corresponding instruction manuals. Plastic bottles, with a volume of 250 ml each, were used to collect samples two to three times a week. 4.1 Total Ammonia ( N H / + NH 3) The terms "ammonia'' and "total ammonia", in this report, refers to the sum of ammonium ion (NH/) and "free" ammonia (NH3). Other works refer to the sum as ammoniacal-N or, simply, NH4. A Lachat Quikchem Automated Ion Analyzer was used to measure the ammonia levels of leachate, mixed liquors, and effluents. Samples were filtered through Whatman #4 filters, preserved using concentrated sulfuric acid (i.e., added till sample altered pH was about 3), and stored in small plastic tubes, at 4 °C, until analysis. The procedure outlined in the Methods Manual for the Lachat Quikchem Automated Ion Analyzer (1987) was followed to determine the ammonia concentrations of interest. Sampling for ammonia analyses was performed every second or third day. An Orion ammonia electrode (Model 95-10) connected to a Cole Parmer Chemicadet Series 5984 pH/mV meter provided fast scanning and immediate estimation of ammonia levels of interest. Samples were prepared and measurements were taken in accordance with the Orion Ammonia Electrode Model 95-10 Instruction Manual. Calibration curves were prepared using four ammonium chloride standardizing solutions. Standardizing solutions were obtained by serial dilution of the 1000 ppm standards to 100 ppm, 10 ppm and 1 ppm standards. Before using, the ammonia standards were adjusted to pH 11 by adding 1 ml of 10 M NaOH solution to each 100 ml of standardizing solution. The ammonia electrode was placed 38 in each standard and each mV reading was taken and recorded. Milhvolt readings (linear axis) were plotted against ammonia concentrations (log axis) on a semilogarithmic scale. Before reading, filtered (Whatman #4) 100 ml samples were prepared similarly to the standards. Sample mV readings were recorded and sample ammonia concentrations were calculated using the calibration curves. Direct ammonia measurements using the ammonia electrode were performed when the systems operated at decreased ambient temperatures. 4.2 Nitrite Plus Nitrate (NOx) NOx concentrations of the leachate, mixed liquors, and effluents were determined using the Lachat Quikchem Automated Ion Analyzer, in accordance with the procedure outlined in the Methods Manual for the Lachat Quikchem Automated Ion Analyzer (1987). Filtered samples (Whatman #4) were preserved with several drops of phenyl mercuric acetate solution and stored in plastic tubes, at 4 °C, until analysis. Sampling for NOx analyses was performed every second or third day. 4.3 Nitrite (N02) Similarly, N 0 2 concentrations of the leachate, mixed liquors, and effluents were detenriined using the Lachat Quikchem Automated Ion Analyzer, in accordance with the corresponding procedure outlined in the Methods Manual for the Lachat Quikchem Automated Ion Analyzer (1987). Filtered samples (Whatman #4) were preserved with several drops of phenyl mercuric acetate solution and stored in plastic tubes, at 4 °C, until analysis. Sampling for nitrite analyses was performed every second or third day. 4.4 Orthophosphate The Lachat Quikchem Automated Ion Analyzer was used to measure the orthophosphate levels of the leachate, mixed liquors, and effluents. Samples were filtered through Whatman #4 filters, preserved using a few drops of phenyl mercuric acetate, and stored in small plastic tubes, at 4 °C, until analysis. The orthophosphate concentrations were determined following the procedure outlined in the Methods Manual 39 for the Lachat Quikchem Automated Ion Analyzer (1987). Sampling for orthophosphate analyses was performed every second or third day. 4.5 Suspended Solids (TSS and VSS) Leachate, mixed liquor and effluent samples were analyzed for total suspended solids (TSS) levels and volatile suspended solids (VSS) levels, one to three times a week. Samples were filtered through glass microfiber filters using a stainless steel filtration apparatus. The filters were placed in aluminum foil filter holders and dried at 103 °C to determine TSS levels and fired at 550 °C to determine VSS levels. Using the filtration apparatus and the aluminum filter holders instead of the ceramic Gooch crucible filtration units, minimized the possibility of errors associated with moisture absorption. 4.6 Chemical Oxygen Demand (COD) COD concentration of the leachate, mixed liquors, and effluents were determined using the closed reflux colorimetric method (APHA, 1993). Samples were filtered through Whatman #4 filters, acidified to a pH of about 2 and stored in plastic bottles, at 4 °C, until analysis. Two milliliters of each sample were pipette into separate vials containing 1.2 ml digestion solution (i.e. with mercuric sulfate to suppress chloride interference) and 2.8 ml of reagent solution. The vials were digested for 2 hours using a HACH block digestor. After digestion, vials were occasionally centrifuged to settle eventual precipitates. Readings were taken using a HACH-DR/200 Direct Riding Spectrophotometer. COD concentrations of the samples were corrected for N 0 2 interference by subtracting 1.1 mg 0 2 /L for each mg N/L of N 0 2 (APHA, 1993). COD tests were conducted one to three times a week. 4.7 Biochemical Oxygen Demand (BOD5) Biochemical Oxygen Demand levels of the leachate, mixed liquors, and effluents were determined about once a month. Samples were filtered through Whatman #4 filters prior to direct pipetting into the 300 ml BOD bottles. Generally, three sample sizes were used in order to cover wider concentration ranges. The BOD bottles were filled with dilution water. HACH Formula 2533 nitrification inhibitor at a 40 concentration of 10 mg/L, and nutrients were added to the dilution water. The dilution water was seeded with 1 ml of aerobic mixed liquor per 10 L of dilution water. A digital dissolved oxygen meter and a self mixing dissolved oxygen probe were used to measure the initial and final oxygen values within the BOD bottles. 4.8 Alkalinity Leachate alkalinity tests were conducted on each batch of leachate collected from the landfill. Mixed liquors were occasionally checked for alkalinity during nitrification inhibition periods. Fifty milliliter samples were filtered trough Whatman #4 filters and titrated to a pH point of 4.5 using 0.02 N sulfuric acid solution (H 2S0 4). During titration, sample pH levels were continuously monitored using an Ag/AgCl combination pH electrode connected to a digital pH meter. 4.9 pH Leachate, mixed liquors, and effluent pH levels were measured on a daily base, using an Ag/AgCl combination pH electrode connected to a portable pH meter. The probe was submerged directly into the leachate from the feed pile, mixed liquors of the reactors, and supernatant of the clarifier. The pH meter calibration was checked before each set of measurements, using a pH 7. buffer solution. A 3-point pH meter calibration was performed, whenever necessary, using pH 7, pH 4, and pH 10 buffer solutions. Cole Parmer Model 7142 pH/pump controllers with Ag/AgCl combination pH electrodes continuously monitored Aerobic #1 reactors pH levels. The controllers were calibrated using pH 7 and pH 10 buffer solutions. The controllers initial pH set point was 7.5. Due to the consistent background noise, frequent differences between the pH values displayed by the controllers and the Aerobic #1 pH levels measured with the portable pH meter were experienced. In addition, the controllers were not equipped to compensate for decreased temperatures. Therefore, the pH levels determined using the portable pH meter were included in the data collection. 41 4.10 Oxidarion-Reduction Potential (ORP) Submerged Broadley-James ORP electrodes, connected to Cole Parmer Chemicadet Series 5984 pH/mV meters, continuously monitored the oxidation-reduction potential within the Anoxic #2 reactors of the systems. The ORP values of the Anoxic #2 reactors were measured every single day. The electrodes were tested using pH 7 and pH 4 buffer solutions saturated with quinhydrone. The millivolt meter was set to display 86 mV with the probe submerged in the pH 7 saturated solution, and was checked for displaying a 263 mV value with the probe submerged in the pH 4 saturated solution. Probes were tested on a regular base using the same procedure (Broadley James Corporation Electrode Specifications). 4.11 Dissolved Oxygen (DO) Dissolved oxygen levels of the Aerobic #1 reactors were continuously monitored using a Yellow Springs Instrument Company Model 54 Dissolved Oxygen Meters with a Yellow Springs Instrument Company Model 5739 Dissolved Oxygen submersible probes. DO levels of the Aerobic #2 reactors of the systems were measured on a daily base. The DO meter was calibrated following the instructions outlined on the back of the instrument. Cahbration was also performed as a function of the ambient temperature. The probe membrane was changed whenever the instrument failed cahbration. 4.12 Temperature The present study was conducted in a temperature controlled:-room. The room temperature was set using the built-in temperature controller. Two additional alcohol thermometers, one placed in the middle of the room and the other one placed at one side of the room, were used to continuously monitor ambient temperature. Mixed liquors temperatures were occasionally measured after each decrease in the ambient temperature. Chapter 5 RESULTS AND DISCUSSION 42 Two laboratory scale, nitrification with pre- and post-denitrification, continuous flow, complete mix, suspended-growth, single sludge, activated sludge systems, generally referred to as Four-Stage Bardenpho process, were used in the present study (Figure 1.1). During the entire study period, the recycle ratios of the systems were adjusted so that mean first anoxic (Anoxic #1) actual hydraulic retention times (AHRT), at average influent flows of 10 L/day (i.e., "base" leachate plus chemical feed into Anoxic #1), were about 1.5 hours for System #1, and 1.7 hours for System #2. The study was conducted in four phases: (1) Base Leachate Phase; (2) Ammonia Loading Phase; (3) pH Phase; and (4) Temperature Phase. The ambient temperature was set to 20 °C during the first three phases of the study, and was varied between 20 °C and 10 °C during the last phase of the study. Formulas and calculation definitions, as well as, explanatory notes are presented in Appendix A. Operating temperature, leachate characteristics, and systems raw and calculated data, of each phase, are recorded in Appendix B. The goal of the Base Leachate Phase was to achieve stable nitrification and denitrification in both systems, and establish optimum operating parameters of the systems, when treating landfill leachate (herein, "base" leachate or base leachate, interchangeable) with average ammonia concentrations of 200 mg N/L and BOD 5 levels of 50 mg/L (City of Vancouver, 1999). Once complete and stable nitrification and denitrification of the base leachate were achieved, influent ammonia concentrations were artificially increased to simulate a leachate (herein, "simulated" leachate or simulated leachate, interchangeable) with an ammonia concentration of up to about 2200 mg N/L. The performances of the two systems, under incrementally increasing ammonia loads up to the 2200 mg N/L influent concentration, were investigated during the second phase of the study, the Ammonia Loading Phase. 43 The third phase, the pH Phase, started when, under influent incoming ammonia concentrations of 2200 mg N/L, stability in the performance of the two systems was observed. Considering that the optimal maximum pH value recommended for denitrification is 8 (EPA, 1993; Metcalf & Eddy, 1991), the aim of the pH Phase was to improve denitrification performance by decreasing, from as high as 9.6 to below 9.2, the pH values in the second anoxic zones of the systems. The final phase of the study, the Temperature Phase, investigated the effect of decreased ambient temperature on the performances of the two systems. The phase ended once nitrification and denitrification within both systems failed. 5.1 Base Leachate Phase The Based Leachate phase involved the initial startup of the systems and the establishment of complete nitrification and denitrification of the base leachate. In addition, the goal of the first phase was also to determine optimal operating parameters (i.e., required methanol additions loads and optimal aerobic dissolved oxygen levels, mainly) and systems means of removing ammonia from landfill leachate (i.e., investigated the occurrence and contribution of air stripping and bacterial assimilation, to the overall performance of the systems, as ammonia removal mechanisms, in addition to nitrification and denitrification). The first phase started with the initial startup of the systems on September 2, 1998, and lasted for 65 days, till November 6, 1998. With the exception of ammonium chloride addition, the systems were fed all required chemical addition from the start. After a period of bacterial acclimatization to the new feed (i.e., leachate instead of sewage), of about seven days, dissolved oxygen levels of the first aerobic zones were adjusted so that the mean concentrations, measured at half reactor depths, were about of 4 mg /L. Occasional dissolved oxygen measurements showed that a minimum concentration of 2 mg/L of oxygen at the bottom of aerobic reactors, and less that 0.5 mg/L oxygen in the anoxic reactors could be reasonably ensured in this condition. 44 5.1.1 Ammonia Levels and Removal Both systems were generating ammonia free effluents by day 20 of the study phase. As shown in Figures 5.1 and 5.2, the influent ammonia concentrations were diluted in the Anoxic #1 reactors, by the recycle lines flows and chemical additions flows to, about 8 times and 7 times, respectively, smaller concentrations. The actual ammonia concentrations in the Anoxic #1 reactors were, most of the time, smaller than the expected concentrations of the reactors (see Appendix A for calculation definition). The consistent difference between the two values showed that ammonia removal in the first anoxic zone occurred on a regular base. Shiskowski et al. (1998b) observed similar first anoxic zone ammonia removal when treating higher strength leachate. He assumed that the ammonia removal was the result of assimilation, under anoxic conditions, by heterotrophic denitrifying organisms. Anoxic #1 ammonia concentration lines of Figures 5.1 and 5.2, respectively, show that by day 52 of the phase, both systems had stable performances, maintaining virtually constant ammonia concentrations in the first anoxic reactors, under slightly variable influent ammonia concentrations. Anoxic #1 ammonia concentrations were constant, even though percentage ammonia removal of both systems was erratic and fluctuated between values of about 8% and 20% (see Base Leachate Phase, Appendix B). Therefore, it was speculated that, in addition to heterotrophic denitrifiers assimilation, a second ammonia removal mechanism, regulated by anoxic ammonia concentrations, was triggered. The presence of increased ammonia concentrations in addition to high organic carbon levels (see CH 3OH:NOx ratios, Base Leachate Phase, Appendix B) may have stimulated ammonia removal through cellular synthesis of heterotrophic aerobes, even though Anoxic #1 dissolved oxygen concentrations were less than 0.5 mg/L (Argaman, 1982). Each increase in influent ammonia concentrations could have, as a result, enhanced aerobic heterotrophic growth, and, consequently, increased ammonia removal through bacterial assimilation; hence, surges in Anoxic #1 ammonia removal rates and percentage ammonia removals (see Appendix A for calculation definitions), under fluctuating, to some extent, influent ammonia concentrations, while mamtaining fairly constant actual ammonia concentrations within the first anoxic reactors. 45 (1/N 6ui) uoj)ej)U33uoo emoujuiv |ejo_ joioeau 46 h/N Bui) uoijejjuaouoo eiuouiuiv lejoi ajeupean aseg (l/N Biu) uonejjuaouoo eiuouiuiv Jopea_ 47 Percentage ammonia removal higher than 90% in the Aerobic #1 reactors of both systems, were obtained starting with the 13th day of the phase. The percentages were virtually 100%, once the ammonia removal performances of the systems stabilized (i.e., day 43 of the phase). Ammonia removal was due to autotrophic nitrifiers assimilation and "free" ammonia stripping in addition to only about 80% nitrification in both systems. Similarly to the Anoxic #1 reactors, some heterotrophic assimilation, using endogenous carbon and methanol occasionally escaping from the first anoxic zones, may have occurred as well. Ammonia levels of the Aerobic #1 reactors of the systems were virtually 0 mg N/L starting with the 43 r d day. Therefore, Anoxic #2 and, subsequently, Aerobic #2 reactors received virtually ammonia free mixed liquors. Consequently, Anoxic #2 and Aerobic #2 ammonia removal rates and percentage ammonia removal values (see Base Leachate Phase, Appendix B) do not reflect reactors ammonia removal capabilities, during this phase. Systems percentage ammonia removals were 100% starting with the first sampling day (i.e., day 10 of the phase). 5.1.2 NOx Levels, Nitrification and Denitrification By day 64 of the study phase, both systems generated effluents with NOx concentration of less than 2 mg N/L (Figures 5.3 and 5.4). The NOx effluent concentrations were a mainly a result of cell lysis within the Aerobic #2 reactors (see negative Aerobic #2 Total Inorganic Nitrogen percentage removals, Base Leachate Phase, Appendix B), due to too high actual hydrauhc retention times. Anoxic #1 and Anoxic #2 reactors had denitrification percentages of over 90%, while Aerobic #1 reactors had nitrification percentages of about 80%, once systems performances were progressively stable and optimized (Figures 5.3 and 5.4). The NOx levels of Aerobic #1 reactors are mainly controlled by nitrification. Ideally, when the incoming ammonia into the first aerobic reactor is completely converted (assuming no bacterial assimilation and stripping occur) to NOx, the NOx concentrations within the reactor should be equal to the ammonia concentrations within the first anoxic reactor. Decreased Aerobic #1 NOx levels signify nitrification inhibition. Subsequently, NOx levels within Anoxic #1 and #2 reactors are controlled by both, 48 uoueoijunuafjo/o pue uoueouuj!N% jojoeay o o o o o o o Bui) uojjejjuaouoo X Q N JOjoeau h/N 6iu) uoijBJjuaouoo X Q N JOjocay 50 (Aep/QOO 6) 6u!peo~i |oueina|/\| (1/N Bui) uouejjuaouoo X Q N Jojoeay (-|/N Bui) uomjjauaouoo X Q N JOjoeay 52 nitrification within the first aerobic reactor and denitrification within the corresponding anoxic reactor. Increasing anoxic NOx concentrations in correlation with increasing aerobic NOx concentrations are a result of denitrification failure and nitrate and nitrite accumulation within the system. Anoxic mixed liquors free of NOx may signify either excellent anoxic denitrification and complete nitrate and nitrite removal, or complete nitrification failure and no NOx loads into the anoxic reactors. Starting on the first phase day, methanol solutions were fed into the anoxic reactors to ensure appropriate anoxic organic carbon levels and stimulate denitrification process. Several anoxic methanol loadings were explored (Figures 5.5 and 5.6). By day 30, methanol loadings to the first anoxic reactors of both systems were about 12 g COD/day while methanol loadings to the second aerobic reactors were about 5 g COD/day. On day 35, strong nitrification inhibition was experienced, as the NOx levels within Aerobic #1 reactors of both systems were considerable low, and methanol loadings were stopped for 24 hours. It was hypothesized that increased aerobic carbon levels, due to unused anoxic methanol escaping into the aerobic reactor, stimulated fast heterotrophic growth (i.e., Aerobic #1 percentages ammonia removal were about 95% for both systems while percentages nitrification were 66% and 20%, respectively) and hindered autotrophic growth (Hanaki et al., 1990a). The hypothesis was validated by fast nitrification recovery in the absence of methanol solution additions (Figures 5.5 and 5.6, Aerobic #1 NOx concentrations lines). Subsequently, methanol loadings were progressively increased until complete anoxic denitrification, in concert with stable nitrification, were observed in both systems. 5.1.3 pH Levels, Alkalinity and Methanol Loadings Reactors pH values were somewhat erratic during incipient phase stages (i.e., first 20 days) in which bacterial populations acchmatized to the new feed (i.e., landfill leachate). Subsequently, pH levels had an almost linear increase in all reactors in response anoxic denitrification establishment, due to methanol addition (Figures 5.7 and 5.8). Increasing Anoxic #1 denitrification performances resulted in alkalinity production increases, hence, anoxic pH levels increases and, consequently, aerobic pH increases, until 53 54 55 stable systems performances were achieved. At stability, the Aerobic #1 reactors had the lowest pH values (i.e., about 7.8) as nitrification processes within the reactors consumed alkalinity. The anoxic reactors had higher pHs (i.e., about 8) as denitrification process generated alkalinity. The natural alkalinity of the leachate, in addition to the alkalinity generated within each system, ensured more than enough alkalinity requirements for nitrification, and no sodium bicarbonate addition was necessary for the treatment of the base leachate (Figures 5.7 and 5.8). 5.1.4 VSS Levels VSS levels within the reactors of both systems fluctuated the first 43 days of the phase (Figures 5.9 and 5.10). Once systems performances improved and eventually stabilized, VSS concentrations were fairly constant. Figures 5.9 and 5.10 do not show any correlation between methanol loadings and VSS levels. Therefore, it was concluded that stable reactor VSS levels and their actual values were due to the estabbshment of reliable nitrification and denitrification, rather than to any pattern in methanol loadings. System #1 VSS concentrations within each reactor slightly fluctuated between about 2000 mg/L and 2200 mg/L, while System #2 concentrations fluctuated between 1700 mg/L and 2000 mg/L, during the last phase period. For both systems, VSS/TSS ratios were about 0.56. It was speculated that the ratios may have been so low as a result of metal precipitate formation. Most interesting is that, for both systems, small fluctuations in Anoxic #1 VSS concentrations coincided with fluctuations in reactor ammonia expected concentrations, starting with day 52 of the phase, hence, at stable performances of the systems. The phenomenon is more evident for System #2 (Figure 5.11) and could support the previously made hypothesis that aerobic heterotrophic growth was triggered by probable increases in reactor ammonia concentrations. In Figure 5.11, it can be observed that increases and decreases in the VSS levels occurred concurrently with increases and decreases in ammonia expected concentrations, while fairly constant Anoxic #1 actual ammonia concentrations were maintained. (~|/6iu) uoneJiuaouoQ SSA JO»oeay 58 (UN Buj) uoj;ej)uaouoo BJUOUIIUV [e\o± jojoeay (~l/6w) uoiicnuaouoQ SSA JOjoeay 5.1-5 COD Loading versus COD Removal Methanol loadings into the anoxic reactors of the systems were progressively increased, until complete denitrification of the NOx loads was observed (i.e., virtually 0 mg N/L anoxic NOx levels). The CH 3OH:NOx ratios (i.e., as mg COD/mg N) were, for the Anoxic #1 reactors, in both systems, about 27:1, while for the Anoxic #2 reactors were about 20:1, in System #1, and 16:1, in System #2. Treating a similar landfill leachate using a MLE system, Carley and Mavinic (1991) observed that at COD:NOx-N ratios higher than 20:1, percentage nitrification was up to 40% reduced, even though overall ammonia removal and denitrification were not seriously affected. In this study, during this phase, indeed percentage nitrification was only 80%, while percentage ammonia removal was 100%. However, at lower methanol loadings, complete denitrification was not achieved. The exact reasons, for these high methanol requirements during the Base Leachate Phase, could not be identified. Figure 5.12 shows that, starting on day 59 of the study phase, under constant methanol loading, System #1, Anoxic #1 COD removal rates progressively increased almost up to the corresponding value of methanol loading, while Aerobic #1 COD removal rates progressively decreased. Hence, increased methanol removal wilhin the first anoxic reactor minimized methanol bleeding into the first aerobic reactor and, consequently, inhibited aerobic heterotrophic growth. During the same period of time (i.e., from day 59 to day 64), System #2, Anoxic #1 COD removal rates (Figure 5.13) were more erratic. However, during the entire phase, increased Anoxic #1 COD removal rates corresponded to decreased Aerobic #1 COD removal rates. Furthermore, when the Anoxic #1 rate value was almost equal to reactor methanol loading value, Aerobic #1 removal rate dropped to zero, or became negative. Therefore, assuming that aerobic heterotrophs used only methanol as organic carbon source, when methanol was entirely consumed in the Anoxic #1 reactor, the Aerobic #1 heterotrophic bacterial population, acclimatized to methanol, failed using other organic carbon sources (i.e., natural BOD 5 of the influent leachate and endogenous carbon); instead of removing COD, they generated COD through cell lysis. (N Bui /aoO Biu) XQN : HO€HO (Aep/aoo 6) 6uipeo-| loueinaiflj (N Btu/aOO 6iu) XQN I HOCHO (Aep/aoo 6) Buipeo-) loueiuaw •8 ajey |BAoiuay fjrjo Jojoeey 62 5.2 Ammonia Loading Phase The objective of Ammonia Loading Phase was to explore the 4-Stage Bardenpho process capabilities of removing nitrogen from a simulated high ammonia landfill leachate (i.e., of up to 2200 mg N/L of ammonia) and to determine adequate methods of progressively acclimatizing the process to increased ammonia loadings. The phase started on November 6, 1998, the 66 th day of the study, and ended on March 6,1999, the 185th day of the study, once stable ammonia removal performance within both systems. The data collected during this phase is presented and discussed as a function of the phase day number (i.e., starting with phase day no. 1 and ending with phase day no. 120), since operational changes (i.e., increased ammonia loadings, primarily) made only during this particular period of time determined systems performances. Influent ammonia levels were artificially increased by adding ammonium chloride solutions into the first anoxic reactors of both systems. Ammonium chloride solution concentrations were adjusted to generate successional incremental increases of about 300 mg N/L in the influent leachate ammonia concentrations. A "simulated" leachate containing ammonia levels of approximately 400, 700. 1000, 1300, 1600, 1800, and 2300 mg N/L was progressively fed to the systems. The actual influent ammonia concentration values were controlled to a certain extend (i.e., plus or minus of up to about 200 mg N/L of ammonia) by inevitable surges in ammonia concentration of the natural landfill leachate (Figures 5.14 and 5.15). The fluctuations in ammonia levels of the base leachate were either the result of seasonal changes (i.e., as a function of the amount of precipitation preceding leachate collection) or of unavoidable nitrification occurring during leachate storage (i.e., whether within the storage container or the feed pile), or both. The first incremental increase of the influent ammonia concentration was preceded by chemical addition flow adjustments to maintain Anoxic #1 AHRTs of approximately 1.5 hours for System #1 and of approximately 1.7 hours for System #2. During the entire phase, phosphorus solutions were adjusted so that a rrrinimum of 2 mg P/L of dissolved ortho-phosphate was ensured within all reactors (Manoharan et al., 1992). Aerobic dissolved oxygen levels were kept always above 2 mg/L (Mavinic, 1997; EPA, 1993). 63 The first attempt at incrementally increasing the mcoming ammonia concentrations (i.e., between November 6, 1998 and December 2,1998) was conducted as per the results of Shiskowski's research (1995). Shiskowski successfully treated incrementally-increased, influent ammonia concentration of up to 1200 mg N/L by increasing, concomitantly with ammonia loading increases, anoxic methanol loadings to match expected nitrite and nitrate production. Therefore, the first four incremental increases in the influent ammonia concentrations (i.e., influent ammonia concentration from of about 400 mg N/L up to of about 1300 mg N/L) were followed, in a day or two, by increases in the methanol loadings to the anoxic reactors. The systems were allowed to acclimatize to the each, newly-increased ammonia loading for about seven days. Considering the findings of the Base Leachate Phase and assuming that most of the ammonia entering the first aerobic reactors would be converted to NOx, CH 3OH:NOx ratios were kept at about 20:1, so that organic carbon denitrification requirements would be met. On the 27 th day of the phase, strong nitrification inhibition (i.e., decreased aerobic NOx production, as shown in Figures 5.16 and 5.17) and ammonia accumulation ( Figures 5.14 and 5.15) were observed within both systems. Fearing that the systems would not recover under the influent ammonia concentration of about 1300 mg N/L (Shiskowski, 1995), the ammonium chloride addition was stopped and the methanol loadings were discontinued for a couple of days, and, than, adjusted to meet base leachate denitrification requirements. Both nitrification and denitrification processes, within both systems, recovered amazingly fast (i.e., 4 days) and on December 7, 1998, (i.e., phase day no. 32) the second attempt to increase the influent ammonia concentrations started. After each incremental influent ammonia concentration increase, the systems were, firstly, allowed to adjust to the increased ammonia loading for a longer period of time than during the first attempt (i.e., about 10 to 14 days), and, secondly, only afterwards, anoxic methanol loadings were increased to meet denitrification requirements (see Ammonia Loading Phase, Appendix B). Nitrification inhibition and systems failure during the first attempt at increasing the influent ammonia concentration was attributed mainly to methanol breakthrough (i.e., higher Anoxic #1 COD and BOD5 levels than the corresponding Aerobic #1 levels, see Ammonia Loading Phase, Appendix B) into first 64 aerobic reactors (Carley, 1988; Hanaki, 1990a; Azevedo, 1993). Furthermore, the data collected during this phase demonstrated that the performance of the two systems was influenced not only by the timing of methanol loadings adjustments, but also by several other interconnected factors that may have ensured, or not, process stability under constant, or changing, operational parameters. The following results discussion is only a humble attempt at explaining the intricate bacterial activity that occurred within this particular type of systems (i.e., nitrification with pre- and post-denitrification processes) when treating progressively increased influent ammonia leachate concentrations. 5.2.1 Ammonia and NOx Levels Ammonia and NOx data are presented together, since a reciprocal correlation between their levels within the systems was observed during the entire phase. The terms "ammonia" and "total ammonia" refer to the sum of ammonium ion (NH/) and molecular ammonia, or "free" ammonia (NH3). The two constituents coexist in an equilibrium regulated by pH and temperature. During the Ammonia Loading Phase, the ambient temperature was maintained at 20 °C and, therefore, the only parameter that influenced their equilibrium was the reactors pHs. Figures 5.14 and 5.15 show that both systems responded very well to the first couple of incremental increases (i.e., of about 300 mg N/L each, on day no. 1 and no. 7, respectively, of the phase) in the influent ammonia concentrations and most of the ammonia was removed in the Aerobic #1 reactors (i.e., Aerobic #1 ammonia levels virtually 0 mg N/L while Anoxic #1 levels increased from about 20 to 50 mg N/L, due to the influent concentration increases). The corresponding increased Aerobic #1 NOx levels from Figures 5.16 and 5.17 demonstrate that the ammonia was mainly removed by nitrification (i.e., Aerobic #1 NOx levels of about 50 mg N/L concomitantly with Anoxic #1 ammonia levels of about the same value). In addition, Anoxic #1 and #2 NOx levels were practically zero, as the incoming loadings of nitrite and nitrate were efficiently denitrified. With the exception of nitrification inhibition and systems failure period, the above described correlation between the two constituents was observed during the entire 65 (~I/N Bui) uoiiejiuaouoo eiuouiuiv IBjoi ejeipeei p3)e|nui;s o o ID CM O O o CM O O m o o o o o in o CO 66 67 68 69 phase. During the failure period, ammonia progressively accumulated within the systems, reaching the highest levels in the Anoxic #1 reactors where, in addition to the influent concentrations, ammonia was returned by both, internal and external recycle lines (Figures 5.14 and 5.15). System #1, with Aerobic #1 ammonia levels over 1500 mg N/L and Anoxic #1 NOx levels practically zero, was more affected by nitrification inhibition than System #2, with Anoxic #1 ammonia levels of about 350 mg N/L and a considerable decreased, but still existent, Aerobic #1 nitrification performance (Figures 5.16 and 5.17). The higher Anoxic #1 and Aerobic #1 AHRTs of System #2 (compared to System #1) may have been the reason for System #2 more stable performance, when both systems were under virtually identical loading changes. Once, both, ammonium chloride and methanol additions were stopped, a peak in NOx levels of both systems occurred, as the existent ammonia within the systems was converted to nitrites and nitrates by a rapidly recovering population of nitrifiers. The NOx concentrations were fairly equal among all the reactors, as methanol loadings were adjusted to meet only base leachate denitrification requirements; hence, the organic carbon levels within the anoxic reactors were insufficient to meet denitrification requirements of the increased NOx concentrations (Figures 5.16 and 5.17). Almost all of the mcoming ammonia was removed within the Aerobic #1 reactors of the systems, when sufficient acclimatization time was ensured between successional influent ammonia concentration increases. Ammonia escaping treatment in the Aerobic #1 reactors was removed by bacterial assimilation within Anoxic #2 reactors and occasional nitrification within Aerobic #2 reactors (i.e., Aerobic #2 ammonia removal of up to about 40 mg N/L for System #1, and of up to about 60 mg /L for System #2). The effluent ammonia concentrations of both systems were virtually ammonia free, from the begirining till the end of the second attempt at incrementally increasing the influent ammonia concentrations (Figures 5.14 and 5.15). As methanol loadings were increased in a more conservative manner (i.e., the anoxic reactors were fed methanol at CH3OHNOX ratios estimated to match denitrification requirements of each antecedent influent ammonia concentration and not the actual one), elevated NOx levels could be observed within all reactors on several occasions (Figures 5.16 and 5.17). However, subsequenuy to each methanol 70 loading increase, the NOx levels within both systems decreased considerably, as a result of fairly stable denitrification. 5.2.2 Alkalinity and Methanol Loadings and Their Effects on Process Performance According to EPA (1993), the nitrification process destroys alkalinity at a theoretical ratio of 7.1 g CaC03 per g N H / - N nitrified, while denitrification process produces alkalinity at a theoretical ratio of 3.57 g CaC03 per g N0 2"-N reduced. Therefore in a pre-denitrification process, theoretically, half of the nitrification alkalinity requirements are provided by the denitrification process. Besides the alkalinity produced by denitrification, nitrification alkalinity requirements of the investigated systems, were ensured by the natural alkalinity of the landfill leachate (i.e., of about 1500 mg CaCOs/L), for up to 400 mg N/L influent ammonia concentration; in addition, sodium bicarbonate was fed to the first aerobic reactor for higher influent ammonia concentrations. The sodium bicarbonate was automatically fed to the aerobic reactor by pH/pump controllers when the aerobic pH levels dropped below 7.5 (i.e., pH value most recommended for appropriate nitrification performance; E P A 1993; Metcalf &Eddy, 1991; etc.). Alkalinity loadings into the systems progressively increased in response to higher Aerobic #1 nitrification rates (Figures 5.18 and 5.19), hence, in response to higher influent ammonia concentrations; at the same time, unit alkalinity consumption per unit N H / - N nitrified ratio remained fairly constant (i.e., mean ratio of about 5:1), even though a little higher than expected (Figures 5.20 and 5.21). Based on the theoretical ratios (EPA 1993), it was expected that denitrification process (i.e., of the Anoxic #1 reactor) would provide half (i.e., about 3.57 CaC03/g NH4+-N) of the nitrification requirements (i.e., of 7.1 g CaC03/g N H / - N nitrified), while, in addition to the natural alkalinity of the leachate, sodium bicarbonate would have to ensure only the remaining of about 3.53 g CaC03/g N H / - N requirements (i.e., instead of 5 g CaC03/g NH/-N). The higher than expected rate may have been the result of incomplete denitrification of the NOx loadings into the Anoxic #1 reactors, hence, decreased denitrification alkalinity production and increased sodium bicarbonate addition into the Aerobic #1 reactors. Poor denitrification performance was either due to insufficient anoxic organic carbon levels (i.e., as a result of methanol loading increases 71 72 (Aep/soQBO 6) 6uipeo-| Auuipnuv o o o o o o o o o o « - o > e o r ~ c o m ' < * n c \ i » - o • * t o CO CM (Aep/N Bui) ajey uoneouuu.N oiqojav 73 (N 6/COOBO 6) peiiujiN ejuouiuiv leioj.: A J I U I I B M I V ("l/N 6ui) uouejujaouoo X Q N Jopeey 74 75 kept one step behind ammonia loading increases, in order to prevent methanol bleeding from the Anoxic #1 into the Aerobic #1 reactor, during incipient stages of the second attempt at increasing influent ammonia concentrations), or due to high pH limited denitrification rates combined with denitrifiers inhibition to elevated "free" ammonia levels (i.e., under high influent ammonia concentrations, at the end of the phase). Figures 5.18 and 5.19 show that each increase in Aerobic #1 nitrification rate, hence, each increase in influent ammonia concentration, had a corresponding increase in alkalinity loadings. The rise in alkalinity loadings slightly decreased within a couple of days, once methanol loadings were adjusted to match, to a certain extent, the increased NOx loadings; hence, denitrification alkalinity production increased. Increased alkalinity loadings meant increased sodium bicarbonate solution flow into the Aerobic #1 reactors, hence, decreased reactors AHRTs; the higher the ammonia loading into the aerobic reactor, the higher the alkalinity loading, and the lower the actual hydrauUc retention time (i.e., assuming effective nitrification performance). If nitrification alkalinity requirements had not been partially provided by the denitrification process occurring within the antecedent anoxic reactor, Aerobic #1 actual hydraulic retention times might have eventually become insufficient to ensure adequate nitrifier growth (Hanaki et al., 1990a). Nevertheless, during the Ammonia Loading Phase, surges in alkalinity loadings, hence, fluctuations in Aerobic #1 reactors AHRTs, continuously threatened the nitrification process performance, and may have contributed, along with other factors (i.e., anoxic methanol loading increases made concomitantly with influent ammonia increases, to match expected aerobic NOx production, before the establishment of consistent NOx production, and without considering potential nitrite accumulation before bacterial acclimatization; Turk and Mavinic, 1989), to nitrification inhibition and systems failure, during the first attempt at increasing influent ammonia concentrations. Methanol solution additions into the anoxic reactors of both systems were used to ensure denitrification organic carbon requirements, as the landfill leachate used in this study was characterized by low levels of biodegradable organics. The COD to NOx requirement ratios, when using methanol as a supplementary 76 (Aep/N Bui) ajea uoiieouuiiuaa oiqojav 79 ( N 6/CJOO 6) paAouiay X Q N : H O C H O (1/N Bui) uoijejjuaouoQ X Q N JojoBay 80 source of organic carbon, vary from the theoretical ratios of 3.7 g COD/g NO3-N and 2.3 g COD/g N 0 2 -N recommended by EPA (1993), to the ratios established by various denitrification studies (e.g., Azevedo, 1993, found a 5:1 ratio, while Carley, 1988, found a 6.5:1 ratio). Throughout this study phase, anoxic methanol loadings were progressively increased to match aerobic NOx production and meet anoxic NOx removal requirements (Figures 5.22 and 5.23). Methanol loading increases made almost concomitantly with influent ammonia loadings increases (at the CH 3OH:NOx ratios of about 20:1, established during the Base Leachate Phase) resulted in methanol bleeding into the Aerobic #1 reactors, and subsequent nitrification inhibition, a probable result of enhanced heterotrophic bacterial growth (i.e., see higher Anoxic #1 COD levels in comparison with corresponding Aerobic #1 COD levels, during nitrification inhibition period, Ammonia Loading Phase, Appendix B) (Hanaki et al., 1990a). Both systems responded well to methanol loading increases made after allowing sufficient time for the establishment of reasonably stable nitrification processes within the Aerobic #1 reactors (i.e., consistent NOx loadings into the anoxic reactors). The methanol loadings into the Anoxic #1 and #2 reactors of System #1 were, for most of the time, equal, as both reactors received estimated equal NOx loads (i.e., the internal recycle line, that returned NOx into Anoxic #1 reactor, had a flow rate of about 40 L/day, theoretically equal with the NOx-laden overflow from the first aerobic reactor into Anoxic #2 reactor). The methanol loadings into the Anoxic #2 reactor of System #2 were higher than the loadings into the Anoxic #1 reactor of the system, since the second anoxic reactor received higher NOx loads (i.e., the internal recycle flow rate was about 30 L/day, while the first aerobic overflow into the second anoxic reactor was about 40 L/day) (Figures 5.22 and 5.23). In addition, the methanol loadings increases were made under more conservative CH3OH:NOx ratios (i.e., lower ratios in accordance with previous studies: Azevedo, 1993, and Carley, 1988). Figures 5.24 and 5.25 show that, by the end of the phase, the Anoxic #1 COD:NOx ratios of both systems were about 5:1, while the Anoxic #2 ratio was about 5:1 for System #1 and about 10:1 for System #2. The higher Anoxic #2 ratio of System #2 was a direct result of decreased denitrification rates within the reactor (Figure 5.25). 81 82 83 san|BA Hd ( D C S | 0 0 t J ; O ( D ( N | 0 0 ' * O ("l/N 6ui) uoi)ej)U93uoo eiuouiuiv ..aa-y.. oixouv 84 85 Progressively increased denitrification performances, within both systems (hence, increased denitrification alkalinity productions) resulted in higher anoxic pH levels. Figures 5.26 to 5.29 show that anoxic pH levels increases, from pHs of about 8 up to pHs of as high as 9.8 (i.e., Anoxic #2 reactor of System #1), determined increased "free" ammonia levels within all anoxic reactors (Benefield et al., 1982; EPA, 1993). In addition to the reactor pH levels, "free" ammonia concentrations within the reactors were also controlled by the reactor total ammonia concentrations at a given time. Even though Anoxic # 1 pH levels were always lower than the Anoxic #2 pH levels (i.e., as a result of dilution by the recycle lines flows), the Anoxic #1 "free" ammonia concentrations (i.e., of about 40 mg N/L, after the last incremental influent ammonia increase) were always higher than the Anoxic #2 "free" ammonia concentrations (i.e. of about 3 mg N/L for System #1 and of about 7 mg N/L for System #2), since the total ammonia concentrations within the first aerobic reactor were always considerably higher than the total ammonia levels within the second anoxic reactor of both systems (Figures 5.28 and 5.29). However, the higher Anoxic #2 pH levels (i.e., pHs of over 9.6 within the Anoxic #2 reactors of the systems compared to the maximum recommended pH = 8 for optimal denitrification; EPA, 1993; Metcalf &Eddy, 1991) seemed to have a more harmful influence on denitrification performance within the respective reactors than the elevated Anoxic #1 "free" ammonia levels on denitrification process within the corresponding reactors (i.e., percentage denitrification of 93% and 76%, respectively, for System #1 and of 52% and 43% respectively, for System #2, at the end of the phase). Furthermore, under virtually equal Anoxic #1 total ammonia levels within both systems (i.e., of about 300 mg N/L, at the end of the phase), hence, equal" free" ammonia levels (i.e., of about 40 mg N/L, at the end of the phase), the higher Anoxic #1 AHRT of System #2 (compared to System #1; 1.7 hours in comparison with 1.5 hours, respectively) appeared to be detrimental, rather than beneficial, to the overall denitrification process within System #2. The higher Anoxic #1 AHRT of the System #2 meant a longer contact time within the reactor, hence, the longer the denitrifiers were exposed to elevated "free" ammonia levels, the stronger they were inhibited. 86 Denitrification inhibition, within the Anoxic #1 reactor of the System #2, subsequently determined NOx accumulation within the system (see Figure 5.17 in comparison with Figure 5.16, from day no. 100 to day no. 120), hence, progressively higher NOx loadings into the anoxic reactors, and, eventually, decreased percentage denitrification. In addition, hypothesizing that higher anoxic pH levels, due to increased denitrification performance and consequently increased alkalinity production, would have eventually meant enhanced denitrification inhibition, the increase in methanol loadings ceased. The continuous recycle (i.e., by the internal recycle lines) of the nitrifiers through the elevated "free" ammonia levels of the Anoxic #1 reactors of the systems probably resulted in nitrate foiming organisms inhibition (i.e., Nitrobacter, nitrite oxidizers); consequently, progressive nitrite accumulation, within both systems, occurred after each incremental increase in the influent ammonia concentration (Figures 5.30 and 5.31). However, nitrifiers acclimatized, to a certain extent, to each successional increase in the Anoxic #1 "free" ammonia levels and, by the end of the phase, were very well able to handle the Anoxic #1 40 mg N/L of "free" ammonia (Turk and Mavinic, 1989). Nitrification processes within the Aerobic #1 reactors of both systems were not affected by the nitrate-formers inhibition to "free" ammonia (i.e., consistent over 100% Aerobic #1 nitrification percentages for both systems). The NO2 /NOX ratios eventually stabilized to a value of approximately 0.8 for the Anoxic #1 and #2, and Aerobic #1 reactors of the systems, and to a value of approximately 0.6 for the Aerobic #2 reactors of the systems (Figures 5.32 and 5.33). By the end of the phase (i.e., day no. 120) System #1 had Aerobic #1 average nitrite levels of about 300 mg N/L, with corresponding NOx levels of400 mg N/L; System #2 had Aerobic #1 average nitrite levels of about 400 mg N/L, with corresponding NOx levels of 600 mg N/L. With the exception of the nitrification inhibition period, when increases in methanol loadings were made without considering the possible nitrite accumulation within the systems (i.e., the methanol requirements for nitrite complete reduction are 2.3 g COD/g N, while the requirements for nitrate complete reduction are 3.7 g COD/N; hence, the CH 3OH:NOx ratio should have been adjusted to meet nitrite reduction requirements rather than nitrate reduction requirements), the overall performance of both systems did not seem to be affected by nitrite 87 88 o o CO (l/N 6ui) uoijBijuaouoo aiuijN JOjoeay 89 v - x * O .2 Z l § < Z -Q ^ O CM 0) O < z CM o 1 c < X O Z CM o CM O o i— CD < X O Z CM" O "8 « 3 O E <o CO CO _ J o i l X o tn < co O !~ 2 To < CD CM .2 i l x ^ O CO C <D < CD CM O o I CO CM 8 1 00 CM | 8 d d d d II it 1 it i i CM CM 1 a: \ CM 90 =*  O — ? o ™ c O < z x o CN o CN X * O .2 2 i § < 2 CN o O :— CD < X O CN O 2 "8 is Z> o £ B •= CD W _ l « LI-'S CD CO o O I— < O X o c < y LL 00 CD CO CN % O o I— CD < (~|/|S| 6ui) uoueJiudouoQ eiuouiiuv |BJOI ajeipean pajeiniuis oney XQN/20N JOjoeey 91 accumulation. On the contrary, high N0 2/NOx ratios resulted in decreased CH 3OH:NOx ratios (i.e., 5:1); this resulted in decreased methanol additions into the anoxic reactors of both systems. 5.2.3 VSS Levels Reactor VSS concentrations were controlled by several variables including influent ammonia concentration (Figures 5.34 and 5.35), methanol loading (Figures 5.36 and 5.37), actual hydraulic retention time, internal and external recycle flows, and performance of reactor intended process. Figures 5.34 and 5.35 show that, as expected, reactor VSS concentration began to steadily increase in response to the first four influent ammonia concentration increases. During nitrification inhibition period, VSS levels became more erratic, with peaks in the Aerobic #2 VSS levels of about 5000 mg /L for System # 1 and of 5100 mg /L for System #2. The high reactor VSS concentration was a result of increased ammonia concentrations reaching the Aerobic #2 reactors, due to nitrification inhibition within the Aerobic #1 reactors. Throughout the recovery period, VSS levels within the systems never did return to the initial levels (i.e., of about 2500 mg/L within System #1 and of about 2000 mg/L within System #2) and were, for the most part, above 3000 mg/L. During the second attempt at incrementally increasing influent ammonia concentrations, System #1 VSS concentrations increased up to values of about 4800 mg/L and 3600 mg/L, respectively, in the anoxic reactors, and about 4000 mg/L and 5200 mg/L, respectively, in the aerobic reactors. System #2 VSS levels increased to approximately 4000 mg/L and 3800 mg/L, respectively, in the anoxic zones, and to 3200 mg/L and 3600 mg/L, respectively, in the aerobic zones. Increased reactor denitrification performance, hence, enhanced cell synthesis, within the Anoxic #1 reactors might have been the reason for Anoxic #1 VSS levels higher than Anoxic #2 VSS levels. High Aerobic #2 VSS concentrations were probably the result of the reactor long hydraulic retention time (i.e., AHRT of 6 hours). Differences of up to 1600 mg/L of VSS between the VSS concentrations of the two systems was common, possibly due to slightly different operating conditions. 92 (~I/N Bui) uouejuiaouoo 94 (Aep/rjoo 6) 6u!peon |oueina|/v (n/Boi) UOI;BJJU90UOO SSA OJXOUV 96 Figure 5.36 shows that, starring with day no. 104 of the last increase in the influent ammonia concentration, under identical methanol loadings, System #1, Anoxic #1 VSS concentrations increased and remained fairly constant at a value of about 4800 mg/L, while the Anoxic #2 VSS concentration sharply decreased from a value of about 4400 mg/L to a value of about 3600 mg/L. The decrease in Anoxic #2 VSS levels was affected by increased reactor pH values and sudden denitrification inhibition. From Figure 5.37, it can be observed that the difference between Anoxic #1 and Anoxic #2, of System #2, VSS concentrations was considerably smaller (i.e., Anoxic #1 VSS of about 4000 mg/L while Anoxic #2 VSS of about 3800 mg IV). Furthermore, the two systems had almost identical Anoxic #2 VSS levels and considerably different Anoxic #1 VSS levels. The almost identical Anoxic #2 VSS concentrations may have been the result of similar denitrifying bacteria inhibition to the increased reactor pH levels (i.e., as both reactors were under virtual identical ammonia and methanol loadings); however, denitrification inhibition within the anoxic reactor of System #1 was experienced suddenly, while denitrification inhibition within the anoxic reactor of System #2 was experienced consistently throughout the phase. The VSS/TSS ratios were, most of the time, approximately 0.8 in System #1 reactors and slightly lower (i.e., about 0.76) in System #2 reactors. Occasional lower ratios occurred mainly during incipient stages of the phase. The effluent VSS concentration fluctuated between about 10 mg/L and 50 mg/L, during system stable performances, with peaks of as high as 328 mg/L, during nitrification inhibition (see Ammonia Loading Phase, Appendix B). 5.2.4 System Performance and Sequenced Reactor Performance Figures 5.38 and 5.39 show that, by day 120, both systems generated ammonia free effluents, with NOx concentrations of 78 mg N/L, and 250 mg N/L, respectively, when treating a simulated leachate with an ammonia concentration of over 2200 mg N/L. Figures 5.40 and 5.41 show that, except for the nitrification inhibition period, system percentage ammonia removals were consistently about 100%, while percentage total inorganic nitrogen removals were erratic, mainly as a result of surges in NOx removal rates. However, on the last day of the phase (i.e., day no. 120), the systems had percentage total inorganic 97 98 99 100 101 nitrogen (TIN) removals of 95%, and 82%, respectively. The higher %TIN removal of System #1 than of System #2 was the result of better overall denitrification performance of System #1 than of System #2 (i.e., the lower Anoxic #1 AHRT of System #lmeant shorter contact time and decreased denitrification inhibition to elevated "free" ammonia, hence, better overall denitrification performance). The performance of each sequenced reactor and the extent of the influence that one reactor performance had over other reactors performances determined the overall ammonia removal capabilities of the systems. (a) Anoxic #1 Reactor Influent ammonia concentrations were diluted in the Anoxic #1 reactor mainly by the internal and external recycle lines. The average maximum ammonia concentration within the reactor was about 300 mg N/L, except during the denitrification inhibition period (Figures 5.42 and 5.43). The highest ammonia concentration of the systems was always, as expected, Anoxic #1 ammonia concentration. About 10% to 20% of the incoming ammonia was removed in the Anoxic #1 reactor, mainly by bacterial assimilation. The percentage ammonia removals were more consistent at lower influent ammonia concentrations and became more erratic at higher influent ammonia concentrations. However, the intended and actual nitrogen removal method of the reactor was the denitrification process. When NOx loadings into the reactor (i.e., NOx was returned from the subsequent aerobic reactor, into the anoxic reactor by the internal recycle line) were completely converted to nitrogen gaseous compounds, reactor NOx levels were virtually zero and percentage denitrification was about 100% (Figures 5.42 and 5.43). Fluctuations in reactor NOx levels and percentage denitrification were a result of influent ammonia increases and subsequent methanol loadings adjustments. Decreased denitrification performance within the reactor resulted in methanol bleeding into the subsequent aerobic reactor and, consequently, some nitrification inhibition, as well as, in decreased alkalinity production. Decreased alkalinity production effected aerobic sodium bicarbonate addition requirements and flows, hence, the actual hydraulic retention time of the subsequent aerobic reactor. On the last day of this phase, the Anoxic #1 percentage denitrification was 93% for System #1, and 52% for System #2. uom;3uu)!uaa°/a pue iBAouiay eiuouiuiv% 103 uoijeoiju^uaQo/o pue iBAOiuay B!UOUIUIV% 104 105 106 (b) Aerobic #1 Reactor Ideally, Aerobic #1 reactor ammonia levels should have been zero while reactor NOx levels should have been identical to Anoxic #1 ammonia levels. Ammonia was removed in the Aerobic #1 reactor mainly by nitrification and some bacterial assimilation, rather than by any air stripping, as the reactor "free" ammonia levels were virtually zero. Percentage ammonia removals reached values of over 90%, even though some of the incoming ammonia escaped treatment (Figures 5.44 and 5.45). Percentage nitrification values were frequently over 100%, as NOx accumulation periodically occurred in the reactor, mainly as a result of poor denitrification performance in the Anoxic #1 reactor (Anoxic #1 NOx returned to the Aerobic #1 reactor, hence, higher NOx levels than expected; see Appendix A for %Nitrification calculation definition). For the most part, during nitrification inhibition, decreased reactor nitrification performance resulted in decreased anoxic NOx loadings, hence, decreased denitrification performance in, both, antecedent and subsequent anoxic reactors. (c) Anoxic #2 Reactor Some of the ammonia escaping treatment in the Aerobic #1 reactor was removed by bacterial assimilation in the Anoxic #2 reactor. Reactor percentage ammonia removal was erratic, fluctuating between 0% and about 90% (Figures 5.46 and 5.47); it was not representative for reactor ammonia removal capabilities, and was mainly controlled by the mcoming ammonia concentration. However, the principal role of the reactor was to denitrify the incoming NOx. Reactor NOx levels were regulated by 3 influential variables: denitrification process performance within the reactor itself, and nitrification and denitrification performances within the antecedent aerobic and anoxic reactors. Decreased Anoxic #1 denitrification performance combined with fairly good Aerobic #1 nitrification performance determined increased NOx loadings into the Anoxic #2 reactor. Anoxic #2 denitrification performance, limited by increased pH levels, incapacitated the reactor to handle the consistently increasing incoming NOx concentrations; hence, the high NOx reactor concentrations, and subsequently decreased reactor percentage denitrification (i.e., of 76% for System #1 and of 43% for System #2, on the last day of the phase). Furthermore, decreased denitrification performance within the Anoxic #2 reactor resulted in further NOx accumulation within the system, due to the eventual return of the NOx, by the external recycle line. uo!)eoijM)!uafj% pue leAOiuay e!UOUiuiv% uo!jBoijuj!uaa% pue |BAoaiey ejuouiuiv% 109 .110 uo!jeouu»!N% pue leAotuay eiuouiuivo/0 I l l (d) Aerobic #2 Reactor Aerobic #2 ammonia and NOx levels were mainly controlled by the performances of all antecedent reactors (Figures 5.48 and 5.49). The system remaining ammonia concentration was nitrified within the Aerobic #2 reactor; however, the surges in reactor percentage nitrification were generated by increases in reactor NOx loadings, rather than by enhanced nitrifying bacteria activity within the reactor. Furthermore, as reactor ammonia loadings always fluctuated, a stable and reliable population of nitrifiers never developed. Therefore, the reactor was not able to handle increased ammonia loads such as those during the nitrification inhibition period. In addition, the reactor was not provided with an additional alkalinity source, hence, alkalinity availability within the reactor was limited. Reactor percentage ammonia removal and percentage nitrification did not really reflect reactor performance. An Aerobic #2 reactor with an AHRT of 1.5 hours (versus the employed Aerobic #2 AHRT of 6 hours) would have been more than adequate to occasionally nitrify the remaining ammonia and to aerate the mixed liquor prior to clarification. 5.3 pH Phase Speculating that decreased anoxic pH values would improve the overall denitrification performances of the systems, decreases in the pH set point of both pH/pump controllers started on March 6, 1999. The pH set point of the controller was adjusted so that the initial Aerobic #1 reactor pH value of 7.6, plus or minus 0.1 pH units, was decreased to a pH value of 7.3, plus or minus 0.1 pH units, over approximate three days (i.e., about 0.1 pH units step-decrease per 24 hours). The decrease in the Aerobic #1 pH value generated a decrease of about 0.2 pH units in the Anoxic #1 pH value and of about 0.4 pH units in the Anoxic #2 pH value, within both systems (i.e. Anoxic #1 pH values down to 8.4 from 8.6, and Anoxic #2 pH values down to 9.2 from 9.6). The initial pH step-decrease in the pH values of the reactors "unbalanced", both, nitrification and denitrification performances within the systems (i.e., sudden nitrification and denitrification inhibition, followed by fast recovery). However, the subsequent pH decreases did not appear to affect the processes 112 to the same extent (i.e., fairly stable nitrification and denitrification process), and the overall performance of both systems started to respond positively to the changed pH. Over a period of time of seven days, the effluent NOx concentrations of the systems decreased from 78 mg N/L to 60 mg N/L, and from 250 mg N/L to 35 mg N/L, respectively, while effluent ammonia concentrations remained practically zero (pH Phase, Appendix B). On the 12th day (i.e., March 17, 1999) of the phase, increased effluent NOx concentrations were observed (i.e., of about 106 mg N/L, and 102 mg N/L respectively), and the speculation was that the denitrifying bacterial population was still acclimatizing, and that, given sufficient time, the effluent NOx concentration would return to the previous values. This reasoning was also based on the fact that both systems were still generating ammonia free effluents and the alkalinity additions into the systems did not decrease (i.e., alkalinity loadings to the systems still of over 90 g CaCOs/day). However, even though the effluent NOx concentrations did not "dramatically" further increase (i.e., only up to about 130 mg N/L and 260 mg N/L, respectively, on the 31 s t day of the phase), on April 5, 1999, residual ammonia concentrations started to appear in the effluents of both systems (i.e., effluent ammonia concentrations of about 70 mg N/L). Furthermore, the decreased Aerobic #1 wasting from 0.500 L/day to 0.250 L/day (i.e., SRT increased from 20 to 40 days), on the 28 th day of the phase, and the progressively decreased methanol loadings, starting with 35 th day of the phase, did not seem to improve the performance of the systems and to facilitate process recovery. By April 12, 1999 (i.e., day no. 38 of the phase), massive ammonia accumulation within both systems was observed and, therefore, on April 13, 1999, the influent ammonia concentrations were decreased from about 2250 mg N/L to about 1250 mg N/L. Considering that the performance of the systems stared to improve immediately after the pH changes, the determinant factor suspected for the failure of the systems was the landfill leachate, somewhat changed, composition (i.e., higher percentage calculated toxicity for March than the average percentage of the preceding months; City of Vancouver, 1999). The system bacterial populations, already under changing operating parameters (i.e., pH changes, mainly), were incapable of simultaneously acclimatizing to an 113 additional severe change. Therefore, in addition to decreased influent ammonia concentrations, on April 13, 1999, the systems were fed from a newly collected batch of leachate. The systems recovered, to a certain extent, over a period of 5 days, and, by day no. 44 of the phase, both systems were fed ammonia concentrations of about 2200 mg N/L, methanol loadings were increased to meet denitrification estimated requirements, and aerobic wasting was reinstated at 0.500 L/day (i.e., SRT = 20 days). Both systems started to generate virtually ammonia free effluents on the 47 th day of this phase. On the last day of the phase (i.e., May 19,1999 and day no. 7) the effluent NOx concentration of each system was about 170 mg N/L; 5.3.1 Ammonia Levels and Nitrification Rate Even though the systems generated effluents containing residual ammonia starting at day no. 31, Figures 5.50 and 5.51 show that ammonia accumulation within the systems began after day no. 19 of the phase. In addition, Figures 5.52 and 5.53 show that, except for the initial acclimatization period being characterized by nitrification rate surges, the first significant decrease in the nitrification rates of the systems also occurred on the 19th day of the phase. Residual ammonia was not observed in the effluents of the systems till the 31 s t day of the study, since ammonia escaping treatment in the Aerobic #1 reactor was nitrified in the Aerobic #2 reactor (i.e., increased Aerobic #2 nitrification rates after day no. 19 of the phase). Once Aerobic #1 nitrification inhibition suddenly increased (i.e., in Figures 5.52 and 5.53, from day no. 28 to day no. 30, the nitrification rate line is sharply descendent) and the limited alkalinity in Aerobic #2 was exhausted (i.e., Figures 5.52 and 5.53 show that Aerobic #2 pH value decreased from about 8.8 to about 6.5), residual ammonia started to appear in the effluents of both systems. Most notable is not only the rapid recovery of both systems but also the fact that, once stable, the systems were able to handle surges in the influent ammonia concentrations of about over 300 mg N/L, without any evident problem. Figures 5.50 and 5.51 show that, at the end of the phase, simulated leachate ammonia concentrations suddenly increased and decreased while the ammonia concentration within the Anoxic #1 reactors remained fairly constant and practically zero within all other subsequent reactors. The (1/N Bui) uoj}ej)uaouoo eiuouiuiv lejoi jopeay 115 116 117 118 corresponding aerobic nitrification rates (Figures 5.52 and 5.53) also increased somewhat; however, the phenomenon within the Anoxic #1 reactor could not be anything else but the result of enhanced bacterial growth of an aerobic secondary heterotrophic population, triggered by increased incoming ammonia levels (see Base Leachate Phase hypothesis). The hypothesis is confirmed, again, by the corresponding increases in the Anoxic #1 VSS levels and relative constant specific denitrification rates from Figures 5.60 and 5.61 (shown later in this chapter). 5.3.2 NOx Levels and Denitrification Rate Ideally, the pH Phase should have ended with the systems generating ammonia and NOx free effluents. Unfortunately, complete anoxic denitrification of the generated NOx concentrations was never achieved and, by the end of the phase, the percentage denitrification of all the anoxic reactors of the systems was only about 50% (Table 5.1 and 5.2) Figures 5.54 and 5.55 show that, even though the systems were under consistently constant operating parameters (i.e., relative constant ammonia loadings and nitrification, constant methanol loadings and fairly constant reactor pHs within 4 days from the first pH step-decrease), the NOx levels of all reactors started to progressively increase beginning on the 10th day of the phase. The phenomenon is more evident for System #2 (Figure 5.55) than for System #1 (Figure 5.54), since System #2 may have experienced stronger inhibition of the heterotrophic denitrifying bacterial population (i.e., similarly to denitrification inhibition at the end of Ammonia Loading Phase; hence, as a result of the longer contact time within the Anoxic #1 reactor of System #2 versus System #1). Comparing the total ammonia data with the corresponding NOx data (i.e., from incipient stages of the phase, Figures 5.50, 5.51, 5.54, and 5.55), the conclusion could be reached that the overall denitrification process within the systems started to fail before the nitrification process. However, anoxic denitrification rate plots-in of Figures 5.56 and 5.57 have almost identical patterns with the aerobic nitrification rates in Figures 5.52 and 5.53. Therefore, the only logical conclusion is that, both nitrification and denitrification experienced simultaneous and similar 119 120 (1/N 6UJ) uouejuuaauoQ X Q N JOjoeay (Aep/N BUJ) aiey uoijeouujiuarj j apeay (Aep/N Bui) aiey UOIJBOIIUJIUSQ jojoBay 123 inhibition, rather than the failure of one process determined the failure of the other (see Ammonia Loading Phase system failure). Denitrification rates eventually increased, and denitrification processes within the systems recovered, to a certain extent, and stabilized. However, notable for the pH Phase are the changes that occurred within the anoxic reactors at the incipient stages of the phase, when both systems responded positively to the changed pH levels of the anoxic reactors, and the performance of the processes at the end of the phase. The most relevant data for anoxic NOx levels in correlation with changes that took place within the anoxic reactors (as a result of the decreased reactor pH) are summarized in Tables 5.1 and 5.2. The tables contain data collected on the last day of the of the previous phase, and on four consecutive sampling days from the beginning and from the end of this phase. On the highlighted 7 t h day of the phase, the systems achieved high percentage removal of total inorganic nitrogen (i.e., 97% and 98%, respectively) concomitantly with relatively low effluent NOx concentrations (i.e., 59 mg N/L and 35 mg N/L, respectively). T a b l e 5 .1 : S y s t e m #1 A n o x i c R e a c t o r p H , N O x a n d % D e n i t r i f i c a t i o n Date Phase Anoxic #1 Reactor Anoxic #2 Reactor Day pH N H 3 NOx %Den. CH 3OH PH N H 3 NOx %Den. C H 3 O H No. (mgN (mgN per** (mgN (mgN per** per L) per L) NOx R e m per L) per L) NOx R e m Mar 05/99 120* 8.6 39 17 93 4 9.8 0 112 76 6 Mar 08/99 3 8.4 27 3 99 5 9.1 3 82 80 6 Mar 10/99 5 8.5 33 51 81 5 9.2 1 140 71 6 Mar 12/99 7 8.6 38 18 92 5 9.2 1 92 79 6 Mar 15/99 10 8.5 29 27 89 5 9.0 1 122 70 7 May 12/99 68 8.5 27 93 62 6 9.1 0 121 70 8 May 14/99 70 8.5 27 83 65 6 9.1 0 105 74 7 May 17/99 73 8.4 25 96 63 6 9.0 0 146 64 8 May 19/99 75 8.4 24 138 53 6 9.0 0 200 57 9 *March 05, 1999, is the 120th day of the Ammonia Loading Phase (i.e., the last day of the previous phase). ••Methanol addition per NOx converted as unit COD per unit N (i.e., mg COD/mg N of g COD/g N). 124 T a b l e 5.2: S y s t e m #2 A n o x i c R e a c t o r p H , N O x a n d % D e n i t r i f i c a t i o n Date Phase Anoxic #1 Reactor Anoxic #2 Reactor Day pH NH 3 NOx %Den. CH 3 OH pH N H 3 NOx %Den. CH 3 OH No. (mgN (mgN per** (mgN (mgN per** per L) per L) NOx R e m per L) per L) NOx R e m Mar 05/99 120* 8.5 34 180 52 6 9.6 7 328 43 10 Mar 08/99 3 8.5 32 13 94 6 9.2 5 35 92 7 Mar 10/99 5 8.6 34 38 83 6 9.2 4 102 78 7 Mar 12/99 7 8.0 14 1 99 17 8.5 17 1 99 21 Mar 15/99 10 8.5 29 49 77 7 9.0 0 93 78 8 May 12/99 68 8.5 25 163 44 8 9.0 0 220 51 12 May 14/99 70 8.5 25 122 51 8 9.1 0 129 70 10 May 17/99 73 8.4 23 126 49 8 9.0 0 164 61 11 May 19/99 75 8.4 23 143 49 7 9.0 0 203 56 11 *March 05, 1999, is the 120 day of the Ammonia Loading Phase (i.e., the last day of the previous phase). **Methanol addition per NOx converted as unit COD per unit N (i.e., mg COD/mg N of g COD/g N). A decrease in the Anoxic #1 pH of about 0.2 pH units effected an average decrease in the reactor "free" ammonia concentrations of about 15 mg N/L. The "free" ammonia concentrations of the Anoxic #2 reactor eventually became zero, as a result of reactor decreased pH (i.e., from about 9.6 to about 9.2); this was concurrent with increased Aerobic #1 nitrification performance (i.e., Aerobic #1 ammonia concentrations virtually zero). Over the initial 7 day period of the phase, percentage denitrification of the anoxic reactors (even though erratic) reached a value of over 90% on at least two occasions, except for the Anoxic #2 reactor of the System # 1, that reached only a maximum of 80%. The methanol to NOx removed ratios slightly increased (i.e., ratios of about 6:1 and 9:1, respectively, and 8:1 and 11:1, respectively) writhin all reactors, even though the adjusted methanol loadings to the Anoxic #1 reactors of both systems were somewhat decreased; the methanol loadings to the Anoxic #2 reactors were similar to the methanol loadings of the previous phase (see pH Phase versus Ammonia Loading Phase, Appendix B). 125 By the end of this phase, the overall denitrification performance of System #2 improved in comparison with the system performance of the previous phase. The NOx levels within the anoxic reactors of the system decreased from 180 mg N/L to 143 mg N/L, and from 328 mg N/L to 203 mg N/L, respectively. The overall denitrification process improvement was a direct result of the increased Anoxic #2 denitrification performance, rather than to Anoxic #1 denitrification performance. The overall denitrification performance of System #1 was considerably decreased in comparison with the system performance of the previous phase. 5.3.3 Specific Denitrification Rates and VSS Levels Figures 5.58 and 5.59 show that, after an initial increase in the specific denitrification rates, the systems experienced progressive decrease in the rates, and, by the end of the phase, the specific denitrification rates stabilized at much lower values than the initial ones (i.e., values of about 0.300 mg N/day/mg VSS in comparison with Anoxic #1 values of about 0.700 mg N/day/mg VSS and Anoxic #2 values of about 0.600 mg N/day/mg VSS). Interestingly, the initial increase in anoxic VSS levels (i.e., from 4800 mg/L and 3600 mg/L to about 8000 mg/L and 7000 mg/L, respectively, and from 4000 mg/L and 3800 mg/l to about 6000 mg/L and 7000 mg/L) did not continue and the VSS levels stabilized, more or less, at their initial values; at the same time, specific denitrification rates decreased during systems failure period, and, than, increased, but never achieved the initial values (Figures 5.60 and 5.61). In addition, the VSS/TSS ratios of all reactors increased from an average value of about 0.8 to values of 0.86 and over. Therefore, even though it is difficult to conclude with absolute certainty, it is fair to speculate that the denitrifying bacterial population developed a persistent toxicity to some landfill leachate constituents. This toxicity prevented the denitrifiers from completely recovering by the end of the phase and to becoming as efficient as in the beghining of the phase (Figures 5.56, 5.57, 5.60 and 5.61). 126 Hd jojoeay ( S S A 6UI/ABP/N BUI) ajey uoneouujjuarj oijpads aojaeay 127 128 129 (I/Bui) SSA JO»oeay (SSA Biu/Aep/N 6tu) ajey uoijeoyujiuaa oijpads jojoeay 130 5.3.4 System Performance Table 5.3 summarizes the overall system performances of the first and last days of the pH Phase. The percentage total ammonia removal of both systems was consistently at 100%, while the percentage total inorganic nitrogen (TIN) removal was erratic during the incipient phase stages and stabilized at about 90% during the last phase period. System #1 effluent NOx concentration decreased to a minimum of 37 mg N/L on the 3 r d day of this phase, when system percentage total inorganic nitrogen removal reached a maximum of 98%. The minimum of 35 mg N/L of NOx effluent concentration was achieved by System #2 on the 7 t h day of the phase (i.e., with %TIN removal of 98%). During the last period of the phase, both systems generated effluents containing over 100 g N/L of NOx. T a b l e 5.3: S y s t e m N i t r o g e n R e m o v a l O v e r a l l P e r f o r m a n c e Date Phase System #1 System #2 Day Effl. %System Removal Effl. %System Removal No. Amm.* NOx Total Total Inorg. Amm.* NOx Total Total Inorg. (mgN (mgN Amm.* Nitrogen** (mgN (mgN Amm. Nitrogen** per L) per L) (%) (%) perL) per L) (%) (%) Mar 05/99 120* 1 78 100 96 0 250 100 86 Mar 08/99 3 0 37 100 98 0 41 100 98 Mar 10/99 5 2 83 100 95 0 64 100 96 Mar 12/99 7 0 59 100 97 0 35 100 98 Mar 15/99 10 3 73 100 96 0 73 100 96 May 12/99 68 0 118 100 93 0 202 100 88 May 14/99 70 0 107 100 94 0 139 100 92 May 17/99 73 0 118 100 94 0 144 100 92 May 19/99 75 0 170 100 90 0 177 100 89 •Ammonia or Total Ammonia refers at the sum of ammonium ion (NH4) and molecular ammonia (NH3). **Total Inorganic Nitrogen (TIN) refers at the sum of ammonia and NOx as N. Figures 5.62 and 5.63 show that bacterial inhibition again (i.e., similarly to the Ammonia Loading Phase system failure period) was more intense within System #1 (i.e., with Anoxic #1 AHRT of 1.5 hours) than within System #2 (i.e., with Anoxic #1 AHRT of 1.7 hours). During the system failure period, System #1 131 132 133 generated an effluent with an ammonia concentration of up to 789 mg N/L, while System #2 effluent ammonia concentration reached a maximum of 672 mg N/L. Once the system recovered, to a certain extent, and stabilized, the effluent NOx concentrations of System #2 were somewhat erratic and varied from 140 mg N/L to 200 mg N/L (Figure 5.63). System #1 effluent NOx concentrations fluctuations were relatively limited in value (i.e., between 120 mg N/L and 170 mg N/L) and occurrence (Figure 5.62). Therefore, System #1 appeared to have a more stable overall performance than System #2. Furthermore, System #2 effluent NOx concentrations were always (i.e., except for the system failure period) higher than System #1 corresponding concentrations; hence, System #1 overall performance was not only more stable but was also more effective than System #2 performance. 5.4 Temperature Phase The main objective of the Temperature Phase was to investigate the performance of the systems under decreased ambient temperatures, when treating influent ammonia concentrations of about 2200 mg N/L. On May 19, 1999 (i.e., last day of the pH Phase), after sampling, the ambient temperature was decreased from 20 °C to 17 °C. By May 21, 1999 (i.e., phase day no. 2), the effluent NOx concentrations of both systems increased from about 170 mg N/L to over 250 mg N/L, while effluent ammonia concentrations remained practically zero (Temperature Phase, Appendix B). After a period of time of 5 days with no change in effluent quality (i.e., on May 24, 1999, and phase day no. 5), the ambient temperature was further decreased to 14 °C. Taking into account the results of previous studies (i.e., Azevedo et al., 1995; Guo, 1992), starting with the subsequent day (i.e., May 25, 1999), Aerobic #1 wasting was stopped, and methanol loadings to the anoxic reactor were progressively decreased. By the 7 t h day of the phase, while the systems were still generating ammonia free effluents, and almost all of the incoming ammonia was removed by nitrification within the Aerobic #1 reactors, the NOx concentrations of the effluents increased to about 300 mg N/L, and to about 600 mg N/L, respectively. 134 was hypothesized that denitrification inhibition might be the determinant factor for the failure of this particular type of system (i.e., nitrification with pre-and post-denitrification processes), when treating high ammonia landfill leachate. Speculating that another temperature step-decrease may have a limited effect on system nitrification performance and would affect mainly denitrification performance, the ambient temperature was further decreased to 10 °C (i.e., on phase day no. 7, after sampling). Starting with the 9 t h day of the phase, effluent residual ammonia concentrations (i.e., of about 40 mg N/L, and 110 mg N/L, respectively, on May 28, 1999) were observed concomitantly with still steadily increasing effluent NOx concentrations. Assuming that only minor nitrification inhibition was occurring, and that, given time, nitrification process would recover, the existing operating parameters and conditions were maintained (i.e., ammonia loadings and operating temperature, mainly), excepting methanol loadings; these were further decreased to match progressively decreasing denitrification performance. In addition, the still relatively high alkalinity additions into the systems (i.e., still of about 90 g CaC0 3 and over) were interpreted as a sign of decreased denitrification and, if not complete, yet adequate nitrification. Starting with the early stages of the 10 °C ambient temperature phase period, phenomena, that had never been experienced before, occurred and generated frequent operating problems. Mixed liquor within all reactors became more viscous, with periodic plugging of the exit tubes of the Aerobic #2 reactors and clarifiers. Excessive foaming within the Aerobic #1 reactors and exit clogging incidents caused "short-circuits" within the systems, on a regular base. Therefore, the data collected during the 10 °C temperature period, may also reflect the effect of the above mentioned operating problems on system performance. By the 21 s t day of the phase, both systems were generating effluent ammonia concentrations of over 1000 mg N/L. Assuming that a decrease in the influent ammonia concentration would enhance nitrification recovery within the systems, the mcoming ammonia concentrations were decreased from 2200 mg N/L to about 1200 mg N/L (i.e., on phase day no. 21). The decreased effluent ammonia concentrations (i.e., to about 270 mg N/L and 70 mg N/L, respectively), within a couple of days, were interpreted as signs of nitrification recovery. Therefore, following the procedure employed during the pH phase system recovery 135 period (i.e., proved to be effective as nitrification recovered within both systems), the influent ammonia concentration was increased in two successional steps, over a period of 14 days. Unfortunately, the nitrification process did not recover and effluent ammonia concentrations progressively increased up to about the initial values (i.e., 1094 mg N/L and to 1056 mg N/L, respectively, on the 47 th day of the phase. Furthermore, a 5 °C increase in the ambient temperature (i.e., on day no. 47, after sampling), did not appear to significantly affect the performance of the systems. In addition, Aerobic #1 wasting was started on the 40 t h day of the phase, at a rate of 0.500 L/day (i.e., SRT = 20 days), and than, continued at a rate of 0.250 L/day (i.e., SRT = 40 days). The daily wasting decreased the frequency of the clogging problems; however, it might have also decreased the chance of system recovery. On July 9, 1999, day no. 51 of the phase and day no. 311 of the study, the loadings to the systems were stopped and the experiments ceased. 5.4.1 Ammonia Levels and Aerobic #1 Reactor %Nitrification Figures 5.64 and 5.65 show that Aerobic #1 ammonia levels were consistently zero at ambient temperatures of 17 °C and that progressively increased starting with 14 °C. From Figures 5.66 and 5.67, it can be observed that a slight decrease in Anoxic #1 percentage denitrification started at ambient temperatures of 17 °C, while the decrease in Aerobic #1 percentage nitrification started only at about 10 °C. The decreased Anoxic #1 percentage denitrification values corresponded to Aerobic #1 NOx progressively increasing, as the NOx escaping treatment in the Anoxic #1 reactor was continuously returned into the Aerobic #1 reactor. Even though incipient denitrification inhibition probably occurred prior to nitrification inhibition (i.e., before the 7 t h day of the phase), the extended and serious failure of the nitrification process within both systems was primarily the result of decreased temperature and not of denitrification inhibition or methanol bleeding into the Aerobic #1 reactor (i.e., methanol loadings were progressively adjusted to meet denitrification decreased requirements) and enhanced heterotrophic growth, as noted earlier in the study. A decrease of approximate 1000 mg N/L in the influent ammonia concentration resulted in decreased effluent concentration (Figures 5.64 and 5.65). The decreased incoming ammonia concentration, indeed, 137 (-|/N Bin) uoijejiuaouoo eiUOUIUJV mol 9JBLiDB3-| p3JB|niUIS (l/N Bui) uouej)ua3UOQ eiuouiuiv I B i o i UJanujg ^ jojOBay uo!ie3M,u)!U8a% i# oixouv pue UOIIBOU,UI!N% V# ojqojav uoneouuj!U8a°/01# oixouv pue uoi)eayu)!N% V# oiqojav 140 resulted in an increased System #2 percentage nitrification value of 80%; however, System #1 percentage nitrification achieved only a maximum value of 21% and the System #1 decreased effluent ammonia concentrations were primarily due to dilution and bacterial assimilation within the system (Figures 5.66 and 5.67). Therefore, the assumption that decreased effluent ammonia concentrations were a result of incipient nitrification recovery within both systems was erroneous. The higher System #2 percentage nitrification (compared to System #1 percentage) may have been the direct result of the longer Aerobic #1 AHRT of the system (i.e., of 1.7 hours versus that of 1.5 hours of System #1); this probably stimulated nitrifiers growth and enhanced nitrification recovery. Nevertheless, increased influent ammonia concentration to initial values (i.e., from 1200 mg N/L back to 2200 mg N/L), at these low operating temperatures, resulted in large-scale nitrification inhibition within both systems (i.e., system percentage nitrification of a maximum 14% and 31%, respectively). 5.4.2 Methanol Loading Effects on Percentage Nitrification and Denitrification The principal ammonia removal method of the systems was the nitrification process occurring mainly in the Aerobic #1 reactors. The nitrification process of the Aerobic #1 reactor could have been influenced, to a certain extent, by the denitrification process occurring in the preceding anoxic reactor (i.e., Anoxic #1 reactor). Anoxic #1 decreased denitrification performance, under consistent methanol loading, might have resulted in methanol bleeding into the Aerobic #1 reactor; hence, there was the possibility of increased heterotrophic growth and nitrification inhibition as a result of increased competition (i.e., between autotrophic and heterotrophic bacteria) for either oxygen (Hanaki et al., 1990b) or ammonia (Carley and Mavinic, 1991). Figure 5.68 shows that System #1, Aerobic #1 percentage nitrification was virtually independent of reactor COD levels of up to about 330 mg/L and this decreased only when the COD levels increased over 330 mg/L (i.e., even though no special trend can be observed). However, increased Aerobic #1 COD concentrations occurred only in the second half of the phase period. Therefore, System #1 nitrification inhibition was mainly a result of decreased temperature and reactor COD concentrations did not have any 141 Figure 5 .68: S y s t e m #1 - T e m p e r a t u r e P h a s e A e r o b i c #1 C O D L e v e l s v e r s u s % N i t r i f i c a t i o n 1000 T -O) c o •0 0 ) o c o o o O O *J— % o 15 o 900 800 700 600 500 400 300 5 200 9 < 100 —I 1 1 1 1 •—I 1— 10 20 30 40 50 60 70 %Nitrification 80 90 100 110 120 1000 x 900 4--Figure 5.69: S y s t e m #2 - T e m p e r a t u r e P h a s e A e r o b i c #1 C O D L e v e l s v e r s u s % N i t r i f i c a t i o n O) E c o 800 700 •£ 600 a> o c o o Q O O MT-% O 15 2 a> < 500 400 300 200 100 4-—I 1 1 1 1 1 h -10 20 30 40 50 60 70 %Nitrification 80 90 100 110 120 oneu D U ! " ' U 3 X O N : HOCHO L# OIXOUV 143 uoueouujiuarjo/o (,# apcouv pue uone3Mj)!N% i.# ojqojav I 1 1 1 1- o CO CD CM O oiiey • * U 3 X O N : H O C H O l# oixouv 144 apparent influence on nitrification performance during the first half of the phase period. Nevertheless, increased reactor COD levels may have limited the nitrifiers' capability to recover at the end of the phase. In addition, Figure 5.69 shows that System #2, Aerobic #1 percentage nitrification values between 30% and 85% were achieved concomitantly with reactor COD levels of about 400 mg/L and that lower percentage nitrification values were actually achieved under much lower reactor COD values; hence, reactor COD levels evidently did not affect Aerobic #1 nitrification performance in System #2. Therefore, it was concluded, again, that decreased nitrification performance was a direct result of nitrification inhibition to low operating temperature, rather than a result of increased heterotrophic growth within the Aerobic #1 reactor (i.e., progressively decreased methanol loading, once denitrification inhibition was observed, minimized methanol bleeding into the Aerobic #1 reactor, hence, decreased ambient temperature became the determinant factor in nitrification inhibition). Figures 5.70 and 5.71 show that, during the first half of this phase, the CH3OH:NOxEntering ratio was progressively decreased to match decreased denitrification requirements for organic carbon. During the second half of the phase, the ratios were slightly increased; this was expected to stimulate, at least, the partial recovery of denitrification. Increased CH30H:NOx E n t e r i n g ratios did not have any obvious beneficial effect on denitrification performance; however, it might have stimulated aerobic heterotrophic bacterial population growth within all rectors and could have limited, both, nitrifying and denitrifying bacterial populations in their ability to recover. 5.4.3 VSS Levels and Specific Nitrification and Denitrification Rates Decreased ambient temperatures resulted in lower nitrification and denitrification rates. However, Figures 5.72 and 5.73, and Figures 5.74 and 5.75 do not show any evident trend of specific nitrification and denitrification rates, respectively, as a function of temperature. Specific utilization rates decreased considerably, once the ambient temperature was 10 °C and the VSS concentration of all reactors decreased (i.e., mainly as a result of decreased bacterial growth rates). Nevertheless, at 10 °C, the rates were erratic and fairly independent of the fluctuation in the VSS levels. Rises in specific nitrification rates may be 145 146 147 148 149 150 151 correlated with two influential changes in the operating parameters that occurred during the second half of the phase period: the decrease of the influent ammonia concentration and the increase of the ambient temperature. However, given the relatively short period of operating time under these particular parameters, the results are not conclusive. Furthermore, the VSS concentration considerably varied from one reactor to another and had specific independent responses to operational changes, once the ambient temperature reached 10 °C. The decreased biomass levels and specific utilization rates of the rectors were not only the result of lower bacterial growth rates under decreased temperatures, but also the result of bacterial inhibition to low temperatures and possibly increased concentrations of suppressing constituents. By the end of the pH phase (i.e., previous phase), as a result of decreased anoxic pH levels (hence, decreased "free" ammonia levels), the N0 2/NOx ratios of the systems decreased from about 0.8 to about 0.4. During the 10 °C period of this phase, the N0 2/NOx ratios of the systems progressively increased up to the value of 0.75, as a result of nitrate forming organisms inhibition at low temperatures. Figures 5.76 and 5.77 confirm the fact that nitrite accumulation was mainly a result of decreased temperatures, since the anoxic pH values were considerable reduced; hence, "free" ammonia bacterial inhibition was minimal (Anthonisen et al., 1976; Randall and Buth, 1984). Increased levels of nitrous acid (HN0 2) associated with low pH levels, may have further inhibited nitrate forming organisms (Anthonisen et al. 1976), as well as, the denitrifying organisms of the anoxic zones (i.e., Beccari et al., 1993). Decreased bacterial growth rates and persistent bacterial inhibition, in addition to somewhat increased organic carbon levels favoring aerobic heterotrophic growth, may have been the prime factors that prevented nitrification and denitrification processes from acclimatizing to the decreased ambient temperatures. 5.4.4 Effluent Total Ammonia and NOx Levels Figures 5.78 and 5.79 show that, once the ambient temperature was decreased to 10 °C, the effluent ammonia concentrations progressively increased. Also, the corresponding increase in effluent NOx concentrations were a result of decreased denitrification performance, rather than improved nitrification. 152 IBAouiay ua6oj}i|s| 0|UB6JOU| |BJOJ.% pue iBAouiay eiuouiuiv ie»oj.% uiajsAs 153 (l/N 6ui) suojjBJjuaouoo X Q N pue eiuouiuiv lejoi Hjanurg o o o CO o o in CM o o o o o m 154 The NOx concentrations remained fairly high during the entire 10 °C phase period, due to still existent, even though limited, nitrification, and virtually no denitrification. The effluent ammonia concentrations decreased in response to decreased influent ammonia concentrations (i.e., influent ammonia concentrations decreased from 2200 mg N/L to 1200 mg N/L, on phase day no. 21). However, once the incoming ammonia levels were increased to the previous values, the effluent ammonia concentrations progressively increased up to about the initial values. The effluent ammonia levels of the last day of the phase are primarily the result of dilution of incoming ammonia within the systems, ammonia removal by aerobic heterotrophic bacterial assimilation, and some nitrification. Therefore, the percentage of total ammonia removal rate was maintained at an average minimum of about 50% and the systems, even though they failed the targeted performance (i.e. > 90% removal), never completely collapsed. 155 Chapter 6 SUMMARY, CONCLUSIONS AND RECOMMENDATIONS 6.1 Summary The prime objective of this research was to investigate the nitrogen removal capabilities of a nitrification process, using pre- and post-denitrification, when treating landfill leachate containing "simulated" ammonia concentrations of about 2200 mg N/L. Two, 4-Stage Bardenpho systems were operated in parallel during the entire experimental period. The external recycle ratio was 3:1 for both systems, while the internal recycle ratio was 4:1 for one system (i.e., System#l), and3:l for the other system (i.e., System #2). The first system had a first anoxic reactor actual hydraulic retention time of 1.5 hours, while the second system had one of 1.7 hours. The first aerobic reactor actual hydraulic retention time was 3 hours for the first system, and 3.4 hours for the second one. The Base Leachate Phase main purpose was to achieve complete nitrification and denitrification of the "base" landfill leachate, containing an average ammonia concentration of 200 mg N/L. The performance of the systems, under successional, incremental ammonia increases, was studied during the Ammonia Loading Phase. Once stability of systems treating mcoming ammonia concentrations of over 2200 mg N/L was observed, denitrification performance was thought to be improved by decreasing the anoxic pH levels, during the pH Phase. The Temperature Phase examined the overall process performance at progressively decreased ambient temperatures, from 20 °C to 10 °C. The most important findings of the four phases of this research are as follows: 1. Complete ammonia removal from the "base" landfill leachate was achieved within 20 days from the begiruiing of the study. The principal ammonia removal mechanism was the nitrification process within the first aerobic reactor; however, bacterial assimilation also played an important role. Effluents free of nitrites and nitrates were generated within 65 days from the beginning of the study. 156 The final CH3OH:NOx ratios required to achieve complete denitrification were unexpectedly high (i.e., about 20:1). 2. The first attempt at incrementally increasing influent ammonia concentrations failed after four incremental increases (each of about 300 mg N/L of ammonia) made at 7 day intervals. Both systems experienced nitrification inhibition at incoming ammonia levels of about 1300 mg N/L and under methanol loadings increased concomitantly with ammonia loadings, to match expected aerobic NOx production (using the CH 3OH:NOx ratios established during the first phase). 3. Nitrification inhibition was a result of two influential factors that concomitantly generated immediate and undesirable responses in bacterial performance. The first factor was increased ammonia concentration within the systems. Increased ammonia concentration in the Aerobic #1 reactor also increased nitrification and, subsequently, sodium bicarbonate addition, hence decreased reactor hydraulic retention time; decreased aerobic actual hydraulic retention time appeared to hinder the growth of nitrifiers. Increased Anoxic #1 reactor ammonia levels, associated with increased pH levels, further suppressed nitrifying bacterial population performance. The recycle of nitrifiers through the elevated "free" ammonia concentrations of the Anoxic #1 reactor effected the nitrate production process, hence, nitrite accumulation. The second factor was methanol loading increases, initiated prior to allowing sufficient time for the establishment of a sustained NOx production, after each incoming ammonia increase. In addition, the methanol loadings adjustments were made without considering the probable nitrite accumulation (i.e., lower organic carbon requirements for nitrite reduction). Therefore, the excess methanol in the Anoxic #1 reactor flowed to the Aerobic #1 reactor and was responsible for enhanced aerobic heterotrophic growth and further nitrifier inhibition. The timing of ammonia and methanol loading increases, with respect to each other and to the corresponding previous loading increase, was decisive; still, the amount of the loading increase also played an important role in nitrification failure. 157 4. During the second attempt at incrementally increasing the influent ammonia concentration, both systems generated ammonia free influents after each incremental increase. The systems were allowed to acclimatize to each incremental ammonia increase for about 14 days and only afterwards, the methanol loadings were increased to match NOx production and meet denitrification requirements for organic carbon. Furthermore, the methanol loading was implemented to yield lower CFf3OH:NOx ratios (i.e., CH 3OH:NOx removed ratios of about 5:1). Within 88 days from the start of the second attempt at increasing influent ammonia concentrations, the systems produced ammonia free effluents with NOx concentrations of 78 mg N/L (i.e., System #1) and 250 mg N/L (i.e., System #2), respectively, when treating simulated landfill leachate with a concentration of over 2200 mg N/L of ammonia. The higher effluent NOx concentration of the second system (compared to the one of the first system) was probably a direct result of the decreased overall denitrification performance of the second system (i.e., possibly due to stronger denitrification inhibition to elevated "free" ammonia levels, since the Anoxic #1 contact time, of 1.7 hours, of the second system, was longer than the one, of 1.5 hours, of the first system). 5. Complete denitrification of the NOx loadings was not accomplished in the Anoxic #1 reactors, due to denitrifying organisms inhibition to elevated "free" ammonia levels and very high pH levels (i.e., pH values of about 8.6). Complete denitrification was also not achieved within the Anoxic #2 reactors, mainly as a result of denitrifier inhibition to extremely high pH levels (i.e., pH values of over 9.6). 6. The overall denitrification performances of the systems started to respond positively to decreased anoxic pH levels (i.e., Anoxic #1 pHs of about 8.4, and Anoxic #2 pHs of about 9.2). However, subsequent changes in the "base" landfill leachate constituents initiated extensive denitrification inhibition within all reactors; therefore, the final percentage denitrification rates were only about 50%. Both systems produced final effluent NOx concentrations of about 170 mg N/L. 158 7. The decreased Anoxic #1 pH levels, hence, decreased "free" ammonia levels, affected a reduction in the N0 2/NOx ratio, hence, reduced nitrifier inhibition. Decreased nitrifying bacteria inhibition resulted in increased nitrification performance and complete ammonia removal in the Aerobic #1 reactor. Furthermore, the systems were able to handle increases in the influent ammonia concentration of over 300 mg N/L, without any evident problem. 8. Nitrification processes of the systems were not affected by decreased ambient temperature from 20 °C to 17 °C, and, subsequently, to 14 °C. Under ambient temperatures of 10 °C, the percentage nitrification rates of the systems decreased to about 10% for System #1, and to 30% for System #2. Nitrification failure was primarily the result of decreased bacterial growth rates and nitrifier inhibition to low temperatures. Incipient denitrification inhibition was observed, starting with the ambient temperature of 17 °C. Percentage denitrification rates of all reactors decreased considerably to less than 5 percent, once the ambient temperature reached 10 °C. A final percentage total ammonia removal rate of about 50% was maintained by both systems. 9. The overall performance of System #1 (i.e., with Anoxic #1 AHRT of 1.5 hours, and Aerobic #1 AHRT of 3 hours ) was more effective than the overall performance of System #2 (i.e., with Anoxic #1 AHRT of 1.7 hours, and Aerobic #1 AHRT of 3.4 hours) during almost the entire experimental period. System #1 usually generated ammonia free effluents with lower NOx concentrations during stability periods (i.e., 78 mg N/L in comparison with 250 mg N/L, at the end of the second study phase, and 170 mg N/L in comparison with 177 mg N/L, at the end of the third study phase). Furthermore, the concentrations of the constituents of interest within System #1 reactors were consistently equal, while within System #2, there was a great deal of fluctuation, even during periods of constant operating parameters and conditions. Therefore, it seems that the higher first anoxic and aerobic hydraulic retention times of System #2 (compared to those of System #1), not only negatively affected the overall denitrification performance of the system (i.e., denitrifiers experienced stronger 159 inhibition, since, within the Anoxic #1 reactor, they were exposed to elevated "free" ammonia levels for a longer period of time), but also enhanced the release of nitrogen through cell lysis (i.e., organic nitrogen, eventually converted into inorganic nitrogen compounds; hence, surges in the concentrations of the constituents of interest). However, System #2 performance was more stable during "stressful" periods, such as that of the Ammonia Loading Phase nitrification failure period, when ammonia accumulation in System #2 was considerably lower than in System #1 (i.e., Anoxic #1 ammonia accumulation of about 353 mg N/L in comparison with 1555 mg N/L). It appears that, during these particular periods, the higher first anoxic and aerobic hydraulic retention times of System #2 (compared to those of System #1) enhanced, to a certain extent, the stability of the system. 6.2 Conclusions From the performance of the pre- and post-denitrification biological process, tested for nitrogen removal capabilities, when treating landfill leachate with "simulated" ammonia concentration of over 2200 mg N/L, and using methanol as supplementary organic carbon source for denitrification, resulted the following conclusions'. 1. In the reactors of the investigated pre- and post- denitrification systems, at least two bacterial populations coexisted (e.g., the autotrophic nitrifiers and the heterotrophic aerobes, in the first aerobic reactor). One population was the majority, or the "dominant" population, which was the one most favored by reactor operational parameters and existing conditions (i.e., type and levels of the constituents within the reactor, pH and alkalinity levels, dissolved oxygen levels in correlation with anoxic and aerobic conditions, ambient temperature). The other population was the minority, or "in waiting" population, which barely survived, waiting for more favorable conditions. 2. Changes in the operational parameters and conditions, under which a reactor performed, generated switches in the equilibrium under which the bacterial populations of this reactor coexisted. Reasonable operational changes enhanced the "in waiting" population activity that completed, rather 160 than outcompeted, the "dominant" population performance (e.g., slightly increased influent ammonia concentrations enhanced ammonia removal through cell synthesis of heterotrophic aerobes within the first anoxic reactor, without affecting denitrification performance). Significant operational changes, increasingly favorable to the minority, stimulated this "in waiting" population of organisms to a point where the "in waiting" bacterial population became the "dominant" bacterial population, resulting in reactor failure (e.g., increasing carbon levels in the first aerobic reactor, not only initiated enhanced heterotrophic population growth, but eventually resulted in a heterotrophic majority and subsequent nitrification failure). 3. Bacterial activity, hence, system performance, was regulated by, both, loadings to the system and corresponding concentrations of constituents vvdthin reactors, at any one time. Therefore, at low ammonia loadings, the performance of the systems was immediately unbalanced even by relative small changes in the concentrations of the reactor constituents, while at high ammonia loadings, the performance of the systems remained stable under similarly increased reactor concentrations (e.g., at influent ammonia concentrations of over 2200 mg N/L, increases of over 300 mg N/L in the incoming ammonia did not have any obvious effect on the performance of the system). 4. Bacterial populations of the reactors appeared to have a "collective memory". Once acclimatized to certain operational parameters and conditions, they acclimatized faster for the second time around to similar parameters and conditions, provided that a relatively short time passed between the two acclimatization periods. 5. In the explored pre- and post-denitrification systems, the biological activity within each antecedent reactor, hence, reactor performance, strongly influenced the performance of each subsequent reactor. The most influential reactor performance was the performance of the first aerobic reactor that controlled not only the performance of the subsequent anoxic reactor, but also the performance of the antecedent anoxic reactor, due to the internal recycle line. 6. The higher Aerobic #1 AHRT of System #2 (i.e., of 3.4 hours compared to that of 3 hours of System #1) did not seem to have any beneficial effect on neither, nitrification performance within itself, nor on the overall performance of the system, during stability periods. On the contrary, the longer AHRTs of the System #2 might have promoted abundant cell lysis and releases of nitrogen. However, during "stressful" periods (e.g., decreased ambient temperature), and, probably as a direct result of its longer AHRTs, the overall performance of System #2 was, not only more stable, but also more efficient. 7. As a result of the influence that one reactor performance had over another reactor, changing one operating parameter or condition in one reactor determined "chain-reaction" responses within the system (e.g., decreased Aerobic # 1 pH level resulted in decreased anoxic pH levels, hence, decreased Anoxic #1 "free" ammonia levels; as a result, there was reduced nitrifier inhibition and an overall increase in Aerobic #1 nitrification performance). 8. Both systems were able to generate ammonia free effluents, not only during stability periods, but also after each incremental ammonia increase, during the second attempt to increase influent ammonia concentration. When treating "simulated" landfill leachate with concentrations of over 2200 mg N/L of ammonia, the ammonia removal performance of the systems was effective and stable even under influent ammonia surges of over 300 mg N/L. 9. Decreased pH levels within the reactors improved the overall performances of the systems. However, as a result of denitrifier inhibition due to combined suppressants (i.e., relatively still high pH levels combined with increased toxicity of the leachate), complete denitrification was never achieved. The systems generated effluent NOx concentrations of about 170 mg N/L and, therefore, percentage total inorganic nitrogen removal was only 90%, while percentage total ammonia removal was 100%. 162 10. The systems experienced incipient denitrification inhibition starting with the ambient temperature of 17 °C, while nitrification inhibition was observed only at about 10 °C. Denitrification inhibition was not expected to occur ahead nitrification inhibition, since, according to the classic theory, nitrifiers (i.e., autotrophs) are more sensitive than denitrifiers (i.e., heterotrophs), to decreases in ambient temperature. As a result of this inhibition and overall process performance, percentage total ammonia removal decreased to about 50%, within both systems. 11. The nitrification process train, with pre- and post-denitrification (i.e., 4-Stage Bardenpho Process), was found to be an effective nitrogen removal method for high ammonia landfill leachates. The process was capable of easily accommodating fluctuations, typical for landfill leachates, in influent characteristics and to efficiently adjust to changes in operational parameters and conditions. Nevertheless, a means of improving denitrification overall performance should be given thought, through further research. In addition, process performance and stability at ambient temperatures of 10 °C, when treating influent ammonia concentrations of over 2200 mg N/L, should be also further investigated. 12. This particular type of biological treatment process may be extremely effective when employed as a nitrogen removal method for other industrial wastes, as they are characterized by very high total inorganic nitrogen concentrations and, probably, low organic carbon levels. However, before deciding whether or not to use such a process, pilot plant trials should be conducted, since some of the constituents of one particular industrial discharge might negatively effect bacterial performance. 6.3 Recommendations Based on the results of the four phases of this research and following up on the conclusions, the recommendations that can be made are: 163 1. The balanced coexistence of bacterial population within the reactor and the reciprocal influences between reactors performances, in a sequenced pre- and post-denitrification process, should not be viewed negatively. Instead, they should be further explored as useful inner-system, self-adjusting mechanisms, that enable the process to properly alter its performance, in response to reasonable changes in operational parameters and existing conditions. The limits of one change could be investigated by increasing and decreasing the variable of interest, while observing process stability, under otherwise constant operational parameters. 2. Further investigations should be conducted to determine appropriate ways of maintaining anoxic pH levels of maximum 8, when treating influent ammonia concentrations of over 2200 mg N/L. Decreased anoxic pH levels will considerably improve denitrification performance and will enhance nitrification stability. Relatively low anoxic pH levels can be maintained by acclimatizing the nitrifying bacterial population of the first aerobic reactor to pH levels below the value of 7, starting with the first step-increase in ammonia loading. 3. The capability of this particular pre-and post denitrification system to treat influent ammonia concentrations higher than 2200 mg N/L should be also explored. Assuming that increased overall denitrification performance is possible, and using similar actual hydraulic retention times (i.e., except that the second aerobic reactor hydraulic retention time should be a maximum 1.5 hours) and recycle ratios (i.e., 4:1 internal, and 3:1 external), the system may be able to generate ammonia and NOx free effluents, when treating incoming ammonia concentrations of over 2200 mg N/L. 4. The effect of decreased temperature on system performance, when treating influent ammonia concentration of over 2200 mg N/L, should be further investigated. Long acclimatization periods after each temperature decrease, correlated with small step-decreases, may be helpful in determining critical points (i.e., ambient temperature) for nitrification and denitrification inhibition and suggest methods to prevent process failure. REFERENCES 164 Albertsen, A, Holz, F., and Martens, J. (1998). Membrane Separation Processes for Leachate Treatment from Landfill Sites, Proceedings of the International Training Seminar: Management and Treatment of MSW Landfill Leachate, Venice, Italy, 2-4 December 1998, CISA, Sanitary Environmental Engineering Centre. Anthonisen, A C , Loehr, R.C., Prakasam, T.B.S., and Srinath, E.G. (1976). Inhibition of Nitrification by Ammonia and Nitrous Acid, Journal of the Water Pollution Control Federation, 48(5), Pages 835-852. APHA (1993). Standard Methods for the Examination of Water and Wastewater. 18th Edition, American Public Health Association, Washington, D C . Argaman, Y. (1982). 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Denitrification of a High Ammonia Leachate Using an External Carbon Source, Environmental Technology Letters, 10, Pages 701-716. Manoharan, R, Harper S.C., Mavinic, D.S., Randall. C.W., Wang, G., and Marickovich, D.C. (1992). Inferred Metal Toxicity During the Biotreatment of High Ammonia Landfill Leachate, Water Environment Research, 64(7), Pages 858-865. Mavinic, D.S. and Randall, C.W. (1990). Inhibition of Nitrification and Denitrification in Biotreatment of a High-Ammonia Municipal Leachate, Report prepared for Environment Canada and Virginia Environmental Endowment Fund, Department of Civil Engineering, University of British Columbia, Vancouver, Canada. Mavinic, D.S. (1997). Biological Waste Treatment, CIVL 569 Course, Department of Civil Engineering, University of British Columbia, Vancouver, Canada. Mavinic, D.S. (1998). Leachate Quality: Effects on Treatability, Proceedings of the International Training Seminar: Management and Treatment of MSW Landfill Leachate, Venice. 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CISA, Sanitary Environmental Engineering Centre. Sawyer, C.N., McCarty, P.L., and Parkin, G.F. (1994). Chemistry for Environmental Engineering. Fourth Edition, New York: McGraw-Hill, Inc. Shiskowski, D.M. (1995). Biological Nitrification and Denitrification of High Ammonia Landfill Leachate Using Pre Denitrification and Pre/Post Denitrification Processes, M. A. Sc. Thesis, Department of Civil Engineering, University of British Columbia, Vancouver, Canada. Shiskowski, D.M. and Mavinic, D.S. (1998a). Biological Treatment of a High Ammonia Leachate: Influence of External Carbon During Initial Startup, Water Research, 32(8), Pages 2533-2541. Shiskowski, D.M. and Mavinic, D.S. (1998b). Pre-Denitrification and Pre- and Post- Denitrification Treatment of High-Ammonia Landfill Leachate, Canadian Journal of Civil Engineering, 25, Pages 854-863. Turk, O. and Mavinic D.S. (1989). Maintaining Nitrite Build-up in a System Acclimatized to Free Ammonia, Water Research, 23(11), Pages 1383-1388. Personal Communications: Mavinic, D.S. (1999), Ph.D., P.Eng., Professor of Civil Engineering, Department of Civil Engineering, University of British Columbia, Vancouver, Canada. APPENDIX A: Formulas and Calculation Definitions & Explanatory Notes 169 Formulas and Calculation Definitions T H E N E G A T I V E L O G A R I T H M O F T H E I O N I Z A T I O N C O N S T A N T F O R A M M O N I U M (pKA) A T D I F F E R E N T T E M P E R A T U R E S : TEMPERATURE TEMPERATURE pK A (°C) (°K) 20 293 9.41 17 290 9.50 15 288 9.57 14 287 9.60 10 283 9.74 where: TEMPERATURE °K = 273 + TEMPERATURE °C pK A = 0.09018 + 2729.92/TEMPERATURE °K (Beriefield et al, 1982) ESTIMATED FREE AMMONIA (mg N/L) = TOTAL AMMONIA (mg N/L) * 10 p H /[10 p K A + 10 p H ] (Benefield et al, 1982; Sawyer et al, 1994) SIMULATED LEACHATE FLOW (L/day) = NH4C1 ANOXIC #1 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) + BASE LEACHATE FLOW (L/day) ANOXIC #1 TOTAL FLOW (L/day) = SIMULATED LEACHATE FLOW (L/day) + Na3P04 ANOXIC #1 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) + CH30H ANOXIC #1 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) SYSTEM TOTAL FLOW (L/day) = ANOXIC #1 TOTAL FLOW (L/day) + NaHC03 AEROBIC #1 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) + CH30H ANOXIC #2 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) ANOXIC #1 OVERFLOW (L/day) = ANOXIC #1 TOTAL FLOW (L/day) + INTERNAL RECYCLE FLOW (L/day) + EXTERNAL RECYCLE FLOW (L/day) AEROBIC #1 OVERFLOW (L/day) = ANOXIC #1 OVERFLOW (L/day) + NaHC03 AEROBIC #1 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) ANOXIC #2 OVERFLOW (L/day) = AEROBIC #1 OVERFLOW (L/day) - INTERNAL RECYCLE FLOW (L/day) + CH30H ANOXIC #2 FEED FLOW (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) AEROBIC #2 OVERFLOW (L/day) = ANOXIC #2 OVERFLOW (L/day) ANOXIC #1 AHRT (hr) = [5 (L)/ANOXIC #1 OVERFLOW (L/day)] * 24 (hr/day) AEROBIC #1 AHRT (hr) = [10 (L)/AEROBIC #1 OVERFLOW (L/day)] * 24 (hr/day) ANOXIC #2 AHRT (hr) = [5 (L)/ANOXIC #2 OVERFLOW (L/day)] * 24 (hr/day) AEROBIC #2 AHRT (hr) = [10 (L)/AEROBIC #2 OVERFLOW (L/day)] * 24 (hr/day) CLARIFIER AHRT (hr) = [4 (L)/AEROBIC #2 OVERFLOW (L/day)] * 24 (hr/day) 170 A E R O B I C #1 S R T (days) = 10 ( L ) / A E R O B I C #1 W A S T I N G (L/day) S Y S T E M S R T (days) = [5 ( L ) * A N O X I C #1 V S S (mg/L) + 10 ( L ) * A E R O B I C #1 V S S (mg/L) + 5 ( L ) * A N O X I C #2 V S S (mg/L) + 15 ( L ) * A E R O B I C #2 V S S (mg/L)]/{AEROBIC #1 W A S T I N G (L/day) * A E R O B I C #1 V S S (mg/L) + [ S Y S T E M T O T A L F L O W (L/day) - A E R O B I C #1 W A S T I N G (L/day)] * E F F L U E N T V S S (mg/L)} INTERNAL RECYCLE R A T I O (L/day) = INTERNAL RECYCLE F L O W (L/day)/SIMULATED L E A C H A T E F L O W (L/day) EXTERNAL RECYCLE R A T I O (L/day) = EXTERNAL RECYCLE F L O W (L/day)/SlMULATED L E A C H A T E F L O W (L/day) NH4C1 F E E D C O N C E N T R A T I O N (g NH4C1/L) = NH4C1 F E E D C O N C E N T R A T I O N (g N/L)/0.2618 (g N/g NH4C1) NaHC03 F E E D C O N C E N T R A T I O N (g CaC03/L) = NaHC03 F E E D C O N C E N T R A T I O N (g NaHC03/L) * 50/84 (g CaC03/g NaHC03) Na3P04 F E E D C O N C E N T R A T I O N (g Na3P04/L) = Na3P04 F E E D C O N C E N T R A T I O N (g P/L)/0.0815 (g P/g Na3P04) A N O X I C #1 CH30H F E E D C O N C E N T R A T I O N (g COD/L) = A N O X I C #1 CH30H F E E D C O N C E N T R A T I O N (ml CH30H/L) * 0.7915 (g CH30H/ml CH30H) * 1.5 (g COD/g CH30H) A N O X I C #2 CH30H F E E D C O N C E N T R A T I O N (g COD/L) = A N O X I C #2 CH30H F E E D C O N C E N T R A T I O N (ml CH30H/L) * 0.7915 (g CH30H/ml CH30H) * 1.5 (g COD/g CH30H) SIMULATED L E A C H A T E T O T A L AMMONIA (mg N / L ) = [NH4C1 F E E D C O N C E N T R A T I O N (g N / L ) * NH4C1 A N O X I C #1 F E E D F L O W (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) * 1000 (mg/g) + L E A C H A T E T O T A L AMMONIA (mg N / L ) * B A S E L E A C H A T E F L O W (L/day)]/SlMULATED L E A C H A T E F L O W (L/day) A N O X I C #1 E S T I M A T E D T O T A L A M M O N I A (mg N / L ) = S I M U L A T E D L E A C H A T E T O T A L A M M O N I A (mg N / L ) * S I M U L A T E D L E A C H A T E F L O W (L/day)/ANOxic #1 O V E R F L O W (L/day) T O T A L A M M O N I A L O A D (g N/day) = S I M U L A T E D L E A C H A T E T O T A L A M M O N I A (mg N / L ) * S I M U L A T E D L E A C H A T E F L O W (L/day) * 1/1000 (g/mg) A L K A L I N I T Y L O A D (g CaC03/day) = L E A C H A T E A L K A L I N I T Y (mg CaC03/L) * B A S E L E A C H A T E F L O W (L/day) * 1/1000 (g/mg) + NaHC03 F E E D C O N C E N T R A T I O N (g CaC03/L) * NaHC03 A E R O B I C #1 F E E D F L O W (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) 0-P04 L O A D (g P/day) = L E A C H A T E 0-P04 (mg P/L) * B A S E L E A C H A T E F L O W (L/day) * 1/1000 (g/mg) + Na3P04 F E E D C O N C E N T R A T I O N (g P/L) * Na3P04 A N O X I C #1 F E E D F L O W (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) A N O X I C #1 CH30H L O A D (g COD/day) = A N O X I C #1 CH30H F E E D C O N C E N T R A T I O N (g COD/L) * CH30H F E E D F L O W (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) A N O X I C #2 CH30H L O A D (g COD/day) = A N O X I C #2 CH30H F E E D C O N C E N T R A T I O N (g COD/L) * CH30H F E E D F L O W (ml/hr) * 24 (hr/day) * 1/1000 (L/ml) 171 A N O X I C #1 E X P E C T E D T O T A L A M M O N I A (mg N / L ) = [ S I M U L A T E D L E A C H A T E T O T A L A M M O N I A (mg N / L ) * S I M U L A T E D L E A C H A T E F L O W (L/day) + A E R O B I C #1 T O T A L A M M O N I A (mg N / L ) * I N T E R N A L R E C Y C L E F L O W (L/day) + A E R O B I C #2 T O T A L A M M O N I A (mg N / L ) * E X T E R N A L R E C Y C L E F L O W ( L / d a y ) ] / A N O X I C #1 O V E R F L O W (L/day) A N O X I C # 1 T O T A L A M M O N I A R E M O V A L R A T E (mg N/day) = [ A N O X I C # 1 E X P E C T E D T O T A L A M M O N I A (mg N / L ) - A N O X I C #1 T O T A L A M M O N I A (mg N / L ) ] * A N O X I C #1 O V E R F L O W (L/day) A E R O B I C #1 T O T A L A M M O N I A R E M O V A L R A T E (mg N/day) = A N O X I C #1 T O T A L A M M O N I A (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day) - A E R O B I C #1 T O T A L A M M O N I A (mg N / L ) * A E R O B I C #1 O V E R F L O W (L/day) A N O X I C #2 T O T A L A M M O N I A R E M O V A L R A T E (mg N/day) = A E R O B I C #1 T O T A L A M M O N I A (mg N / L ) * [ A E R O B I C #1 O V E R F L O W (L/day) - I N T E R N A L R E C Y C L E F L O W (L/day)] - A N O X I C #2 T O T A L A M M O N I A (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day) A E R O B I C #2 T O T A L A M M O N I A R E M O V A L R A T E (mg N/day) = A N O X I C #2 T O T A L A M M O N I A (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day) - A E R O B I C #2 T O T A L A M M O N I A (mg N / L ) * A E R O B I C #2 O V E R F L O W (L/day) S Y S T E M T O T A L A M M O N I A R E M O V A L R A T E (mg N/day) = T O T A L A M M O N I A L O A D (g N/day) * 1000 (mg/g) - E F F L U E N T T O T A L A M M O N I A (mg N / L ) S Y S T E M T O T A L F L O W (L/day) % A N O X I C #1 T O T A L A M M O N I A R E M O V A L = { A N O X I C #1 T O T A L A M M O N I A R E M O V A L R A T E (mg N / day ) / [ A N O X I C# l E X P E C T E D T O T A L A M M O N I A (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day)]} * 100 % A E R O B I C #1 T O T A L A M M O N I A R E M O V A L = { A E R O B I C #1 T O T A L A M M O N I A R E M O V A L R A T E (mg N/day)/[ANOXIC #1 T O T A L A M M O N I A (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day)]} * 100 % A N O X I C #2 T O T A L A M M O N I A R E M O V A L = { A N O X I C #2 T O T A L A M M O N I A R E M O V A L R A T E (mg N / d a y ) / [ A E R O B I C #1 T O T A L A M M O N I A (mg N / L ) * [ A E R O B I C #1 O V E R F L O W (L/day) - I N T E R N A L R E C Y C L E F L O W (L/day)]]} * 100 % A E R O B I C #2 T O T A L A M M O N I A R E M O V A L = { A E R O B I C #2 T O T A L A M M O N I A R E M O V A L R A T E (mg N/day)/[ANOXIC #2 T O T A L A M M O N I A (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day)]} * 100 % S Y S T E M T O T A L A M M O N I A R E M O V A L = { S Y S T E M T O T A L A M M O N I A R E M O V A L R A T E (mg N/day)/[TOTAL A M M O N I A L O A D (g N/day) * 1000 (mg/g)]} * 100 A E R O B I C #1 N I T R I F I C A T I O N R A T E (mg N/day) = A E R O B I C #1 NOx (mg N / L ) * A E R O B I C #1 O V E R F L O W (L/day) - A N O X I C # 1 NOx (mg N / L ) * A N O X I C # 1 O V E R F L O W (L/day) A E R O B I C #2 N I T R I F I C A T I O N R A T E (mg N/day) = A E R O B I C #2 NOx (mg N / L ) * A E R O B I C #2 O V E R F L O W (L/day) - A N O X I C #2 NOx (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day) A E R O B I C #1 S P E C I F I C N I T R I F I C A T I O N R A T E (mg N/day / mg V S S ) = A E R O B I C #1 N I T R I F I C A T I O N R A T E (mg N/day)/[AEROBIC #1 V S S (mg/L) * 10 ( L ) ] A E R O B I C #2 S P E C I F I C N I T R I F I C A T I O N R A T E (mg N/day / mg V S S ) = A E R O B I C #2 N I T R I F I C A T I O N R A T E (mg N/day)/[AEROBIC #2 V S S (mg/L) * 10 ( L ) ] % A E R O B I C #1 N I T R I F I C A T I O N = { A E R O B I C #1 N I T R I F I C A T I O N R A T E (mg N/day)/[ANOXIC #1 T O T A L A M M O N I A (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day)]} * 100 172 % A E R O B I C #2 N I T R I F I C A T I O N = { A E R O B I C #2 N I T R I F I C A T I O N R A T E (mg N/day)/[ANOXIC #2 T O T A L A M M O N I A (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day)]} * 100 A L K A L I N I T Y : T O T A L A M M O N I A A D D E D (g CaC03/g N ) = A L K A L I N I T Y L O A D (g CaC03/day)/TOTAL A M M O N I A L O A D (g N/day) A L K A L I N I T Y : T O T A L A M M O N I A N I T R I F I E D (g CaC03/g N) = A L K A L I N I T Y L O A D (g CaCO3/day)/[AER0BlC #1 N I T R I F I C A T I O N R A T E (mg N/day) * 1/1000 (g/mg)] A N O X I C #1 NOx L O A D (mg N/day) = L E A C H A T E NOx (mg N / L ) * B A S E L E A C H A T E F L O W (L/day) + A E R O B I C #1 NOx (mg N / L ) * I N T E R N A L R E C Y C L E F L O W (L/day) + A E R O B I C #2 NOx (mg N / L ) * E X T E R N A L R E C Y C L E F L O W (L/day) A N O X I C #2 NOx L O A D (mg N/day) = A E R O B I C #1 NOx (mg N / L ) * [ A E R O B I C #1 O V E R F L O W (L/day) -I N T E R N A L R E C Y C L E F L O W (L/day)] A N O X I C #1 D E N I T R I F I C A T I O N R A T E (mg N/day) = A N O X I C #1 NOx L O A D (mg N/day) - A N O X I C #1 NOx (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day) A N O X I C #2 D E N I T R I F I C A T I O N R A T E (mg N/day) = A N O X I C #2 NOx L O A D (mg N/day) - A N O X I C #2 NOx (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day) A N O X I C #1 S P E C I F I C D E N I T R I F I C A T I O N R A T E (mg N/day / mg V S S ) = A N O X I C #1 D E N I T R I F I C A T I O N R A T E (mg N/day)/[ANOXIC # 1 V S S (mg/L) * 5 ( L ) ] A N O X I C #2 S P E C I F I C D E N I T R I F I C A T I O N R A T E (mg N/day / mg V S S ) = A N O X I C #2 D E N I T R I F I C A T I O N R A T E (mg N/day)/[ANOXIC #2 VSS (mg/L) * 5 ( L ) ] % A N O X I C #1 D E N I T R I F I C A T I O N = [ A N O X I C #1 D E N I T R I F I C A T I O N R A T E (mg N/day)/ANOXIC #1 NOx L O A D (mg N/day)] * 100 % A N O X I C #2 D E N I T R I F I C A T I O N =' [ A N O X I C #2 D E N I T R I F I C A T I O N R A T E (mg N/day)/ANOXIC #2 NOx L O A D (mg N/day)] * 100 A N O X I C #1 CH30H : NOx E N T E R I N G (mg COD/mg N) = A N O X I C #1 CH30H L O A D (g COD/day) * 1000 (mg/g)/ANOXIC #1 NOx L O A D (mg N/day) A N O X I C #1 CH30H : NOx R E M O V E D (mg COD/mg N) = A N O X I C #1 CH30H L O A D (g COD/day) * 1000 (mg/g)/ANOXIC #1 D E N I T R I F I C A T I O N R A T E (mg N/day) A N O X I C #2 CH30H : NOx E N T E R I N G (mg COD/mg N) = A N O X I C #2 CH30H L O A D (g COD/day) * 1000 (mg/g)/ANOXIC #2 NOx L O A D (mg N/day) A N O X I C #2 CH30H : NOx R E M O V E D (mg COD/mg N) = A N O X I C #2 CH30H L O A D (g COD/day) * 1000 (mg/g)/ANOXIC #2 D E N I T R I F I C A T I O N R A T E (mg N/day) A N O X I C #1 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day) = A N O X I C #1 E X P E C T E D T O T A L A M M O N I A (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day) + L E A C H A T E NOx (mg N / L ) * B A S E L E A C H A T E F L O W (L/day) + A E R O B I C #1 NOx (mg N / L ) * I N T E R N A L R E C Y C L E F L O W (L/day) + A E R O B I C #2 NOx (mg N / L ) * E X T E R N A L R E C Y C L E F L O W (L/day) - A N O X I C #1 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * A N O X I C # 1 O V E R F L O W (L/day) 173 A E R O B I C #1 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day) = A N O X I C #1 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day) - A E R O B I C #1 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * A E R O B I C #1 O V E R F L O W (L/day) A N O X I C #2 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day) = A E R O B I C #1 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * [ A E R O B I C #1 O V E R F L O W (L/day) - I N T E R N A L R E C Y C L E F L O W (L/day)] - A N O X I C #2 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day) A E R O B I C #2 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day) = A N O X I C #2 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * A N O X I C #2 O V E R F L O W (L/day) - A E R O B I C #2 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * A E R O B I C #2 O V E R F L O W (L/day) S Y S T E M T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day) = T O T A L A M M O N I A L O A D (g N/day) * 1000 (mg/g) + L E A C H A T E NOx (mg N / L ) * B A S E L E A C H A T E F L O W (L/day) - E F F L U E N T T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * S Y S T E M T O T A L F L O W (L/day) % A N O X I C #1 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L = { A N O X I C #1 T O T A L I N O R G A N I C N I T R O G E N ' R E M O V A L R A T E (mg N/day)/[ANOXIC #1 E X P E C T E D T O T A L A M M O N I A (mg N / L ) * A N O X I C #1 O V E R F L O W (L/day) + L E A C H A T E NOx (mg N / L ) * B A S E L E A C H A T E F L O W (L/day) + A E R O B I C #1 NOx (mg N / L ) * I N T E R N A L R E C Y C L E F L O W (L/day) + A E R O B I C #2 NOx (mg N / L ) * E X T E R N A L R E C Y C L E F L O W (L/day)]} * 100 % A E R O B I C #1 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L = { A E R O B I C #1 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N /day) / [ANOXIC #1 T O T A L I N O R G A N I C N I T R O G E N (mg N/L) * A N O X I C #1 O V E R F L O W (L/day)]} * 100 % A N O X I C #2 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L = { A N O X I C #2 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N / d a y ) / [ A E R O B I C #1 T O T A L I N O R G A N I C N I T R O G E N (mg N / L ) * [ A E R O B I C #1 O V E R F L O W (L/day) - I N T E R N A L R E C Y C L E F L O W (L/day)]]} * 100 % A E R O B I C #2 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L = { A E R O B I C #2 T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day)/[ANOXIC #2 T O T A L I N O R G A N I C N I T R O G E N (mg N/L) * A N O X I C #2 O V E R F L O W (L/day)]} * 100 % S Y S T E M T O T A L I N O R G A N I C N I T R O G E N R E M O V A L = { S Y S T E M T O T A L I N O R G A N I C N I T R O G E N R E M O V A L R A T E (mg N/day)/[TOTAL A M M O N I A L O A D (g N/day) * 1000 (mg/g) + L E A C H A T E NOx (mg N / L ) * B A S E L E A C H A T E F L O W (L/day)]} * 100 A N O X I C #1 E X P E C T E D C O D (mg/L) = [ L E A C H A T E COD (mg/L) * B A S E L E A C H A T E F L O W (L/day) + A E R O B I C #1 COD (mg/L) * I N T E R N A L R E C Y C L E F L O W (L/day) + A E R O B I C #2 COD (mg/L) * E X T E R N A L R E C Y C L E F L O W (L/day) + A N O X I C #1 CH30H L O A D (g COD/day) * 1000 (mg/g)]/ANOXIC #1 O V E R F L O W (L/day) A N O X I C #1 COD R E M O V A L R A T E (g/day) = { [ A N O X I C #1 E X P E C T E D COD (mg/L) - A N O X I C #1 COD (mg/L)] * A N O X I C #1 O V E R F L O W (L/day)} * l/iooo (g/mg) A E R O B I C #1 COD R E M O V A L R A T E (mg/day) = [ A N O X I C #1 COD (mg/L) * A N O X I C #1 O V E R F L O W (L/day) - A E R O B I C #1 COD (mg/L) * A E R O B I C #1 O V E R F L O W (L/day)] * 1/1000 (g/mg) 174 Explanatory Notes * In maintaining consistency with previous research (Azevedo, 1993; Guo, 1992; Shiskowski, 1995) for comparison and guidance, system internal and external recycle ratios were calculated by dividing recycle flows to simulated leachate flows (i.e., landfill leachate flow plus ammonium chloride flow) and not to total chemical addition flows to the first anoxic reactors (i.e., including phosphorus and methanol flows). * Reactors actual hydraulic retention times, expected and evaluated concentrations (i.e., when was the case), and removal values were calculated considering the total flow into the reactor of interest (i.e., including total chemical addition flow, where was the case), hence, considering the dilution factor contribution to the process. * Solids retention times were calculating considering only the actual first aerobic reactor daily wasting, hence, ignoring the withdrawal of about 250 ml of mixed liquor from each reactor on sampling days (i.e., about every second or third day). * Constituents concentrations within Aerobic #2 reactors were used in all calculations involving the external recycle lines of the systems. * In the following appendices (i.e., Appendix B, C, D, and E) negative ammonia removal rates and denitrification values signify accumulation of ammonia and NOx, respectively, within the reactor of interest, while negative nitrification values signify accumulation of NOx in the reactor preceding the reactor of interest. * In the following appendices (i.e., Appendix B, C, D, and E) the ammonia removal percentage, nitrification percentage, and denitrification percentage are calculated considering the actual constituent concentrations within the reactors of interest, that sometimes may be higher then the expected concentrations, mainly due to cell lysis, hence, percentage values higher than 100%. * It should be noted that, in the following appendices (i.e., Appendix B, C, D, and E): (a) ammonia removal rates, total inorganic removal rates, denitrification and nitrification are controlled by two principal variables while all calculated percentages are controlled by three principal variables; (b) long SSRTs (i.e., usually over 61 days) and "upset" process performance (i.e., bacterial inhibition, system failure) detennined cell lysis, hence, release of organic nitrogen eventually converted into inorganic nitrogen compounds, hence, increased nitrogen concentration within the reactor of interest, hence negative removal rates, denitrification and nitrification values, and negative and over 100% calculated percentages; (c) the results of the analyses of low concentration constituents were a function of instrument set detection limit (e.g., a measured value of 0.99 may represent an actual value of 0.00, when instrument detection limit is plus or minus 1.00), hence, calculated rates and percentages involving low concentrations are not always representative for process performance. * In the following appendices (i.e., Appendix B, C, D, and E) the symbol "d" was used for day or days; the symbol was used interchangeably with the terms. * The data in the following appendices (i.e., Appendix B, C, D, and E) were collected, as a rule, after the systems were under the same operating parameters for at least 24 hours, and the operating parameters of the systems were changed, usually, only after sampling. 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