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Changes in aerobic digester performance with the use of methanol for Biological Nutrient Removal (BNR)… Koh, Jeff Jae Hong 2001

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CHANGES IN AEROBIC DIGESTER PERFORMANCE WITH THE USE OF METHANOL FOR BIOLOGICAL NUTRIENT REMOVAL (BNR) IN A FULL-SCALE SEQUENCING BATCH REACTOR (SBR) by J E F F J A E H O N G K O H B.A.Sc. (Civil Engineering), University of British Columbia, 1997 A T H E S I S S U B M I T T E D I N P A R T I A L F U L F I L M E N T O F T H E R E Q U I R E M E N T S F O R T H E D E G R E E O F M A S T E R O F A P P L I E D S C I E N C E in T H E F A C U L T Y O F G R A D U A T E S T U D I E S D E P A R T M E N T O F C I V I L E N G I N E E R I N G We accept this thesis as conforming to the required standard THE UNIVERSITY OF BRITISH COLUMBIA March 2001 © Jeff Jae Hong Koh, 2001 U B C Special Collections - Thesis Authorisation Form Page 1 of 1 In presenting t h i s thesis i n p a r t i a l f u l f i l m e n t of the requirements for an advanced degree at the U n i v e r s i t y of B r i t i s h Columbia, I agree that the Lib r a r y s h a l l make i t f r e e l y a v a i l a b l e for reference and study. I further agree that permission for extensive copying of th i s thesis for s c h o l a r l y purposes may be granted by the head of my department or by his or her representatives. I t i s understood that copying or p u b l i c a t i o n of t h i s thesis for f i n a n c i a l gain s h a l l not be allowed without my written permission. Department of The U n i v e r s i t y of B r i t i s h Columbia Vancouver, Canada Date http://www.library.ubc.ca/spcoll/thesauth.html 3/2/01 ABSTRACT The primary goal of this research was to determine whether the use of methanol in the Sequencing Batch Reactors (SBR) in the District of Kent Wastewater Treatment Plant affected the aerobic digester performance. This was achieved by monitoring four key performance areas of sludge digestion: solids reduction performance; nutrient balance -influx, accumulation, and re-release; stability of digested sludge; and dewaterability of digested sludge. The secondary objective of this project was to investigate general performance issues surrounding the biosolids handling units of the District of Kent Wastewater Treatment Plant in response to the expected increase in incoming flow in the future. Two parallel process trains, Experimental and Control, consisted mainly of the two SBRs and the two aerobic digesters were established to provide means for direct comparison of the effect of methanol. Methanol was injected into the Experimental S B R at dosages of 27, 54 and 81 L/day. The first two dosages were run for one month each and the third for two months. In addition, an extra month of no methanol injection was studied. The methanol injection into the Experimental S B R resulted in a significant increase in the solids level in the SBRs , and the sludge wasting to the Experimental digester was increased accordingly by the plant operators. A s the methanol dosage increased from 27 L/day to 54 L/day and finally to 81 L/day, the sludge wasting escalated as well . A s the wasting kept increasing, so did the strain on the coarse bubble diffusers to supply sufficient oxygen; when the methanol dosage was at 81 L/day, the dissolved oxygen level in the digesters became inadequate for proper aerobic digestion. The resulting volatile suspended solids reduction was relatively low at 22 and 20% for Control and Experimental digesters, respectively. i i The total nitrogen influx into the Experimental digester followed that of the volume of S B R waste sludge; as the nitrogen content in the biomass of the two S B R waste sludges was very similar, except when methanol was first introduced and when it was discontinued. The nitrogen accumulation in the digesters was not a concern, as the total nitrogen present in the digesters was largely controllable by the adjusting the wasting of digester sludge. The increased solids, resulting in short sludge ages, and induced low dissolved oxygen levels, were found to limit hydrolysis of organic nitrogen into soluble forms, and therefore limited the release of ammonia back into the S B R , via the decant. The Experimental S B R waste sludges contained much more organic phosphorus in the biomass than the Control S B R waste sludge, but contained less orthophosphate in the bulk liquid. Overall, the total phosphorus influx was greater in the Experimental digester, due to the increased wasting. Accumulation of total phosphorus was also controlled largely by adjusting the digester sludge wasting. The increase in incoming total phosphorus in the Experimental digester, resulted in significant amounts of total phosphorus being re-released and re-entering the S B R via decant and filtrate. The amount of re-release was high enough to raise the incoming total phosphorus load into the S B R by up to 39%. The specific oxygen uptake rate (SOUR) and volatile suspended solids reduction was used to determine sludge stability. The specific oxygen uptake rate (SOUR) showed that it was affected by the sludge age of the samples. N o sludge sample taken during the course of the experiment met the required S O U R value of 1.5 mgO2 /1000mgVSShr. Wi th respect to the effect of methanol injection on the sludge stability, the methanol addition to the Experimental S B R (causing shorter sludge ages in the Experimental digester) was found to worsen the stability of the resulting digested sludge. Capillary suction time (CST) test was used exclusively to determine sludge dewaterability. The methanol injection into the Experimental S B R appeared to have no effect on the dewaterability of the digested sludge, as samples from both sides had very similar C S T results. It was found that the dewaterability was affected not only by the total suspended i i i solids concentration of the digested sludge, but even more by the temperature of the digested sludge, that controls the filtrate viscosity. Overall, there were no direct effects of the methanol addition to the sludge digestion process, but the increase in sludge wasting to the digester resulted in a detrimental affect on all parts of the digester performance that were monitored, except nitrogen re-release and sludge dewaterability. Filamentous microorganisms did cause sludge bulking in both digesters during the experiment, and Nocardia spp., was identified as the filamentous microorganism most common in the digester sludge. Although there are many factors that could cause the excessive growth of filamentous microorganisms, three possible reasons were identified for this plant: low dissolved oxygen in digester; long sludge age; and low food-to-microorganism ratio ( F / M ratio). iv TABLE OF CONTENTS Abstract i i List of Tables v i i i List of Figures ix Acknowledgements x i 1 Introduction 1 1.1 Significance 1 1.2 Objectives 3 2 Literature Review 5 2.1 Sequencing Batch Reactor 5 2.2 Biological Nutrient Removal 7 2.2.1 Biological Nitrogen Removal 7 2.2.2 Biological Phosphorus Removal 9 2.3 Methanol 10 2.4 Aerobic Sludge Digestion 12 2.4.1 Solids Reduction 13 2.4.2 Nutrient Re-release 14 2.5 Sludge Stabilization 17 2.6 Sludge Dewaterability 19 3 Methods and Materials 22 3.1 Experimental Design and Setup 22 3.1.1 Pre-Experiment Operations 22 3.1.2 Full-Scale Experiment 25 3.2 Sampling and Analysis 29 3.2.1 Sampling Plan 30 3.2.2 On-line Monitoring Plan 32 3.2.3 Analytical Methods 34 v 3.2.3.1 Total Suspended Solids (TSS) and Volatile Suspended Solids (VSS) 35 3.2.3.2 Ammonia (NH3), Nitrate and Nitrite (NOx), and Orthophosphate (P04) 36 3.2.3.3 Total Kjeldahl Nitrogen (TKN) and Total Phosphorus (TP) 37 3.2.3.4 Soluble Chemical Oxygen Demand (CODS) 38 3.2.3.5 Volatile Fatty Acids ( V F A s ) 39 3.2.3.6 Oxygen Uptake Rate (OUR) Test : 39 3.2.3.7 Capillary Suction Time (CST) Test 40 3.2.3.8 Microscopic Examination for Identifying Filamentous Microorganisms 41 3.2.4 Quality Assurance and Quality Control 42 3.2.5 District of Kent Wastewater Treatment Plant Data Collection Program 43 4 Results and Discussion 45 4.1 District of Kent Wastewater Treatment Plant 45 4.1.1 Challenges of a Full-Scale Experiment 45 4.1.2 Pre-Run Results: District of Kent Wastewater Treatment Plant Profile 47 4.2 Solids Production in S B R 51 4.3 Solids Reduction Performance 54 4.4 Nutrient Balance - Influx, Accumulation, and Re-Release 62 4.4.1 Nitrogen 62 4.4.2 Phosphorus 73 4.5 Sludge Stability 86 4.6 Sludge Dewaterability : 93 4.7 General Performance Issues 99 4.7.1 Sludge Bulking 100 4.7.2 Volatile Fatty Acids (VFAs) Production 107 4.7.3 Seasonal Variations .109 4.7.4 Scale-Up Issues 114 5 Conclusions 118 5.1 Summary and Conclusions 118 5.1.1 Effect of Methanol in S B R on Aerobic Digester Performance 118 5.1.2 General Performance Issues 121 v i 5.2 Recommendations for Further Research 123 6 References 124 Appendix A - l : Increase in S B R Wasting Due to Methanol Addition 129 Appendix A - 2 : Total Suspended Solids Reduction in Digesters 131 Appendix A - 3 : Volatile Suspended Solids Reduction in Digesters 137 Appendix A - 4 : Sludge Age of Digesters 143 Appendix A - 5 : Temperature and Dissolved Oxygen in Digesters 147 Appendix B - 1 : Nutrients in S B R Wasting 149 Appendix B-2: Nutrient Re-release via Digester Decant 155 Appendix B-3 : Nutrient Re-release via Filtrate 163 Appendix B-4: Nutrient Levels in Digesters 171 Appendix C: Specific Oxygen Uptake Rate of Digester Waste Sludge 177 Appendix D : Capillary Suction Time (CST) of Digester Waste Sludge 179 Appendix E : Other Relevant Data 180 Appendix F-1: Plant Data January 1999 183 Appendix F-2: Plant Data February 1999 186 Appendix F-3: Plant Data March 1999 189 Appendix F-4: Plant Data Apr i l 1999 192 Appendix F-5: Plant Data May 1999 195 Appendix F-6: Plant Data June 1999 198 Appendix F-7: Plant Data July 1999 201 Appendix F-8: Plant Data August 1999 204 v i i LIST OF TABLES Table 1: Operation System of the Control and Experimental S B R s 25 Table 2: Summary of Experimental Runs .29 Table 3: Summary of Sampling Plan for Experiment and Control 31 Table 4: Sample Locations and Sample Types 32 Table 5: On-line Monitored Parameters 33 Table 6: Sample Handling and Preservation 34 Table 7: District of Kent Wastewater Treatment Plant Laboratory Analyses 44 Table 8: Influent and Effluent Profiles (Jan. - Mar. 1999) : 48 Table 9: Profile of Septage After 7 day Aeration and Typical Septage 48 Table 10: Pre-Run S B R and Aerobic Digester Operating Characteristics 49 Table 11: Aerobic Digester Waste Sludge Profile 50 Table 12: Mass Balance Analysis 55 Table 13: Evidence of Denitrification in Experimental Digester Decant and Filtrate 71 Table 14: U S E P A Sludge Stabilization Options 87 Table 15: Suspended Solids Levels of Digester Decant 101 Table 16: Possible Causes of Sludge Bulking in Digesters 105 Table 17: Volatile Fatty A c i d Concentration in Digester Decant and Filtrate 108 Table 18: Kent Prison Influent Profile (Nov 98 - Jan 99) 115 v i i i LIST OF FIGURES Figure 1: Pre-Experiment Schematic of Treatment Plant 23 Figure 2: Modif ied Schematic of Treatment Plant For Experiment 27 Figure 3: Wasting of S B R Sludge 52 Figure 4: Increase in Solids Production in Experimental S B R 53 Figure 5: Volatile Suspended Solids Reduction in Digesters 56 Figure 6: Sludge Age of Digesters 57 Figure 7: Comparison of the Two Digesters with Theoretical Model 60 Figure 8: Average Dissolved Oxygen Levels and V S S Reduction 61 Figure 9: Total Nitrogen Entering Digesters Dai ly 64 Figure 10: Total Nitrogen Content in S B R Waste Sludge Biomass 65 Figure 11: Total Nitrogen and Total Nitrogen Concentration in Digesters 67 Figure 12: Organic Nitrogen and Soluble Nitrogen in Digesters 68 Figure 13: Total Nitrogen and Soluble Nitrogen Re-release from Digesters 70 Figure 14: Ammonia Re-release and Influent Ammonia Entering SBRs 72 Figure 15: Total Dai ly Phosphorus Wasted from S B R s 74 Figure 16: Organic Phosphorus Content in S B R Waste Sludge Biomass 75 Figure 17: Orthophosphate Concentration in S B R Waste Sludge 77 Figure 18: Total Phosphorus Mass and Total Phosphorus Concentration in Digesters 78 Figure 19: Average Daily Digester Sludge Wasting 79 Figure 20: Total Phosphorus and Orthophosphate in Digester Recycle Streams 81 Figure 21: Total Suspended Solids in Digester Recycle Streams 82 Figure 22: Increase in Phosphorus Re-release with Methanol Dosage 84 Figure 23: Orthophosphate Re-release and Orthophosphate in Influent Entering the S B R 85 Figure 24: Specific Oxygen Uptake Rate of Digester Waste Sludges @ 20 °C 89 Figure 25: S O U R and Sludge Age of Control Digester Waste Sludge 90 Figure 26: S O U R and Sludge Age for Experimental Digester Waste Sludge 91 Figure 27: Capillary Suction Time Results for Digester Waste Sludges 94 ix Figure 28: Effect of Total Suspended Solids in Sludge on Sludge Dewaterability 96 Figure 29: Effect of Temperature on Dewaterability of Control Digester Sludge 97 Figure 30: Effect of Temperature on Dewaterability of Experimental Digester Sludge 98 Figure 31: lOx Photograph of Control Digester Sludge (June 16, 1999 - Run 3) 102 Figure 32: 40x Photograph of Control Digester Sludge (June 16, 1999 - Run 3) 102 Figure 33: lOx Photograph of Experimental Digester Sludge (June 15, 1999 - Run 3).. 103 Figure 34: 40x Photograph of Experimental Digester Sludge (June 15, 1999 - Run 3).. 103 Figure 35: Average Daily Temperature of Digesters 110 Figure 36: Average Weekly p H and Sludge Age of Digesters I l l Figure 37: Seasonal Influent Flow Variation 113 x ACKNOWLEDGEMENTS I would like thank the Department Civil Engineering at the University of British Columbia, for giving me the opportunity to continue learning and to complete this program. I would like to express my sincere gratitude to my supervisors, Dr. D.S. Mavinic, Dr. W.K. Oldham, and Dr. A. Meisen, for their invaluable advice, constructive criticism, and constant support throughout the course of my research. It has truly been a rewarding experience working with all my supervisors. I wish to thank Nuno Louzeiro, whom I braved the hot summer and cold winter of the Fraser Valley together, along with our "cop car", the Chevy Caprice. I am very thankful for his assistance, cooperation, and for the good times. I would also like to thank all the people who were a part of this research: Ian Gardner, Steve Nuttall, and Keith Paisley, from the District of Kent, B.C., for allowing the research to take place at their facility, for their assistance, valuable advice, and their friendship; Paula Parkinson, Susan Harper, and Scott Jackson, from the Department of Civil Engineering, for their assitance and advice. I would like to thank my parents, and my brother Jason, for their constant love, encourage-ment, and ceaseless prayers. I am also thankful to Pastor David Oh and the rest of my brothers and sisters at Philadelphia church for their support of prayers. Above all, I thank Lord Jesus for his unconditional love, sacrifice, and grace. By his grace alone I am what I am. I dedicate this work to him. xi 1 INTRODUCTION 1.1 Significance In that past 10 years, environmental issues have received much publicity and public interest around the world, as the environment throughout the world has continued to deteriorate with increased human activity. Now, with the public outcry for conservation of the environment being as loud as ever, governments have responded with comprehensive recycling programs, gas emissions control policies, and other policies in attempt to protect the land, air and waters. Since nitrogen and phosphorus are limiting nutrients in biomass growth, they have been found to accelerate the eutrophication of lakes and reservoirs. Not only is the increased aquatic vegetation and algae aesthetically unsightly, they may interfere with the beneficial uses of water resources. Furthermore, a high level of nitrogen is known to have other adverse effects, such as ammonia and nitrite toxicity toward aquatic life forms (including fish), depletion of dissolved oxygen in receiving waters, and nitrate contamination in ground water supplies (US E P A , 1976). Currently, the Canadian Water Guidelines suggest the total ammonia concentration in a wastewater discharge effluent to be no higher than 1.81 mg N / L , and nitrate at 0.02 mg N / L (Ministry of Supply and Services, 1995). Wi th the public becoming increasingly sensitive about environmental issues, the discharge regulations put forth by the government are expected to become more restrictive in the future, resulting in added pressure for many wastewater treatment plants to rethink their strategy to treat the incoming wastewater. To accommodate the current and future regulations, many treatment plants are adapting their wastewater treatment methods to Biological Nutrient Removal (BNR) processes. B N R 1 processes are now the choice of many engineers over chemical addition and other nutrient removal techniques when upgrading or designing new wastewater treatment plants. Sequencing Batch Reactors (SBRs), a type of bioreactor gaining popularity due to its relatively smaller land requirements, can provide the necessary anaerobic, anoxic, and aerobic zones required for B N R . Methanol has been widely accepted as an economical and effective external carbon source for denitrification. However, The use of methanol for B N R processes has only had limited research. Likewise, insufficient research has been done to profile the effect of methanol in the rest of the wastewater treatment facility. The major by-product of wastewater treatment is sludge or biosolids. The waste sludge from the bioreactors is most commonly sent to either aerobic or anaerobic sludge digesters for further treatment. The primary role of the digesters is to reduce the mass of solids and to stabilize the incoming sludge. Sludge digestion is a key part of wastewater treatment, with sludge management now representing 40 to 50 percent of total plant costs (Mavinic, 1998). Since the consequences of poor performance would normally appear as a poor quality final product or i f the decant supernatant or filtrate is recycled, it would affect the rest of the treatment units. A good performing sludge digestion operation should provide high (40-50 %) suspended solids reduction performance, high biological stability, and good dewaterability. Aerobic digesters are often preferred to anaerobic digesters for plants that are designed to handle less than 5 Mgal /d (1.7 x 10 4 m 3/day) mainly because of lower capital cost. Moreover, aerobically digested sludge is more odorless and biologically stable than anaerobically digested sludge (Metcalf and Eddy, 1991). Handling and disposal of surplus sludge is an expensive process in wastewater treatment. Therefore, sludge digestion and the subsequent dewatering process are a very important part of the process train. Continuing efforts to improve the performance of the liquid treatment operation in B N R processes is undoubtedly important. However, this emphasis 2 on just the liquid treatment has often led to insufficient focus on the biosolids handling unit processes (Kempton et al., 1999). This research reported on here was conducted at the District of Kent Wastewater Plant in Agassiz, British Columbia, Canada. This treatment plant currently receives combined sewage of municipal wastewater from a community of 5000 people plus local septage, and stormwater. In the near future, the District of Kent Wastewater Treatment Plant is expected to receive the wastewater from the Kent Max imum Security Prison and a nearby First Nations Reservation. A s a result, the wastewater inflow is expected to increase significantly, thereby increasing the strain on the treatment facilities. Moreover, traditionally prison wastes are stronger than municipal waste, due to the reduced water consumption per capita in prisons. 1.2 Objectives The primary objective of this research was to determine whether the use of methanol, for denitrification purposes in the Sequencing Batch Reactors (SBRs), affects the aerobic digester performance itself. Four major areas were monitored to profile the effect of methanol on the aerobic digesters: 1) Solids reduction performance; 2) Nutrient Balance - Influx, accumulation, and re-release; 3) Stability of digested sludge; and 4) Dewaterability of digested sludge. The secondary objective of this study was to investigate general performance issues surrounding the biosolids handling units of the District of Kent Wastewater Treatment Plant, in response to the future increase in incoming flow. 3 This research is the biosolids management component of an on-going study at the University of British Columbia to explore the potential of methanol for Biological Nutrient Removal (BNR) purposes. Therefore, the scope of this research is limited to the biosolids treatment and management aspect of wastewater treatment. 4 2 LITERATURE REVIEW This chapter summarizes the findings of studies that were previously done and available in literature that are relevant to this research. Although this study is a part of a joint project to examine the use of methanol for biological nutrient removal in sequencing batch reactors, the scope of this research is limited to the biosolids management aspect of wastewater treatment. Therefore, this chapter w i l l emphasize more on topics related to sludge management and aerobic digester performance. 2.1 Sequencing Batch Reactor The Sequencing Batch Reactor (SBR) is a unique secondary treatment system; it is a f i l l -and-draw type reactor system, involving a single, complete-mix reactor in which all steps of the activated sludge process occur. It can also be described as a time-oriented, periodic system, which can satisfy different treatment objectives by simply modifying the application and duration of mixing and aeration in a single tank (Metcalf and Eddy, 1991). Since the 1970's, the S B R technology has improved drastically with the advent of improved aeration devices and computers, along with constant research from the environmental engineering research community. Currently the S B R technology is being adopted rapidly, worldwide, for activated sludge treatment at small and medium-sized municipal wastewater treatment plants (Rusten et ah, 1993). The main components of a S B R are the main tank, inlet, outlet, mixing and aeration system, and controller. A typical operating sequence for a S B R is composed of the following steps: f i l l , react (aeration), settle (sedimentation/clarification), draw (decant), and idle (with solids wasting). M i x e d liquor remains in the reactor during all cycles, thereby eliminating the need for separate secondary sedimentation tanks. Average design 5 hydraulic retention times range from 12 hours, where the objective is to meet organic and suspended solids reduction, to 24 hours, when flow rates are highly variable and nitrification, denitrification, and phosphorus removal is also required (US E P A , 1986). The S B R has some distinct advantages over traditional activated sludge systems, which include the following (Ketchum, 1997, and Louzeiro, 1999): • A n increased flow rate can be treated for a given tank volume, since only one tank is required to accomplish flow equalization, biological reactions, and biomass settling (clarification); • SBRs have smaller land requirements and are more economical, due to the elimination of a separate clarification tank; • S B R tank serves as an equalization basin during the f i l l phase, and therefore is able to handle peak flows and shock loadings, without a considerable drop in effluent quality and without the need to extra tanks; • S B R does not require sludge recycling (RAS) between tanks and therefore reduces the number of pumps and extra piping required; and • The clarification process in an S B R is usually more efficient than in continuous-flow systems. Most of the disadvantages associated with S B R systems stem from the fact that the technology is new, compared to conventional activated-sludge systems. Lack of research has made it very difficult to predict the behavior of a given system (Louzeiro, 1999). Furthermore, as the S B R system gets larger, there is an increasing degree of sophistication associated with the timing units and level sensors required to control the process parameters. A s a result, the S B R is primarily suitable for smaller flow rates (US E P A , 1986). 6 2.2 B i o l o g i c a l N u t r i e n t R e m o v a l With the increase in scientific knowledge in wastewater treatment, it became increasingly apparent that secondary treatment, to reduce the biological oxygen demand in wastewater was not enough. Since the early 1970's, the number of wastewater treatment facilities capable of advanced wastewater treatment have grown remarkably (Metcalf and Eddy, 1991); thus only encouraged and demanded more scientific research into advanced wastewater treatment. The removal of nutrients has been gathering more attention compared to other advanced treatment areas in wastewater treatment. Nitrogen and Phosphorus are limiting nutrients in biomass growth, and they have been found to accelerate the eutrophication of lakes and reservoirs and stimulate the algal growth. Not only are the increased aquatic vegetation and algae aesthetically unsightly, they may interfere with the beneficial uses of water resources. Furthermore, a high level of nitrogen is known to have other adverse effects such as ammonia and nitrite toxicity toward aquatic life forms (including fish), depletion of dissolved oxygen in receiving waters, and nitrates concerns in public health, because of its serious and occasionally fatal effects on infants (US E P A , 1976). A strict discharge guideline exists in North America and in many parts around the world. Numerous processes have been developed to separate the nutrients from the discharge effluent. They can be largely classified into two groups; chemical addition and biological nutrient removal. Biological nutrient removal, or B N R , has been gaining wide popularity among design engineers over chemical addition, mostly because of the cheaper cost and the production of non-chemical, stable end products. 2.2.1 Biological Nitrogen Removal Most nitrogen found in surface waters is derived from land drainage (3-24 lb of N/acre/year, 3.5-27 kg of N/ha/year) and dilution of wastewater effluents. Feces, urine, and food processing discharges are the primary sources of nitrogen in domestic waste with 7 a per capita contribution in the range of 8-12 lb of N/year (3.5-5.5 kg of N/year) (Viessman et al., 1993). The nitrogen forms of interest are organic, inorganic and gaseous nitrogen. To summarize briefly, hydrolysis and bacterial decomposition releases ammonia by deamination of nitrogenous organic compounds, Organic N ^ NFL,"1" (Bacterial Decomposition) Continued aerobic oxidation results in nitrification, N H 4 + + 0 2 N 0 2 " (Nitrifying Bacteria) N 0 2 " N 0 3 " (Nitrifying Bacteria) Denitrification occurs with heterotrophic metabolism in an anaerobic or anoxic environment, N 0 3 ~ N 2 (g) (Bacterial Denitrification) These three reactions, in sequence, define the biological nitrification-denitrification process. In biological nitrification, under aerobic conditions, ammonia in the wastewater is transformed to nitrite (N0 2 ") and then to nitrate (N0 3~) by nitrifying bacteria Nitrosonomas and Nitrobacter, respectively. The equations of the reactions are as follows (Metcalf and Eddy, 1991): <Nitrosonomas> 55 N H 4 + +760 2 + 109HCO 3 " C 5 H 4 0 2 N + 54N0 2 " + 5 7 H 2 0 + 1 0 4 H 2 C O 3 <Nitrobacter> 400NO 2 " + N H 4 + + 4 H 2 C 0 3 + H C 0 3 " + 1950 2 C 5 H 4 0 2 N + 3 H 2 0 + 400NO 3~ According to the above equations, a large amount of alkalinity is consumed and therefore extensive nitrification can drop the p H in wastewater. A p H range of 7.5 to 8.6 exists for 8 an optimal reaction; therefore, abundant alkalinity may be required in the system to prevent the decrease of p H outside this range (Metcalf and Eddy, 1991). Biological denitrification, which is the second step that follows nitrification in the biological nitrogen removal process, is completed under anaerobic or anoxic conditions. The microorganisms that are responsible for denitrification are heterotrophic bacteria Achromobacter, Aerobactor, Alcaligenes, Bacillus, Brevibacterium, Flavobacterium, Lactobacillus, Micrococcus, Proteus, Pseudomonas, and Spirillum. Denitrification reactions require an organic carbon source. In wastewater treatment, denitrification follows nitrification so the organic carbon that is available is usually lysed cell mass during endogenous decay. The use of endogenous carbon may result in low denitrification rates and therefore, limited overall nitrogen removal (Leung et al., 1994). Denitrification requires the use of hydrogen ions and restores some alkalinity in the process, thus compensating for the alkalinity lost during nitrification. The denitrifiers appear to operate at an optimal p H range of 7 to 8 (Metcalf and Eddy, 1991). 2.2.2 Biological Phosphorus Removal Most phosphorus entering surface waters is from human-generated wastes and land runoff. Contributions from non-point sources in surface drainage vary from 0 to 15 lb of P/acre/year (0-17 kg of P/ha/year), depending on land use, agricultural practice, fertilizer additions, topography, soil conservation practices, and other factors. Domestic waste contains approximately 3.5 lb of P/capita/year (1.6 kg of P/capita/year), of which about 60% is from phosphate builders used in synthetic detergents (Viessman et al., 1993). The most common forms of phosphorus are organic phosphorus, orthophosphates (FLPCV, HPO4 2 ", PO43"), and polyphosphates. A l l polyphosphates in aqueous solution gradually hydrolyze to orthophosphates. Phosphorus is utilized by microorganisms during cell 9 synthesis, and orthophosphate is available for biological metabolism without any further breakdown. Acinetobactor are one of the most prominent phosphorus accumulating organisms (PAOs) responsible for phosphorus removal. These microorganisms release the stored phosphorus in the presence of volatile fatty acids ( V F A s ) under anaerobic conditions. If the anaerobic zone is followed by an aerobic zone, microorganisms wi l l uptake the phosphorus above usual normal levels. Then, the P A O s are either wasted via sludge wasting, or removed and treated in a side stream, to release the excess phosphorus. 2.3 Methanol Methanol ( C H 3 O H ) , as the chemical formula suggests, is a single-carbon-chain compound. It is usually made from natural gas, and flammable, volatile, and extremely toxic. Being a single-carbon-chain compound, which is the simplest form of carbon compound, methanol can be readily consumed by microorganisms that require a carbon source during any biological reaction. For this reason, studies have shown that methanol is an effective external carbon source required for denitrification. For example, Tam et al. (1994) and Carley et al. (1991), in their respective studies demonstrated that after acetate and propionate, methanol was most effective in producing improved denitrification results. The other benefit of methanol is its relatively cheap price, compared to the other substrates; that makes methanol an attractive option for denitrification purposes. Generally, it is understood that the V F A s , which are required for biological phosphorus removal (BPR), can only be the substrate that the P A O s wi l l use, to release the phosphorus during primary release of phosphorus in the anaerobic basin. Other forms of substrate require fermentation to be converted to V F A s , before they can be utilized. However, findings by Carucci state that organic substrates, other than V F A s , can be used as 10 substrates for B P R without being converted to V F A s (Carucci et al., 1999). The effectiveness of methanol as a substitute for V F A s has been studied with mixed results. Tarn et al. (1992) and Randall et al. (1997), found in their respective experiments that methanol addition had a negligible effect on improving phosphorus removal. On the other hand, Jones et al. (1987) found evidence to conclude that methanol supported B P R . Overall, there has not been enough research done to profile the effectiveness of methanol for phosphorus removal, especially under full-scale plant operation. The expected solids production in the bioreactors, when methanol is injected, can be estimated in two ways, the theoretical model and the kinetic model. The theoretical model, based on the following nitrate removal reaction, is: N 0 3 " + 1.08 C H 3 O H + FT 0.065 C 5 H 7 0 2 N + 0.47 N 2 + 0.76 C 0 2 + 2.44 H 2 0 where C H 3 O H is the methanol added and C 5 H 7 0 2 N is the biomass or sludge. Using this model, Louzeiro (1999) reported that approximately 0.17 kg of solids were produced per L of methanol. The kinetic model was based on the following formula: dt ( d L h ^ a b • S v d t J •V where: V = volume, m 3 ; S = volatile solids concentration, mg V S S / L ; L = substrate concentration, mg C O D / L ; t = time, day; a = yield coefficient, mg V S S / mg C O D ; and, b = endogenous decay coefficient, day"1. This kinetic model can be simplified, when introducing a net yield coefficient, a', to represent the yield and endogenous decay, to the following: dt f ,dh\ a — V d t J •V Louzeiro, set up a full-scale experiment at the District of Kent Wastewater Treatment Plant, in Agassiz, B . C . , to inject methanol into a full-scale sequencing batch reactor (SBR) 11 and using the historical data of the S B R , found that 0.44 kg of bio-solids were produced per L of methanol (Louzeiro, 1999). 2.4 Aerobic Sludge Digestion Sludge digestion entails the microbiological and biochemical transformation of organic biomass in wastewater sludges to harmless and safe end products. The primary objective of sludge digestion is solids reduction. Currently, the two most widely used sludge digestion methods are anaerobic digestion and aerobic digestion. Aerobic sludge digestion has several clear advantages and disadvantages. The advantages are summarized as follows (Lue-Hing etal, 1998 and Mavinic , 1998): • low capital cost; • ease of operation; • volatile solids reduction is approximately equal to that obtained anaerobically; • a biologically stable end product is produced; • end product is odour free; • lower B O D in supernatant decant; • recovery of more basic fertilizer values in sludge; and • high pathogen reductions. The disadvantages are summarized as follows: • high power requirements for mixing and aeration; • only suitable for small plants with less than 5 M G D or 220 Us; • volatile solids reduction varies with temperature; • loss of methane production; and • sludge produced has poor mechanical dewatering characteristics. 12 2.4.1 Solids Reduction The aerobic digestion process involves inducing endogenous respiration of the biomass, after all the soluble substrate is completely utilized by the microorganisms. The aerobic digestion of the waste activated sludge may be considered as a continuation of the activated sludge process. The phenomenon of endogenous respiration can be described by the following equation (Lue-Hing et ai, 1998): C5H7O2N + 8.250 2 5 C 0 2 + H + + NO3 + H3PO4 + 1.5H 2 0 + inert residuals Metcalf and Eddy present a slightly different more simplified equation to describe endogenous respiration (Metcalf and Eddy, 1991): C 5 H 7 0 2 N + 7 0 2 5 C 0 2 + H + + N 0 3 " + 3 H 2 0 + inert residuals The cell tissue is oxidized aerobically to carbon dioxide (CO2), water ( H 2 0 ) , and ammonia (NH3). The ammonia is subsequently oxidized further to nitrate (NO3) as digestion ensues. However, not all of the cell tissue is biodegradable; only 75 to 80 percent can be oxidized. The remaining 20 to 25 percent is composed of inert compounds and organic compounds that are not biodegradable. The kinetics of endogenous respiration can be represented in a first order reaction: rxd = -k d * X v where, r X d is the rate of endogenous decay (mg/L-d); kd is the endogenous rate coefficient (d"1); and X v is the volatile solids concentration (mg/L). Reports of kd for full-scale aerobic digestion systems are limited and vary with great magnitude. This is probably because kd is a sludge characteristic and each wastewater treatment facility would have its own unique kd. Reported values of k^ for bench scale 13 systems are readily available compared to the full-scale systems. Krishnamoorthy summarized several studies to report a k d range 0.016 to 0.426 d"1 (Krishnamoorthy et al, 1989). Koers studied the kd in cold temperatures (5, 10, 20 °C) in both bench scale and full scale studies and reported a range of 0.0049 to 0.0402 d"1 (Koers, 1979). The successful application of aerobic digestion of wastewater sludges depends on the type of sludge and the temperature. While these two are not controllable, the primary controlling design parameter in aerobic digestion is the detention time. Typical detention times for aerobic digesters range from 15 to 30 days (Lue-Hing et al, 1998). Given normal conditions, an aerobic sludge digester should be able to achieve 40 to 50 % reduction of volatile solids. The theoretical quantity of oxygen required to oxidize the biomass (C5H7O2N) and completely oxidize the ammonia released to nitrates is 1.98 pounds of oxygen per pound of biomass (0.9 kg of oxygen per kg of biomass). However, the results from full-scale and bench-scale studies indicate the oxygen requirements ranging from 1.74 to 2.07 pounds of oxygen per pound of biomass (0.79 to 0.94 kg of oxygen per pound of biomass) (US E P A , 1979). The actual specific oxygen utilization rate, mass of oxygen per mass of volatile solids per hour, is reported to be a function of total sludge age and liquid temperature. The oxygen residual or the dissolved oxygen level should be maintained at 1 mg/L or above, under all operating conditions (Ahlberg et al, 1972); otherwise, the biological activity of the bacteria is hindered, and therefore, the digester w i l l experience poor solids reduction. 2.4.2 Nutrient Re-release Bioreactors achieving biological nutrient removal (BNR) are expected to produce waste sludge that is high in nutrient content (especially phosphorus), compared to waste sludge from conventional bioreactors. Phosphorus, more so than nitrogen, is expected to be present in higher concentrations in the waste sludge since the biological phosphorus 14 removal mechanism involves removing the phosphorus accumulating organisms (PAOs), following the aerobic phosphorus uptake stage via the waste sludge. The majority of the phosphorus uptake is stored in the biomass in long chains of polyphosphate. The polyphosphate is unstable and maybe released into the liquid phase under anaerobic conditions or under prolonged periods of aeration (Barnard et al., 1998). Tonkovic (1999) summarized various studies of phosphorus re-release and determined that intermittent aeration reduces phosphorus re-release by 50%, compared to that of continuous aeration. In addition, biosolids from B N R treatment plants, operating under long sludge ages (e.g. 40 days), were found to leach very little phosphorus even under continuous aeration conditions. It was suggested that the phosphorus was in a mineralized form contained within the lysed cell wall or as inorganic granules (polymeric metal phosphate complexes) within the cell proper (Tonkovic, 1999). Anaerobic digestion, compared to aerobic digestion has been found to solubilize up to 60% of the total phosphorus in the B N R waste sludge (Murakami et al., 1987). Zhao et al. (1998), from their bench scale two-stage intermittent aeration biological removal process, observed phosphorus release in the presence of nitrate and dissolved oxygen, during the air-off period. This mode of release of phosphorus, called the secondary release of phosphorus, has a detrimental effect on the efficiency of the overall phosphorus removal process. Zhao et al. suggested that a relatively long aeration cycle (with long air-off period) should be avoided in the two-stage intermittent aeration process. According to Barnard et al. (1998), the secondary release of phosphorus may result from a long hydraulic retention time in the aerobic zone, inadequate supply of organic compounds in the wastewater, and insufficient activity of the fermentation bacteria in the anaerobic zone. Numerous studies conducted by the researchers at the University of British Columbia proved that the monitoring of oxidation-reduction potential (ORP) in the reactors is beneficial to optimizing a B N R process, because it enables for making the distinction 15 between anoxic, anaerobic, and aerobic periods (Louzeiro, 1999). Wi th the improved control of aeration input in the reactors, it may be an effective means to control the secondary release of phosphorus (Koch et al., 1985). Nitrogen removal is mostly achieved through the denitrification process, which transforms the nitrate to nitrogen gas and releases it into the atmosphere. Therefore, depending on the extent of denitrification achieved in the bioreactor, the amount of nitrates present in the waste sludge wi l l vary. However, considering that the majority of the nitrogen wi l l be in the form of organic nitrogen contained in the cell mass, because of organic carbon removal, the nitrates wi l l only be of minor significance (Kempton et al., 1999). During aerobic digestion, the organic nitrogen in the biomass is hydrolyzed to ammonia during cell lysis and then is nitrified to form nitrate. This is much more prevalent in continuous aeration compared to intermittent aeration. Under intermittent aeration, organic nitrogen is hydrolyzed to ammonia and then nitrified to nitrates as well . However, under anoxic conditions, when the aerators are inactivated, the nitrates are denitrified to nitrogen gas. This brings about significantly lower nitrates levels in the supernatant decant and filtrate compared to that from a continuously aerated digester. However, denitrifiers require specific conditions that need to be satisfied before the denitrification process is induced. The optimal p H for denitrifiers exists between 7 and 8 (Metcalf and Eddy, 1991). The presence of dissolved oxygen wi l l suppress the enzyme system needed for denitrification. Several studies have indicated that a critical dissolved oxygen concentration of approximately 0.2 mg/L exists, above which denitrification is minimal. It is also suggested that the denitrification activity is also adversely affected by extremely low dissolved oxygen concentrations (Lie et al., 1994). If these conditions are not met little or no denitrification may occur which, in sludge digestion where decant is usually recycled back into the influent stream, could result in more nitrate released into the wastewater flow. This increases the possibility of increased presence of nitrate in the effluent. 16 2.5 Sludge Stabil ization Currently, several methods of ultimate municipal sludge utilization or disposal are currently practiced. These include land application, distribution and marketing, landfilling, incineration, and ocean disposal. In all of these except incineration, there is some slight risk to humans from exposure to the pathogenic microorganisms that these sludges may contain. Government regulations in many countries now require specific reduction in pathogens, before they are applied to land. Reduction in vector attraction of municipal wastewater sludges is a relatively new concept that originated with the U S E P A ' s 1979 regulation for sludge disposal. In 1979, the US E P A acknowledged that vectors, which are attracted to the sludge, are capable of transporting infectious agents and therefore, poses as a possible threat to public health safety. Moreover, this was reflected by the inclusion of vector attraction reduction in the latest U S E P A regulations for the management and disposal of sewage sludge, 40 C F R Part 503 - Standards for the use or disposal of sewage sludge (US E P A , 1993). Both vector attraction reduction and pathogen reduction in sludge can be achieved in large part, by stabilizing the sludge. In addition, the elimination of offensive odours and the inhibition of the potential for putrefaction can be achieved via sludge stabilization. The success in achieving these objectives is related to the effects of the stabilization operation or process on the volatile or organic fraction of the sludge. The means to eliminate these potential health hazards and nuisance conditions, through stabilization, are the biological reduction of volatile content, the chemical oxidation of matter, the addition of chemical to the sludge to render it unsuitable for the survival of microorganisms, and the application of heat to disinfect of sterilize the sludge (Metcalf and Eddy, 1991). Aerobic digestion can achieve sludge stabilization since the volatile content is destroyed via endogenous decay in the aerobic digestion process. The primary measure of degree of stabilization by sludge digestion is percent volatile solids reduction. The 1979 U S E P A 17 regulations recommend a minimum 38 percent reduction in volatile solids as achievable and indicative of stabilization for both anaerobic and aerobic digestion. However, application of the 1979 regulation in ensuing years showed that, sometimes, sludges were stable and did not attract vectors, but the requirement for 38% volatile solids reduction could not be met. (Farrell et al., 1996) The use of specific oxygen uptake rate (SOUR) has been recommended by E ikum et al. (1977) as a reliable method for indicating sludge stability provided temperature effects are taken into consideration. The oxygen uptake rate (OUR) can be used as a measure of biological activity since microorganisms in sludge use oxygen as they consume food. The combination of the O U R data with the concentration of the volatile suspended solids yields the S O U R . Jeris et al. (1985), E ikum et al. (1977), and Koers (1979) monitored S O U R with progressing sludge digestion, and these authors observed that S O U R typically increases briefly in the early stages of digestion and then gradually decreases as the duration of the digestion period increases. Moreover, It was found that the sludge did not putrefy or cause offensive odours when the S O U R was less than 1.2 mg0 2 /h r /gVSS. The use of S O U R as a measure of sludge stability is fully implemented in the 1993 U S E P A regulations for sludge disposal. According to the U S E P A , aerobically digested sludge is considered stable i f the volatile solids reduction is greater than 38 percent or the S O U R of the sludge is less than 1.5 mg0 2 /hr /gVSS at 20 °C (US E P A , 1993). Ahlberg et al. (1972) examined aerobically digested sludges at a number of wastewater plants and obtained a correlation with length of time of aeration and S O U R . For digesters operated at temperatures over 10 °C, S O U R fell to -2.0 mg0 2 /hr /gVSS after a total sludge age of 60 days and to 1.0 mg0 2 /hr /gVSS after -120 days sludge age. Aerobic digestion is usually carried out at ambient temperatures that can range from 5 to 30 °C. Consequently, i f S O U R s are used to indicate stability, there must be a way to adjust the S O U R measurement to a standard condition. E ikum et al. (1977) recommended that 18 the following equation to be used to adjust the S O U R determined at one temperature to another, based on the Arrhenius equation (Farrell et al, 1996): S O U R , T\ _ QT\-T2 S O U R T 2 Where, S O U R T i = Specific oxygen uptake rate at T i ; S O U R T 2 = Specific oxygen uptake rate at T 2 ; and 9 = Temperature sensitivity coefficient. Available literature indicate that the temperature sensitivity coefficient of volatile solids destruction depends to some extent on the type of sludge being digested. Koers (1979) found the value of 0 to be less than 1.072 at temperatures greater than 15 °C, whereas Eikum et al (1977) determined 0 to equal 1.112 for primary sludges. Grady et al. (1980) have reviewed the data of several investigators and recommended that 0 be determined experimentally. If this is not possible, they recommend that 0 = 1.05 be used for digestion of waste activated sludges. 2.6 Sludge Dewaterabil i ty Sludge dewatering is a physical/mechanical unit operation used to reduce the moisture content of sludge. Some of the reasons cited for the advantages of sludge dewatering are as follows (Metcalf and Eddy, 1991): • the costs for trucking sludge to the ultimate disposal site become substantially lower when sludge volume is reduced by dewatering; • dewatered sludge is generally easier to handle than thickened or l iquid sludge; • dewatering is required normally prior to the incineration of the sludge to increase the energy content by removal of excess moisture; 19 • dewatering is required before composting to reduce the requirements for supplemental bulking agents or amendments; • in some cases, removal of the excess moisture may be required to render the sludge odorless and non-putrescible; and, • dewatering is required prior to landfilling to reduce leachate production at the landfill site. Sludge dewaterability describes how easily the water can be separated from the solids content in the sludge. Naturally, the sludge dewaterability has a direct effect on the performance of the sludge dewatering operation. The dewaterability of the sludge depends on many factors listed as follows (Lotito et al, 1993): • fluid properties such as bound water content, viscosity, ionic strength, density; • particle properties such as size distribution, shape, surface potential, surface area, density; and, • sludge properties such as solids concentration, permeability, yield strength, and electrokinetic properties. Amongst the various factors that influence sludge dewaterability, particle size distribution is considered as the most significant. Particle size distribution describes the distribution of particles in terms of size. Fine particles, when using a centrifuge or a gravitational thickener to dewater sludge, descend very slowly because of their small size. The amount of fine particles also affects the amount of adhered water. The smaller a particle is, the bigger the specific surface, and, depending on its hydrophobicity, more water is bound per volume fraction. Therefore, the increase of fine particles negatively affects the sludge dewaterability. However, the inner floe structure must be taken into account, to fully understand the dewatering of sewage sludge (Olboter et al, 1993). There are several ways to quantify sludge dewaterability, but there seems to be no universally accepted means to evaluate the ease with which sludge wi l l release its water. 20 The specific resistance to filtration (SRF) test was the first widely used technique. This test is based on an analysis of pressure drop for flow through a porous medium using the Darcy equation. The resulting sludge characterization parameter is related to permeability. It is understood that the resistance to filtration originates from the hydraulic resistance of water flowing in the pores of the filter cake. Hence, the particle size distribution of the slurry seems a most important parameter (Agerbaek et al, 1993). This technique, although based on sound theory, has not gained great popularity, however, due to its time consuming and cumbersome nature. In addition, it has been suggested that the S R F is a poor test for estimating actual vacuum or pressure filter performance (Vesilind, 1988). The capillary suction time (CST) test is a quick and easy method for determining sludge dewaterability. This test has been found to be especially useful for profiling the effect of chemical conditioners on sludge dewaterability. Unfortunately, the C S T test is a purely empirical test and is not based on a theoretical analysis of sludge dewaterability. For this reason, many researchers have denounced the C S T test as a research tool only. However, research carried out by Unno et al. (1983) and Vesi l ind (1988) have shown that a mathematical relationship between the S R F and C S T exist, when the C S T is considered with other variables such as solids concentration and fluid viscosity; this is then reported as the filterability constant, therefore validating the use of C S T as a tool for sludge dewaterability. 21 3 METHODS AND MATER IALS The change in aerobic digester performance with methanol addition for biological nutrient removal in a full-scale sequencing batch reactor (SBR) was examined in this research. This full-scale demonstrative experiment included four experimental runs. This section describes the materials, procedures, and analytical methods that were employed in this study. It should be noted that the scope of this research is limited to the sludge management aspect of wastewater treatment. Therefore, the details of the full-scale experiment design involving the SBRs and the methanol injection, which were a part of a separate study by Louzeiro (1999) in this joint project, are only briefly discussed in this section. 3.1 Experimental Design and Setup 3.1.1 Pre-Experiment Operations This research was conducted at the District of Kent Wastewater Treatment Plant in Agassiz, British Columbia, Canada from January 1999 to August 1999. This plant has recently undergone a major retrofit in 1993, under the supervision of Stantec Consultants. The facility has a design flow of 5400 m 3/day, and the average flow was 1100 m 3/day in 1998 and 1200 m 3/day in 1999. This wastewater treatment plant primarily treats the municipal wastewater generated by nearly 5000 people in Agassiz, British Columbia. The treatment plant's flow sequence includes: rotary screens, an air-degritter, two parallel S B R reactors with built-in clarifiers, a chlorine contact tank with SO2 de-chlorination facility, provision for final effluent discharge into the Fraser river, two aerobic digesters, a belt filter press plus polymer feed system, and a screw conveyor. The screw conveyor sends the dewatered sludge into a storage warehouse. The stored sludge is then transported once a month by truck to a designated landfill. Two sludge lagoons are located on-site for 22 emergency storage of the digested sludge. Local septage is also dumped into a separate holding tank and is bled into the influent flow after being aerated for, on average, 7 days. Influent wastewater is constantly flowing into the SBRs, but the magnitude of the flow varies depending on the operation of the pumps at the headworks. Hence the SBRs influent feed is "semi-continuous", while the SBRs effluent withdrawal and sludge wasting are intermittent. The SBRs operate on a four-hour cycle, and the operation of the two SBRs is offset by two-hours. The waste sludge from the SBRs is pumped through a common underground pipeline and directed at the exit to the targeted aerobic digester. The plant operators control the SBR wasting rate manually to manage the Total Suspended Solids (TSS) level in the SBRs. Figure 1 shows the general schematic of the District of Kent Wastewater Treatment Plant. S B R s Influent Effluent Filtrate S B R Waste Sludge Dewatering Facility Decant Cl Contact Tank Aerobic Digesters l-J Digester Waste Sludge Dewatered Sludge Primary Treatment Effluent Discharge Figure 1. Pre-Experiment Schematic of Treatment Plant 23 The aerobic sludge digesters were managed under a batch operation mode. Unlike the SBRs, the two aerobic digesters were not operated in parallel. The waste sludge from both SBRs was fed exclusively into one aerobic digester. The digesters were continuously being aerated during this time. Once the digester reached capacity, the waste sludge from the SBRs were then directed toward the second digester, while the first digester that had reached capacity was being aerated. When the second digester was nearly full, the first digester under aeration was wasted and the digested waste sludge was sent to the belt filter press for dewatering. This then allowed the first digester to receive the waste sludge from the SBRs again, while the second digester was aerated. The time required for the digester to fill was wholly dependant on the volume of waste sludge coming from the SBRs. Moreover, the wasting rate of the SBRs generally conformed to the incoming influent flow to the treatment plant. Since the digester sludge was not wasted and dewatered until the digester reached capacity, the sludge age of the digester varied depending on the time of year. As preparation for digester sludge wasting and subsequent dewatering, the plant operators suspended aeration of the target sludge digester and allowed the sludge to settle overnight. The next morning, the supernatant was decanted via a telescoping weir to a well. Once the decant volume reached a certain level in the well, the installed pumps were automatically activated to pump the decant into the influent stream. The decant was blended with the influent that was being channeled to the SBRs for further treatment. Following the completion of decanting, the settled digester sludge was pumped to the dewatering facility, where the digested sludge was mixed with polymer and fed into the belt filter press. The screw conveyor sent the dewatered sludge cake into the storage area where it was stored until they were transported off-site. The two aerobic digesters at the treatment plant were equipped with coarse bubble aerators. The air from one source pump was shared between the two aerators in the digesters, an air-degritter in the primary treatment facility, and a small aerator in the septage holding tank. Plant operators controlled the airflow in to the digesters manually, but due to the 2 4 inconsistency in the air used in the other two sinks, it was difficult to supply a steady amount of air to both digesters. 3.1.2 Full-Scale Experiment The full-scale experiment designed for this research included the establishment of two parallel process trains, Experimental and Control. The two S B R s and the two aerobic digesters at the District of Kent Wastewater Treatment Plant proved to be a good fit for this proposed experimental design. B y establishing two parallel units, it allowed for direct comparison of results of the two sides since both were subject to the same conditions and operating guidelines. The main difference between the two flow streams was that the Experimental S B R had methanol injection. The methanol was mixed into the Experimental S B R following the completion of aeration to promote improved B N R . Table 1 explains the operation strategy of the Control and Experimental SBRs . The Control S B R was not modified from the existing pre-experiment settings. Table 1. Operation System of the Control and Experimental SBRs Cycle Time (minutes) Control SBR Stage Air On/Off Condition in SBR Experimental SBR Stage Air On/Off Condition in SBR 0 - 1 0 Idle and Waste Off Anoxic or Anaerobic Idle and Waste Off Anoxic or Anaerobic 10-110 Aeration On Aerobic Aeration On Aerobic 110-125 Settle Off Anoxic Methanol Addition Off Anoxic 125 - 170 Settle Off Anoxic or Anaerobic Settle Off Anoxic 170 - 240 Decant Off Anoxic or Anaerobic Decant Off Anoxic or Anaerobic 25 The methanol injection system was composed of a pump, a mixer, and an electrical timer. The pump was used to inject the methanol from the methanol tote tank (where the methanol is stored), to the propeller of the mixer that is installed in the S B R . The mixer was specifically installed in the Experimental S B R to allow the methanol to contact the mixed liquor throughout the entire volume of the S B R . A timer was used to turn on the mixer at the end of the aeration phase (110 t h minute), and turn it off 15 minutes later (125 t h minute) (Louzeiro, 1999). In order to form two parallel trains, modifications to the District of Kent Wastewater Treatment Plant were needed. Under the pre-experimental setting, as shown in Figure 1, there would be no segregation of the waste sludge from the respective SBRs , the decant from the respective digesters, and the filtrate produced from dewatering the waste sludge of the respective digesters. This meant that the waste sludge from the Experimental S B R would be discharged into the Control digester along with the waste sludge from the Control S B R , and the decant and filtrate from the Experimental digester would be recycled into the Control S B R , as well as the Experimental S B R . As mentioned earlier in Section 3.1.1, the decant from the digesters is introduced and blended into the influent stream. In order to direct the decant into the desired S B R tank, a surface pipe with valves were constructed. Given that the decanting of the two digesters are done at separate times, two separate pipes were not found to be necessary. This modification of flow path allowed for complete segregation of the two decants. Figure 2 portrays the modified schematic of the District of Kent Wastewater Treatment Plant during the experiment. Actuator valves, which can be programmed and automated, were installed to each digester so that the incoming S B R sludge could be directed to the desired digester. This configuration allowed the proper digester to receive the S B R sludge during wasting that occurs once every cycle. The wasting volumes have ranged from 1.5 to 6.6 m 3 per cycle 26 over the duration of the research. The S B R sludge is pumped from the S B R to the digester through an underground pipe of 10 cm diameter and a total length of approximately 70 m, prior to reaching the actuator valves in the digesters. When the pump stops wasting from the S B R , the sludge that remains in the underground piping stays until the next batch of wasting starts from the opposite S B R . A s a result, approximately 0.55 m 3 of the sludge from the previous wasting was mixed with the next batch of wasting. Methanol y Tote Tanks Methanol Influent -Mixer Filtrate Decant Primary Treatment Valves Effluent Cl Contact Tank Experiment Digester Control Digester Actuator Valves Experiment S B R Control S B R S B R Waste Sludge Digester Waste Sludge Dewatered Sludge Effluent Discharge Figure 2. Modified Schematic of Treatment Plant For Experiment The treatment plant is configured such that the filtrate from dewatering is collected into an underground well that is directly underneath the belt filter press. The filtrate is then pumped and transported via an underground pipe and discharged into the influent flow 27 stream. Therefore, the filtrate could not be fed into the appropriate corresponding S B R tank and were blended with the influent and distributed back into both SBRs . A s explained above, the plant configurations at the District of Kent Wastewater Treatment Plant made the complete separation of the Experimental and Control process trains difficult. The impracticably and cost of the necessary construction that is required to overcome the problem was determined to be greater than the expected minor effect that the slight flow integration should have on the overall experiment. Moreover, such uncontrollable circumstances were expected since this project was a full-scale, real-world, demonstration study. Three methanol dosages were applied to the Experimental S B R , 0.3, 0.6 and 0.9 L/min respectively. These dosages were chosen in an attempt to determine the optimal methanol dosage for biological nutrient removal. Since there are six cycles daily, the daily methanol dosages would then be 27, 54, and 81 L/day respectively. The average volume of the S B R is 875 m 3 ; hence, the approximate methanol concentration in the Experimental S B R for each dosage was 4.1, 8.1, and 12.2 mg C H 3 O H / L (Louzeiro, 1999). This full-scale experiment was comprised of four experimental runs. Prior to the start of Run 1, a baseline was established for the parameters of interest in both process trains. These results can be found later in Section 4. The first two dosages were applied for one month, and the third dosage for a two-month duration. Following the four-month methanol injection period, the system was run for one month without methanol injection. Table 2 below lists the experimental runs that were undertaken in this full-scale experiment. 28 Table 2. Summary of Experimental Runs Run Duration Methanol Dosage Pre-run Establish Baseline 0 L/Day 1 4 weeks 27 L/Day 2 4 weeks 54 L/Day 3 8 weeks 81 L/Day 4 4 weeks 0 L/Day Unlike the pre-experiment operations, the aerobic digesters were operated in a more methodical fashion by following a regular schedule. At the start of the experiment, the two digesters wasted enough sludge that the two digesters would start with the same volume level of 100 m 3 , which was about half of the maximum capacity of the digester of 200 m 3 . For the first two methanol dosages from the start of the experiment, decanting, wasting and dewatering of digesters were done every two weeks to conform to the low rate of incoming S B R waste sludges. However, with the increase of influent flow and methanol dosage, the S B R wasting also increased, and from the third dosage on the decanting, wasting and dewatering were undertaken every week until the completion of the experiment. In attempt to synchronize the operation between the Experiment and Control sides, the decanting, wasting and pressing of both digesters were carried out on consecutive days. 3.2 Sampling and Analysis A comprehensive strategy for sampling and analysis was planned for this study and is described in detail in this section. 29 3.2.1 Sampling Plan Detailed summaries of the sampling plan employed in this full-scale experiment are shown in Tables 3. For both the Experimental and Control process trains, all sampling was done on consecutive days, when each digester was decanted, wasted, and dewatered. The main objective of this research was to monitor the change in aerobic digester performance with methanol addition. Two of the major factors to profile the aerobic digester performance were solids reduction and nutrient re-release and accumulation. To obtain results for solids reduction performance, it was necessary to establish a mass balance around the aerobic digesters, by analyzing the amount of solids that were incoming to, wasted from and present in the digester. The nutrient re-release and accumulation in the digesters required the analysis of the different nitrogen and phosphorus levels, such as ammonia (NH3), nitrates and nitrites (NOx), total kjeldahl nitrogen ( T K N ) , orthophosphate (P0 4 ) , and total phosphorus (TP). To investigate any signs of fermentation during the digester's overnight settling period before wasting and dewatering, the digester decant and belt press filtrate were analyzed for volatile fatty acids (VFAs) . The list of samples used in this study and their actual sample locations along with the sample types are shown in Table 4. A l l locations were sampled multiple times during the pre-run stage in the experiment to determine the number of samples required. A l l point source samples showed negligible differences in results, hence, one sample was taken during all subsequent sampling days. A l l grab samples were repeated three times and blended to make one sample for analysis. 30 Table 3. Summary of Sampling Plan for Experiment and Control Analytical Parameters Sample Description Total Suspended Solids (TSS) Volatile Suspended Solids (VSS) Influent Sewage SBR Waste Sludge Digester Sludge Digester Waste Sludge Digester Decant Filtrate Sludge Cake Final Effluent Ammonia (NH 3) Nitrate and Nitrite (NO x ) Orthophosphate (P0 4) Soluble COD (COD s ) Influent SBR Waste Sludge Digester Sludge Digester Waste Sludge Digester Decant Filtrate Final Effluent Total Kjeldahl Nitrogen (TKN) Total Phosphorus (TP) Influent SBR Waste Sludge Digester Sludge Digester Waste Sludge Digester Decant Filtrate Sludge Cake Final Effluent Volatile Fatty Acids Digester Decant Filtrate 31 Table 4. Sample Locations and Sample Types Sample Source Sample Location Sample Type Influent (into SBRs) Flow channel outside of primary treatment facility Grab Samples SBR Waste Sludge Tap at pipe containing SBR waste sludge Point Source Digester Sludge Digester Grab Samples Digester Waste Sludge Tap at pump containing digester waste sludge Point Source Digester Decant End of recycle pipe discharging decant Point Source Filtrate End of pipe discharging filtrate in belt filter press Point Source Sludge Cake Belt from Belt filter press Grab Samples Final Effluent End of pipe discharge under manhole Point Source Each sampling day required and followed a strict schedule, to ensure that for every week, each sample was collected at or near the same time. This was undertaken to synchronize the sampling as much as possible and minimize the possible source of error. 3.2.2 On-line Monitoring Plan Several on-line automated means of data gathering were used in this experiment to monitor certain parameters. Temperature is a critical parameter that controls the rate of biological reactions and controls what microorganisms can thrive in an environment. Nitrifying bacteria are extremely sensitive to various inhibitors including pH (Metcalf and Eddy, 1999). Temperature and pH of both Experimental and Control digesters were measured on a continuous basis by submerging on-line probes into the aerobic digesters. The data was collected using an 8-channel data logger and software manufactured by Lakewood Systems Ltd. of Ontario, Canada. The temperature probes used in this 32 experiment were also manufactured by Lakewood Systems Ltd. The p H probes used were products of Hanna Instruments. A l l probes were cleaned weekly, using warm soapy water. The calibration of the p H probes was done once a month according to Standard Methods ( A P H A , 1996). The central computer at the District of Kent Wastewater Treatment Plant was used to monitor and record the following parameters: influent flow rate, effluent flow rate, digester level, decant volume, and digester sludge wasting volume. Table 5 summarizes the parameters that were monitored during the experiment. Table 5. On-line Monitored Parameters Monitored Parameter On-line Instrument Frequency Digester Temperature (°C) Temperature Probe Continuous Digester pH pH Probe Continuous Influent Flow (m'Vday) Central Computer Daily SBR wasting rate (mVcycle) Central Computer Daily Digester Level (m3) Central Computer Daily Decant Volume (mJ) Central Computer Bi-weekly/Weekly Digester Sludge Wasting Volume (m3/day) Central Computer B i-weekly/Weekl y The efficiency of the pump that transfers the waste sludge from the S B R to the digester was suspected to vary with the solids level in the sludge. It was thought to be important to determine the correlation between wasting time and wasting volume. Furthermore, the daily S B R waste sludge volume recorded by the plant operator was found to be outdated, since the correlation was several years old and had not checked recently (Nuttall, 1999). To determine the exact volume of S B R waste sludge that was pumped into the digester, the change in digester level was recorded after each wasting. Using the recorded data, a 33 correlation between wasting time and wasting volume was determined. This process was repeated every 4 weeks for verification. 3.2.3 Analytical Methods Table 6 summarizes how the samples were handled in accordance with the desired analysis. The sample preservation was done based on the Standard Methods ( A P H A , 1996) and the guidelines provided by a Canadian professional environmental laboratory (Analytical Service Laboratory Ltd.). Filtered samples were obtained using a vacuum filter (Whatman No.934-AH). The oxygen uptake rate (OUR) test and the capillary suction time (CST) test required the tests to be done on-site, in order to guarantee the best possible results. A l l of the other analyses were carried out at the University of British Columbia Environmental Engineering Laboratory. A l l analytical procedures used in this research followed the guidelines outlined in Standard Methods ( A P H A , 1996). Table 6. Sample Handling and Preservation Analytical Test Filtered Storage Method Preservative N O x / P 0 4 / N H 3 / C O D s Yes Up to 28 days at 4 °C Add sulfuric acid to pH <2 T K N / T P No Up to 28 days at 4 °C Add sulfuric acid to pH <2 VFAs Yes Up to 1 month at 0 °C Add 0.1 m L o f 2% phosphoric acid TSS / VSS No Up to 24 hours at 4 °C None Oxygen Uptake Rate No Analyze immediately None Capillary Suction Test No Analyze immediately None 34 3.2.3.1 Total Suspended Solids (TSS) and Volatile Suspended Solids (VSS) The TSS and V S S samples were immediately preserved at 4 °C and analysis was carried out at the U B C Environmental Engineering Laboratory within 24 hours of the sampling time. It is imperative that the solids analysis be performed immediately since with time, the solids level may decrease as it becomes anaerobic and becomes digested, and thus resulting in a faulty TSS and V S S values. The samples were brought to room temperature before the analysis was performed. The following is the summary of the analytical steps performed: • an aluminum dish with a glass fiber filter paper (Whatman No.934-AH) placed inside the dish was fired in a muffle furnace at 550 °C for 15 minutes. After firing, the dish was transferred to a dessicator where it was allowed to cool. This pre-firing step is necessary to ensure that the glass fiber filter paper w i l l not suffer any loss of weight due to ignition during firing of the sample, which would bring about an incorrect result; • the tare weight of the aluminum dish and the fired glass fiber filter paper was measured and recorded; • the fired glass fiber filter paper was placed on top of the head of the vacuum filter and the filter paper was wetted with distilled water in order to completely seal the r im area of the vacuum filter head; • 10 ml for sludges, 100 ml for decants and filtrates, and 1 L for influent and effluent samples were prepared and were mixed thoroughly; • the sample was filtered slowly through the fired glass fiber filter on the vacuum filter apparatus; • following filtering, any residual solids that remained in the inner walls of the vacuum filter head were scraped and placed on the fiber filter that contains the filtered solids; • the filter was placed on the aluminum dish; • the sample volume, sample description, and dish identification was recorded; 35 • the aluminum dish and the filter paper were placed in the 103 °C oven to dry overnight; • the dish was transferred to a dessicator where it was allowed to cool, and was weighed; the increase in mass represents the total suspended solids of the sample; • the dish was fired in a muffle furnace at 550 °C for 15 minutes; and • after firing, the dish was allowed to cool in a dessicator, and was weighed;.the loss of mass represents the volatile fraction of the suspended solids that was removed via ignition. 3.2.3.2 Ammonia (NH3), Nitrate and Nitrite (NOx), and Orthophosphate (P04) Total nitrogen consists of organic Nitrogen, ammonia ( N H 3 ) , nitrate ( N 0 3 ) , and nitrite ( N 0 2 ) . The total of nitrate and nitrite is also recognized as NOx- It is understood that the growth rate of Nitrosomonas, which oxidizes ammonia to nitrite, is less than that of Nitrobacter, which oxidizes nitrite to nitrate. Therefore, nitrite accumulation should not occur unless Nitrosonomas is inhibited. For this reason, the N O x concentration was considered as NO3 in this experiment. Orthophosphate is one of the forms phosphorus exists in aqueous form along with polyphosphate and organic phosphorus. The soluble inorganic phosphorus is predominantly orthophosphate. The following is the summary of the analytical steps performed: • the filtered samples that were preserved at p H of under 2 at 4 °C were withdrawn into vials for analysis using the Lachat automated ion analyzer; • the operation of the automated ion analyzer was performed by a qualified laboratory technician; and • QuikChem method No. 10-107-06-1-F was used for ammonia, method No. 10-107-04-1-E for N O x , and No. 10-115-01-1-D for orthophosphate. (Lachat Instruments, Inc., 1990) 36 3.2.3.3 Total Kjeldahl Nitrogen (TKN) and Total Phosphorus (TP) Total Kjeldahl nitrogen ( T K N ) is the total of the organic and ammonia nitrogen. Therefore, the amount of organic nitrogen in sample can be calculated using the T K N and ammonia results. Total phosphorus is the measurement of all phosphorus forms present in a sample. The organic phosphorus content can be determined by subtracting the inorganic phosphorus concentration from the total phosphorus. The following is the summary of the analytical steps performed: • QuikChem digestion and extraction method No . 10-107-06-2-D was used for T K N and No. 10-115-01-1-C was used for T P (Lachat Instruments, Inc., 1990); • 0.5 m L of sample was used for all sludge samples and 5 m L was used for decant, filtrate, influent and effluent samples; • the sample was withdrawn using a pipet into a digestion tube; distilled water was used to rinse any sample solids that attached to the inner wall of the tubes; • 10 m L of sulfuric acid digestion solution was added and mixed; • 3 or 4 boiling rocks were added to the digestion tube to prevent vigorous boiling during the digestion process; • the digestion tubes were placed in a preheated block digester for 3.5 hours at 130 °C to vaporize all water content; followed by 2.5 hours at 380 °C; • the samples were allowed to cool and ammonia-free distilled water was used to f i l l the digestion tube while shaking to ensure mixing with the digested sample; • the mouth of the tube was then closed using a rubber plus and the tube was overturned three or four times to allow for good mixing; and • 5 m L was withdrawn from the digestion tube for T K N and T P analysis by a qualified technician using the Lachat automated ion analyzer. 37 3.2.3.4 Soluble Chemical Oxygen Demand (CODs) The amount of organic substrate in decant and filtrate were of great interest in this full-scale experiment, since both were recycled and reintroduced into the SBRs, along with the influent. Therefore, it was important to calculate the amount of soluble organic substrate that was contained in the decant and filtrate. The COD test is widely accepted as a means of measuring the organic strength of wastewater. Compared to the BOD test, it only requires a fraction of the time and is a easier test to perform. The following is the summary of the analytical steps performed: • COD vials were prepared with reagents for a maximum COD concentration of 200 ppm; • 2 mL of each sample was placed in the vials and closely capped; samples that exceed the 200 ppm COD range were diluted before injecting into vial; • the vials were shaken vigorously to ensure good mixing; • standards of 0, 10, 20, 50, 100, and 200 ppm were prepared with potassium hydrogen phthalate; • the vials were digested in a block digester for 2 hours, then were allowed to cool to room temperature; and • the vials were wiped and cleaned in preparation for analysis; • a spectrophotometer set at 600nm wave length was used read the absorbance of light by the sample vial; • the spectrophotometer was calibrated according to the 0 ppm standard and the absorbance of the other standards was measured; • a calibration curve of absorbance vs. COD concentration was constructed; and • samples vials were measured for absorbance and COD was determined using the calibration curve. 38 3.2.3.5 Volatile Fatty Acids (VFAs) Volatile fatty acids (VFAs) , especially acetic acid, are the best-known fermentation products to support biological phosphorus removal (Mino et al., 1987). A n anaerobic zone is required as a fermentation step to produce V F A s by normally occurring facultative organisms (Comeau et al., 1986). The aerobic sludge digester is not aerated during overnight settling and during decanting, wasting and dewatering. Since the decant and filtrate are recycled into the SBRs , it was beneficial to investigate whether any V F A production could occur inside the digester, and thus assist in improved biological phosphorus removal. The samples filtered specifically for V F A analysis were preserved with 2% phosphoric acid and were frozen immediately. The samples were analyzed within one month of sampling. The following is the summary of the analytical steps performed: • the frozen samples were allowed to thaw and extracted into test vials; • analysis was performed by qualified laboratory technician using the computer-controlled Hewlett-Packard 5880A Gas Chromatograph (Hewlett-Packard Company, 1995), equipped with a flame ionization detector (FID); and • the Supelco G C Bulletin 7 5 I G was used as the guideline and helium was used as the carrier gas. 3.2.3.6 Oxygen Uptake Rate (OUR) Test The oxygen uptake rate (OUR) test was used to monitor the biological stability of the sludge. Dividing the O U R rate by the V S S of the sample yields the specific oxygen uptake rate (SOUR). S O U R is a commonly used and widely accepted indicator of sludge stability along with the reduction of volatile solids (Tonkovic, 1999). Since the S O U R is the measure of biological activity, it is beneficial to perform the test immediately after sampling. The sample of digester waste sludge was aerated within 5 minutes from time of sampling and the test was conducted no later than 15 minutes from time of sampling at the 39 in-house laboratory of the District of Kent Wastewater Treatment Plant. The following is the summary of the analytical steps performed: • 500 m L of the digester waste sludge sample was poured into a beaker, capped the beaker; • sample was immediately aerated using a fish pump and a magnetic stirrer; • a basin of controlled temperature was used to keep the sample at constant temperature; • the dissolved oxygen (DO) probe was calibrated and the probe was placed in the sludge sample being aerated; • the sludge sample was aerated until the dissolved oxygen reading has stabilized; turned off the fish pump and the magnetic stirrer, • the initial D O reading and temperature were recorded; • using a stop watch, the D O was recorded every 30 seconds; data was recorded over a 15 minute period or until D O became limiting, whichever occurred first; and • the oxygen consumption rate (OUR) is the drop in D O divided by the time duration; the test was repeated two more times with fresh sludge samples. 3.2.3.7 Capillary Suction Time (CST) Test To compare the dewaterability of the Control and Experimental digester waste sludges, the capillary suction time (CST) test was chosen. The specific resistance to filtration (SRF) test was also available, but the C S T test can be done in a much shorter period. Although there are, individuals who argue that the C S T test is purely an empirical test and is not based on a theoretical analysis of sludge dewaterability, several researchers have provided a model which links the C S T in a linear correlation with S R F (Vesilind, 1988), thus giving merit to the C S T test as a research tool for sludge dewaterability. It is commonly understood that the sludge sample must be tested as soon as possible since biological degradation of the sludge w i l l affect the result of the C S T test (Triton Electronics Ltd. , 40 1998). The sludge sample was tested within 5 minutes after sampling to ensure the best possible results. The following is the summary of the analytical steps performed: • the sample was mixed to ensure homogeneity; • the filter paper was placed on the C S T unit with the rough side on top and the electrode block was placed on top of the filter paper; • the funnel was inserted into the electrode block, then rotated and applied downward pressure to insure even contact on the filter paper; • ensure the timer is reset to zero seconds; and • pour sludge sample into funnel reservoir, and record resulting C S T time. 3.2.3.8 Microscopic Examination for Identifying Filamentous Microorganisms Microscopic examination of the sludge is necessary to assess the condition and macrostructure of the sludge floe and to determine the quantities and identification of the filamentous microorganisms. The samples were taken from the sludge digesters as a grab sample, and the sample was brought to the treatment plant in-house laboratory for immediate examination. The sample was stained using the Gram stain - modified Hucker method described in Standards Methods ( A P H A , 1996). The following is the steps taken to stain and observe with a microscope: • prepared thin smears of a sludge sample on microscope slide and thoroughly air dried; • stained for 1 minute with crystal violet solution; rinsed for 1 second with running water; • stained for 1 minute with iodine solution; rinsed for 1 second with running water; • decolorized by adding ethanol (95%) drop by drop and allowing it to smear for 30 seconds; rinsed and blot dried; • examined under microscope at lOx magnification and 40x magnification with direct illumination; 41 • inspected color of sample; blue-violet is Gram-positive and red is Gram-negative; and • inspected the shape, length, diameter, and branching to identify the filamentous microorganism. 3.2.4 Quality Assurance and Quality Control Quality assurance and quality control (QA/QC) programs are vital to ensure that analytical data are legitimate. Experienced, qualified laboratory technicians at the U B C Environmental Engineering Laboratory performed the majority of the laboratory analyses in this research, and conducted analyses with a stringent QA/QC program. The technicians also perform routine maintenance and calibrations of the analytical instruments with ready standards. Historical records of test samples are kept to assess the reliability of the various analyses. These test samples are used regularly to identify and fix any instrument problems. The following is a summary of the QA/QC measure that was taken during sampling and analyses to assure the validity of the analytical results: • Frequent calibrations using single-point checking were applied with a standard solution, Single-point checking requires the preparation of known concentrations of a standard solution. The results are compared with the previous standard reading, and i f the error is within 10 percent the analysis of samples were proceed. If the results are outside this range, a secondary standard must be prepared and assessed similarly. If the second standard also fails, the calibration problem must be corrected before any further sample measurements are made (Ramamoorthy, 1997); • a test solution of a known concentration was analyzed to check the calibration of the analytical instrument; • field blanks were taken to detect any contamination; 42 • during transport all samples were kept in an ice box and were promptly transferred to a refrigerated (at 4 °C) dark room for storage upon arrival to the U B C Environmental Engineering Laboratory; and • duplicate samples were routinely taken to check the precision and that the sample size is adequate. 3.2.5 District of Kent Wastewater Treatment Plant Data Collection Program The District of Kent Wastewater Treatment Plant laboratory is well equipped to handle various types of analyses. Trained laboratory technicians perform an extensive amount of sample collection and analysis. Historical data are available from January 1997 for the parameters listed in Table 7. 43 Table 7. District of Kent Wastewater Treatment Plant Laboratory Analyses Parameter Sampling Location Frequency Flow (nrVday) Influent Daily Decant, digester waste sludge Bi-weekly/weekly Volume level (m3) Digesters Daily Temperature (°C) Influent, effluent, SBRs, digesters Daily pH Influent, effluent, SBRs, digesters Daily DO (mg/L) Influent, effluent, SBRs, digesters Daily Alkalinity Influent, effluent Monthly N H 3 (mg/L) Influent, effluent Twice weekly P 0 4 (mg/L) Influent, effluent Twice weekly TSS, VSS (mg/L) SBRs Daily Influent, effluent, primary effluent, SBR waste sludges, digesters Twice weekly BOD5 (mg/L) Influent, effluent Daily COD (mg/L) Influent, effluent Daily T K N (mg/L) Influent, effluent Bi-weekly C l 2 (mg/L) Effluent Daily S 0 2 (mg/L) Effluent Daily Colour (FAU) Effluent Daily Turbidity (pt) Effluent Daily Fecal coliform (count/100ml) Effluent Bi-weekly F / M Ratio SBRs Weekly 44 4 RESULTS AND DISCUSSION This full-scale research experiment is a part of a full-scale demonstration study at the University of British Columbia to examine the feasibility of using methanol for biological nutrient removal purposes. This full-scale experiment was designed to study the effect of methanol addition in the SBR on the aerobic digestion unit. The results obtained from this experiment are detailed and discussed in this section. Included in this section is the report on the digester sludge reduction performance, nutrient re-release and accumulation, sludge stability, and sludge dewaterability. Moreover, the difficulties facing a full-scale experiment, the solids production in SBR due to methanol addition, and general performance issues such as, sludge bulking, VFA production, seasonal variations, and scale-up potential are discussed in this section. 4.1 D i s t r i c t o f K e n t W a s t e w a t e r T r e a t m e n t P l a n t The joint study with Louzeiro (1999), to examine the feasibility of using methanol for biological removal, was carried out at the District of Kent Wastewater Treatment Plant, in Agassiz, BC, Canada. This section profiles some of the challenges that surfaced during the study, and the results of the pre-run analysis to characterize the plant and establish a baseline. 4.1.1 Challenges of a Full-Scale Experiment Bench-scale or pilot-scale experiments are often used in environmental engineering research. Any new design modification or advancement in process technology is first evaluated in a bench-scale setup rather than a full-scale setup. The low cost and the ability to control variables make the bench-scale setup an ideal way to test new ideas in 45 environmental engineering. However, a process verified in bench-scale cannot be considered a complete success until it is assessed in a full-scale setup, since the bench-scale experiment is operated under controlled circumstances (hardly the case in the real world of wastewater treatment). Controlled environment and steady-state are unrealistic conditions in many full-scale wastewater treatment plants. Numerous uncontrollable variables such as weather, climate, incoming flow, existing plant configurations, and even human error make the need for a trial demonstration study in a full-scale setup necessary for any new process technology to gain approval and ultimately be marketed. This full-scale demonstration study, as expected, brought about many, real-world problems, that were mostly uncontrollable or highly impractical. The following are the notable variables and difficulties that were observed in this project: • Plant Configuration. The plant had to be modified from its original operating system specifically to segregate the flows of the SBR waste sludge, digester decant and filtrate, to create two parallel process trains. As documented in section 3.1.2, to completely segregate the Control and Experiment SBR waste sludges, and to reroute the filtrate to the appropriate SBRs, would have required costly and impractical construction work for a temporary setup; • Overnight Settling. The aerobic sludge digesters were allowed to settle overnight with the aerator off before the digester is decanted, wasted and dewatered. Overnight settling was unpredictable as floating sludge, sludge bulking, and good settling were observed at different times for both aerobic digesters during this experiment; • Weather. Changing temperature and precipitation are the main variables that affect the sludge digestion process. As seasons change ambient temperature and fluid temperature change gradually. Temperature affects biological activity. Precipitation which could also affect temperature, mostly affecting incoming flow in the form of storm water; 46 • Influent flow. Overall trend of the incoming influent flow volume was mostly predictable on a season-to-season basis, but the influent characteristics could change without notice; and • Human Factor. Plant operator work hours often restricted the time available for sludge dewatering therefore limiting the amount of digester sludge that can be wasted and dewatered. In addition, plant operators were trained and experienced professionals but the operators and the research team members were not immune to occasional mistakes during the duration of this study (i.e. one dewatering day that had to be cancelled as the polymer had run out without any in reserve); Most results in this experiment were difficult to accurately quantify in the form of a formula or a model, due to these variables and a non-steady state condition. However, by compiling weekly data to produce an average monthly result for each run, the results were representative of the overall trend and the main goal of profiling the change in aerobic digester performance was achieved. 4.1.2 Pre-Run Results: District of Kent Wastewater Treatment Plant Profile Prior to Run 1, when the first methanol injection started, a two-month pre-run period from January to March 1999 was undertaken to establish a baseline for various parameters of interest at the plant. Table 8 characterizes the influent, effluent of the District of Kent Wastewater Treatment Plant. The influent profile conforms to a typical medium-strength municipal sewage, except for a high orthophosphate concentration. The effluent was of excellent quality and well under the legal limits of effluent discharge in British Columbia, Canada. Local septage, which was transported and dumped into an aerated holding tank, is profiled and compared to typical values found in septage, in Table 9 (Metcalf and Eddy, 1991). Septage was aerated 47 for 7 days, on average, before it was bled into the influent flow stream to reduce the oxygen demand and thus reduce the load on the SBRs. Table 8. Influent and Effluent Profiles (Jan. - Mar. 1999) Parameter (units) Influent (mean ±std deviation) Effluent (mean ± std deviation) Total Suspended Solids T S S (mg/L) 211 (± 83) 4.4 (± 3.4) Biochemical Oxygen Demand B O D (mg/L) 252 ( ± 8 1 ) 4.2 (± 4.6) Chemical Oxygen Demand C O D (mg/L) 420 (± 165) 13 (± 12) Ammonia N H 3 - N (mg/L) 33.6 ( ± 8 . 8 ) 0.2 (± 0.3) Orthophosphate P 0 4 - P (mg/L) 11.8 ( ± 2 . 6 ) 2.0 (± 0.5) Table 9. Profile of Septage After 7 day Aeration and Typical Septage Parameter (Units) Typical Septage Septage Sample after 7 days aeration Total Suspended Solids T S S (mg/L) 15000 11600 Volat i le Suspended Solids V S S (mg/L) 7000 8300 Soluble Chemical Oxygen Demand C O D s (mg/L) 30000 (COD) 261 Ammonia N H 3 - N (mg/L) 400 0.07 Nitrate N 0 3 - N (mg/L) Not Available 0.19 Orthophosphate P 0 4 - P (mg/L) Not Available 0.75 Total Kjeldahl Nitrogen T K N (mg/L) 700 188 Total Phosphorus T P (mg/L) 250 56 Average Weekly Volume of Septage Dumped (mVweek) Not Applicable 25 4 8 The average weekly septage volume of 25 m was an estimate based on the plant operator's experience that about half of the septage holding tank (approximately 50m 3 capacity) was filled when the septage was dumped (Nuttall, 1999). 25 m 3 per week can be translated into 3.6 m per day and, considering that the average influent flow over the duration of the experiment was approximately 1160 m" per day, it was determined that the impact of the reduced-strength septage to the influent flow and the experiment was minimal. The operating characteristic of the two SBRs , and two aerobic digesters during the pre-run period, are profiled in Table 10. The SBRs were operated extremely conservatively, similar to an extended aeration process, with sludge ages of over 60 days. According to the plant operator, the target operating total suspended solids level in the SBRs is about 4000 mg/L (Gardner, 1999). The recommended operating total suspended solids level in the SBRs is 3500 mg/L, as suggested by Stantec Consulting at the time of plant retrofit in 1993 (Stantec Consulting, 1993). Table 10. Pre-Run SBR and Aerobic Digester Operating Characteristics Parameter (Units) SBRs Digesters Temperature (°C) 12.9 12.3 PH 6.6 6.5 Dissolved Oxygen (mg/L) 3.0 2.9 Total Suspended Solids TSS (mg/L) 4400 16800 Volatile Fraction (%) 79 73 Sludge Age (days) 64 Not Available 49 The aerobic digester waste sludges sampled during the pre-run stage are outlined in Table 11. A s indicated, the aerobic digesters showed erratic values in terms of volatile suspended solids reduction percentage. This was primarily because the decanted and wasted volumes of the digester varied widely from week to week. Considering the time restrictions with respect to the plant operators' work hours, the overnight settling characteristic of the digester sludge and the amount of solids in the digester generally determined how much would be decanted and wasted. Table 11. Aerobic Digester Waste Sludge Profile Parameter (Units) Sample Date 1/20/99 3/10/99 Total Suspended Solids TSS (mg/L) 17500 16200 Volatile Fraction (%) 73 76 Volatile Suspended Solids Reduction (%) 52 25 Sludge Age (days) 60 31 Specific Oxygen Uptake Rate SOUR (mg02/gVSShr) 2.0 Not Available Capillary Suction Time CST (sec) Not Available 53 Solids content of Sludge Cake (%) 15 15 As described in Section 3.1.2, there were four experimental runs in this full-scale experiment. These runs are differentiated by the injected methanol dosage. Runs 1 and 2 started in March 11 t h 1999, with a methanol dosage of 27 and 54 L/day, respectively, and were applied for 4 weeks each. Run 3, with a methanol dosage of 81 L/day, was applied for 8 weeks. Run 4 with no methanol dosage, was run for 4 weeks until August 4 t h 1999; this was undertaken to observe what happens after methanol addition is discontinued, 50 since the changes in the Experimental S B R could take some time to be reflected in the Experimental digester. 4.2 Solids Production in SBR The addition of a short-chain carbon compound, like methanol ( C H 3 O H ) , would cause extra biomass growth in the Experimental S B R and result in an increase in total suspended solids (TSS) level in the Experimental S B R . There would thus be increased wasting toward the Experimental digester, since the treatment plant operators control the S B R wasting rate to manage the TSS level in the SBRs . The wasting of Experimental and Control S B R sludges is shown in Figure 3. The volume of S B R sludge wasting is determined by developing a correlation between the wasting rate and the wasting volume. A s aforementioned in Section 3.2.2, the existing correlation developed by the plant operator was inappropriate for this research project. Therefore, a new correlation was determined by recording the increase in digester level after each S B R sludge wasting. The operators require time to observe and assess the situation with the S B R ' s TSS level, before any action is taken to adjust the wasting rate. The immediate wasting rate is not an exact indicator of the actual difference in solids produced in the SBRs; however, it is still representative of the overall trend. Figure 3 was generated using multiple wasting rates to generate one representative wasting rate per run. From Figure 3, the discrepancy between the experiment S B R wasting rate and the control S B R wasting rate is apparent and, it shows that the wasting to the Experimental digester increased with increasing methanol dosage. During the two, 1-month periods in Run3, the Experimental S B R wasted, on average, 34 and 41% more than the Control S B R . Following the 4-month run of methanol, the two wasting rates became nearly identical after one month. 51 (^Bp/§jj) spips papuadsng IBJOX o ON p p q p p p p p p p p > r i o > / " > 0 > o o i f " i o < r i o > / " i (teafit) spijos papuadsng jmox 53 The increase in solids production was determined by taking the difference between the total amount of solids wasted from the Experimental S B R with methanol addition, and the Control S B R without methanol addition. Figure 4 plotted the increase in solids production in the Experimental S B R against the applied methanol dosage. The linear relationship shows about 0.39 kg of solids produced per L of methanol. This, however, is not based on a rigorous and detailed analysis; as such, it is presented here as an approximation. According to Louzeiro, based on the same full-scale experiment and settings, the theoretical model, to examine solids production, revealed an estimation of approximately 0.17 kg of solids produced per L of methanol, while the kinetic model presented an estimation of approximately 0.44 kg of solids produced per L of methanol. (Louzeiro, 1999) The observed result of 0.39 kg of solids produced per L of methanol is very close to that predicted by Louzeiro's kinetic model. While the theoretical model was a generalized universal model, the kinetic model utilized the historical data of the District of Kent Wastewater Treatment Plant to derive the yield coefficient and therefore produced a treatment-plant-specific result that can only be applied to the results obtained from this plant. The proximity of the kinetic model and the observed model supports the claim that the observed model based on wastage to the digester, was representative of the overall trend and provides merit to this full-scale experimental setup and results. 4.3 Solids Reduction Performance Solids reduction is the single most important performance characteristic of sludge digestion. Transporting and disposal of sludge end products is a significant cost item, therefore, decreasing the amount of solids that needs to be dealt with, is an important goal for designers and treatment plant operators alike. Total suspended solids (TSS) destruction and volatile suspended solids (VSS) destruction are the usual parameters that quantify the solids reduction in the digesters. V S S reduction has received a wider acceptance as a better 54 digester performance parameter since the majority of the solids reduction takes place in the biodegradable content of the sludge. Table 12 summarizes the mass balance analysis that was used to estimate the V S S reduction and sludge age in the digesters and is based on the widely accepted work of Koers (1979). Table 12. Mass Balance Analysis Parameter Formula VSS destroyed VSS reduction % Sludge age = Total VSS incoming - Total VSS removed from digester (decant and wasting) - Change in total VSS in digester = (VSS destroyed / Total VSS incoming) x 100 % = Total VSS in digester / Total VSS removed daily from digester (decant and wasting) Figure 5 shows the V S S reduction achieved by both Control and Experimental digesters over the four runs. Figure 6 plots the sludge age of both Control and Experimental digesters. Figures 5 and 6 show the Control values increasing and the Experimental values decreasing in the stages, pre-run and Run 1. A s mentioned in Section 3.1.1, the operational scheme of the aerobic digesters before this study was different than the scheme during the study. Prior to the start of the experiment, the Experimental digester did not receive and did not waste any sludge for an extended period of time, since the S B R waste sludge was sent to the Control digester; this directly affected the V S S reduction and sludge age of the pre-run values that were used to determine the baseline. 55 The V S S reductions fluctuated wildly over the course of the experiment, with an upper reduction rate of 54% and a lower rate of 20% for both digesters. The suspected reason for the observed troughs in Figure 5, where both digesters registered the lowest V S S reductions, was the high solids load coming into the digesters from the SBRs , as profiled in Figure 3, thereby shortening the sludge age (see Figure 6). The sludge age has a direct affect on the solids reduction performance of the sludge digestion process, since the longer sludge age means that the digesting biomass was kept in the digester for a longer period of time. This point is further supported by the U S Environmental Protection Agency's conclusion, based on both full-scale and pilot-scale studies, that solids destruction is primarily a direct function of both liquid temperature and sludge age (US E P A , 1979). Figure 6 shows that the sludge age results are consistent with the solids reduction and incoming S B R waste sludge flow. The Experimental digester generally had a shorter sludge age than that of the Control digester, with the difference varying from approximately 6 days to up to 29 days. Extra solids produced from methanol addition in the Experimental S B R resulted in the Experimental digester receiving more S B R waste sludge than the control digester, which resulted in the increase of sludge digester wasting, and therefore shortened the sludge age. However, during the two months of Run 3, the V S S reduction of the Experimental and Control digesters were nearly identical, despite a clear difference in sludge age for the two sides. In order to investigate further, the V S S reduction results, with the corresponding, temperature and sludge age, were plotted against the industry standard theoretical curve, that depicts V S S reduction as a function of temperature and sludge age (Water Pollution Control Federation, 1985), shown in Figure 7. According to the figure, both the Control and Experimental digesters generally performed to the expected V S S reduction percentages at the corresponding temperature x sludge age values, except the for the 1st month of Run 3 of methanol injection. The V S S reduction was much lower than it should be for that time period. This meant there were reasons, other than low sludge age, for the poor solids destruction performance. 58 There are several process variables that need be satisfied to ensure good aerobic digestion such as p H , mixing, and temperature range. The most significant however, is the oxygen requirement for the destruction of cell mass. Without the necessary dissolved oxygen in the system, the oxidation of the cell mass cannot be achieved. The sludge biomass which is represented by the empirical equation C 5 H 7 N O 2 , has a hypothetical oxygen requirement of 1.98 pounds to oxidize 1 pound of cell mass. The results of full-scale and pilot-scale studies indicate that the observed oxygen requirements were 1.74 to 2.07 pounds per pound of volatile solids degraded (Lue-Hing et al., 1998). It is a widely accepted fact, that a minimum dissolved oxygen (DO) level of 1.0 mg/L should be maintained in the aerobic digester at all times to ensure effective sludge digestion (Ahlberg et al., 1972). Figure 8 shows the dissolved oxygen levels of the digesters and the corresponding V S S reductions over the duration of this full-scale experiment. Figure 8 clearly shows that the V S S reduction declined sharply as the D O level dropped to under 1.0 mg/L, and recovered once the D O recovered. The low and nearly identical V S S reductions that were displayed by both the Experimental and Control digesters during the 1st month of Run 3 (as shown in Figure 5), can be explained by the insufficient dissolved oxygen in both digesters; this situation appears to have hindered V S S reduction so much that the comparatively longer sludge age of the Control digester did not result in a better V S S reduction than the Experimental digester. Wi th the increase of incoming sludge from both Control and Experimental S B R to the digesters (as shown in Figure 3), most likely the D O was depleted faster than the coarse bubble aerator could supply it, thus resulting in the drop in D O levels. Moreover, given that the sludge age of the S B R shortens as the wasting volume from the S B R increases, the waste sludge being pumped into the digesters was fresher or more biologically active, thereby, utilizing the oxygen more quickly. 59 U > 2 U > £ U « Q W I + t T (q/Sui) U ^ S A X Q pdAfOSs iQ o o o o o o o o o o ON od r-* >/S Tf' fN o" (%) uoipnpaj SSA An increase in incoming solids, that the digester cannot handle, has proved to be detrimental to aerobic digester solids reduction performance. Since methanol increases the production of solids in the bioreactor, methanol injection would amplify the problem of dropping DO levels, thus causing limited solids reduction (in addition to the expected lower solids reduction due to shortened sludge age). 4.4 Nutrient Balance - Influx, Accumulation, and Re-Release Biological nutrient removal (BNR) in wastewater treatment means that nutrients are separated from the influent flow and removed via the waste sludge, except for denitrification. Therefore, sludge digesters of BNR wastewater plants are often facing increased influx of nutrients, especially phosphorus. Careful operation and management of sludge digesters in BNR plants is required, since scenarios where the accumulation of nutrients in digesters can bring about significant amounts of nutrient re-release into the bioreactors via decant and filtrate, and therefore increase the burden on the BNR process'. Therefore, the performance of the digesters in regards to nutrients revolves around the control of nutrient recycle into the flow system. This section summarizes the findings and observations that were made during this study in regards to the influx, accumulation, and re-release of nitrogen and phosphorus in the two aerobic digesters. 4.4.1 Nitrogen Waste sludges from BNR bioreactors generally have their nitrogen contained in the cell matter of microbial organisms formed because of organic carbon removal. (Kempton et al., 1999) This was evident in this study, as the soluble nitrogen (NH 3, N0 2 , and N0 3 ) accounted for less than 1% (in mass) of the total nitrogen entering the digesters via the SBR waste sludges. 62 Figure 9 shows that, in general, the Experimental S B R wasted more total nitrogen to the digester than the Control S B R . The trend of both S B R s look strikingly similar to that of the volume of S B R waste sludge discussed in Section 4.2, except for the trough for the Experimental side in Run 1 and Run 4. The observed peak by both S B R s at Run 3 of methanol injection, indicate that the difference in total nitrogen mass wasted was most likely due to more solids wasted in the Experimental S B R . The claim that the reason for additional nitrogen influx into the Experimental digester was due to the additional S B R sludge wasting was supported by Figure 10, which shows the nitrogen content in the S B R waste sludge biomass. In fact, there was very little difference between the two sludges in terms of nitrogen content except for Run 1 and Run 4 (due to acclimatization, explained later in this section). This suggested that even with the introduction of methanol, there was no apparent change in the composition of the biomass in regards to the synthesis and storage of nitrogen in organic form in biomass. During Run 1 of methanol injection, the Experimental digester received only a limited amount of nitrogen compared to the Control side. Furthermore, this is also reflected in the low nitrogen content in the Experimental S B R waste sludge biomass, as shown in Figure 10. This observation could be the result of the microorganisms in the Experimental S B R acclimatizing to the sudden introduction of methanol into the bioreactor. The acclimatization period is followed by active synthesis by the microorganisms, due to the added carbon in the form of methanol. The observed drop in nitrogen content in biomass in Run 4 coincides with the discontinuation of methanol injection. This could be another shock response to the microorganisms, which had been acclimatized to the presence of methanol and had been utilizing it; as such, the microorganisms faced another condition change with the discontinuation of methanol. With the different amount of nitrogen entering the digesters, the total nitrogen mass and the total nitrogen concentration was monitored, as shown in Figure 11. 63 ( S S A 3 L U / N § U I %) ssBiuoiq I I I J U S J U O . I U S S O J ; ; ^ Figure 11 shows that despite a large discrepancy in total nitrogen influx for the two digesters (as evidenced in Figure 9), the total nitrogen concentration and the total nitrogen mass in the digesters were similar. The total nitrogen concentration was generally somewhat higher for the Experimental digester, with the difference of no more than 16% throughout the experiment. The digesters were limited to a maximum liquid volume capacity of 200 m each and, as the volume of the incoming S B R waste sludge increased, the wasting of the digester sludge had to be increased accordingly. This resulted in both digesters, having similar percent nitrogen content in biomass, while operating under similar volume levels. This explains (according to Figure 11) the effect of methanol addition with respect to inconsequential nitrogen accumulation on the Experimental digester. However, given that the operating sludge ages of the two digesters were different from each other, it was suspected that the type of nitrogen present in the digesters could be different. The notion was proved valid with the examination of the total soluble and organic nitrogen content in the digesters, as shown in Figure 12. The organic nitrogen represents the nitrogen content in the intracellular environment and total soluble nitrogen, consisting of ammonia and nitrate, represents the nitrogen content in the extracellular environment. The total nitrogen content in the sludge biomass ranged from 6 to 11% for both digester sludges, which is close to the 8 to 10% that was reported by Tonkovic (1999). Under aerobic conditions, the organic nitrogen, that is in the cell tissue, is hydrolyzed to ammonia and released into the liquid phase, in a reaction commonly known as ammonification (Tonkovic, 1999). The ammonia in turn, under aerobic conditions, goes through nitrification where nitrifying bacteria such as Nitrosomonas and Nitrobacter transform the ammonia into nitrates. Since both reactions require an aerobic environment, the dissolved oxygen level in the digesters greatly affected whether the nitrogen would be in organic or soluble state. 66 ( T / N %m) J3;sa8i(j ui U 0 I 1 B J 1 U 3 3 U O 3 uoSoj i i f j I B J O X o ^ U c/3 (%) sseuioig ui uagojji^j rejox p p p o p o P ^ o csi oo -^f o ( N ( N — — (N 8J[) sj3jsa§iQ ui uaSojjifyj o | q i i | o § jo SSBJV JBJOX 68 o >< o in to § z U H t t C/3 > x Z U H ^ As illustrated in Figure 8 and discussed in Section 4.3, the dissolved oxygen level was well below the required minimum of 1.0 mg/L during the 2 months of Run 3. Consequently, the soluble nitrogen content was at its lowest among the four runs, as the low dissolved oxygen level in the digesters inhibited the hydrolysis of organic nitrogen to ammonia and the nitrification of the existing ammonia. This point was further supported by the fact that the organic nitrogen content in the digester sludge biomass was at its highest during this time, meaning that the ammonification of organic nitrogen was at its lowest, over the duration of the experiment. During Runs 1 and 2, the Experimental digester contained more soluble nitrogen than the Control, mainly due to a combination of higher dissolved oxygen concentration and longer sludge age. Beginning with Run 3, where both sides received about the same amount of dissolved oxygen, the longer sludge age of the control side was the reason for the higher soluble nitrogen level, since longer sludge age means longer digesting times. The observations made can be again be confirmed with the inspection of the total soluble nitrogen and ammonia nitrogen re-release via supernatant decant and belt press filtrate, for both digesters, as shown on Figure 13. The trends shown in Figure 13 were strikingly similar to that of the total soluble content trends in Figure 12. The Experimental digester recycle contained more nitrogen then the Control digester, mostly due to the greater volume of decanting and sludge wasting of the Experimental side. Denitrification is a biochemical reaction that occurs under anoxic or anaerobic conditions; it reduces nitrates to nitrite and ultimately to gaseous forms of nitrogen, such as N 2 and N 2 0 . Since the digesters were allowed to settle overnight prior to decant and wasting, denitrification could have and did occur in the digesters whenever there was not enough dissolved oxygen, (as summarized in Table 13). During the 2 months of Run 3, there were almost negligible amounts of nitrates found since they were converted into nitrogen gas following denitrification. .69 70 Table 13. Evidence of Denitrification in Experimental Digester Decant and Filtrate Run Proportion of Nitrate in Total Soluble Nitrogen (%) Dissolved Oxygen Level (mg/L) Pre-run 98 6.6 Run 1 90 3.3 Run 2 91 3.9 Run 3 - 1st 3 0.3 Run 3 - 2nd 27 0.9 Run 4 79 2.2 Figure 14 compares the daily incoming influent ammonia level with the ammonia content of decant and filtrate recycle for both the Control and Experimental digesters. Ammonia is the type of nitrogen that is most closely monitored for the incoming influent, due to its toxicity and burden of oxygen demand for nitrification. Results showed that the ammonia concentration in the decant and filtrate ranged from 2 mg/L to 8 mg/L, but the total mass of ammonia recycled did not exceed 0.2 kg per day (while the influent brought in at least 3 kg of ammonia daily). Therefore, nitrogen re-release had virtually no impact on overall plant operations, as far as ammonia nitrogen was concerned. In summary, the addition of methanol into the Experimental S B R brought about several different characteristic changes to the digester and to the digester sludge but the impact of the nitrogen re-release was found to be negligible. 71 o OO OO 72 4.4.2 Phosphorus In the BNR process, phosphorus removed from the liquid phase is incorporated into the waste sludge and thereby removed from the process. Although some of the phosphorus in the waste sludge is used for metabolic purposes, typically 1.5 to 2 percent (dry weight), the bulk of the phosphorus is stored as long chains of polyphosphate (Kempton, 1999). This polyphosphate is unstable under some conditions and may be discharged to the liquid phase under anaerobic conditions or under extended periods of aeration. (Wentzell et al, 1984) Orthophosphate is formed by hydrolysis of the polyphosphate, and is the form of phosphorus that is of most interest environmental engineers since it can be readily used for biological metabolism. (Manahan, 1994) Figure 15 shows the total daily phosphorus entering the digesters via the SBR waste sludge. The general trend appeared to be similar to the SBR sludge wasting volumes, except for the seemingly increased magnitude of the discrepancy of total phosphorus influx into the two digesters. Examination of Figure 16, which illustrates the organic phosphorus content in biomass, showed that the cause of this discrepancy was due to the higher phosphorus content in the Experimental SBR waste sludge; the Experimental digester was receiving greater volumes of waste sludge that contain higher phosphorus content. 73 (^cp/d 3^ 1) siuoqdsoqj | B J O X Suiuioaui ^ I J B Q (SSA %m/d Sui %) ssBiuoiq ain ui juajuoa sruoqdsoiu The phosphorus content in the S B R waste sludge biomass was as high as 3.8% for the Experimental side during Run 2, with the difference between the two sides reaching a maximum of 1.3% during the 1st month of Run 3. Although more work is needed to delineate the phosphorus mass balances between the S B R tanks and the digesters, these results are consistent with the work of Louzeiro (1999), based on the same site, who reported that improved phosphorus uptake was observed in the Experimental S B R receiving methanol. Moreover, the difference in sludge age of the two S B R s could also be, partially responsible for this trend. A s discussed in Section 4.2, the Experimental S B R wasted more sludge than the Control S B R , meaning that the Experimental S B R has a shorter sludge age than the Control S B R . Since the organic phosphorus can be released under extended periods of aeration, the longer sludge age in an aerobic environment of the Control S B R may have prompted more cell lysis and the subsequent release of polyphosphates into the liquid phase therefore resulting in the biomass having less organic phosphorus in the cell mass. This claim is further supported in Figure 17, which shows that the orthophosphate concentration in the Control S B R waste sludge was generally higher than that of the Experimental side, especially in Runs 2 and the 2nd month of Run 3. Figure 18 shows that, despite a large discrepancy in total phosphorus influx for the two digesters, the total phosphorus concentration and the total phosphorus mass observed were similar, very much like the situation with nitrogen described in Section 4.4.1. The digesters were limited to a maximum liquid volume capacity of 200 m 3 each and as the volume of the incoming S B R waste sludge increased, the wasting of the digester sludge had to be increased accordingly. The wasting of the digesters was not consistent, as shown in Figure 19, thus resulting in the maximum total phosphorus being observed at different times; Run 2 for the Experimental and 1st month of Run 3 for the Control. 76 The wasting volume of the digester was affected by several day-to-day variables, such as overnight settling, weather, and time constraints. The effects of wasting are readily seen when Figures 18 and 19 are compared. A n y change in wasting are directly observed in the next run; for instance, the peak of the total phosphorus level in the digester was observed when the previous run period's wasting was low, and as the wasting was increased, the phosphorus level also decreased. Similar to nitrogen, the effect of methanol addition in the S B R appeared to have little effect on phosphorus accumulation in the digester, since it was readily influenced by the digester sludge wasting process. Although the methanol addition in the Experimental S B R eventually resulted in accumulation of phosphorus in the Experimental digester, the increased wasting of digester sludge due to the extra influx of sludge (as shown in Figure 19), resulted in increased volume of decant and filtrate re-entering the flow path, hence more phosphorus. This is shown in Figure 20. Moreover, a trend of increasing phosphorus re-release, and increased differences between the Experimental and Control sides, with increasing methanol dosage, was observed. In Run 4, where the methanol injection had been discontinued, there still remained a substantial amount of extra phosphorus re-release in the Experimental digester compared to the Control. The most likely reason for this is that any change in the main operation is slow to take affect in a sludge digester, due to the time variable involved in plant operation. The differences between the total phosphorus and orthophosphate levels in the digester decant and filtrate (in Figure 20), represent the non-aqueous organic form of phosphorus. Since organic phosphorus is contained within the cell mass, the increase in organic phosphorus in the recycle streams may be due to the increased presence of suspended solids in the digester decant and filtrate. Figure 21 shows the average total suspended solids levels of the recycle streams for both digesters, and it can be seen that, in general, as the suspended solids levels rose, the organic phosphorus level in both Experimental and Control sides rose as well (in Figure 20). These elevated solids levels may be due to sludge bulking during the experiment (discussed in later section). 80 r °° I -00 ( ^ B p / J §JJ) 3SB3|3J-3J j IBJOX (top/fty) 31EJ1HJ p U B J U B D 3 Q J31S3§IQ UI ggX ^ H B G 33BJ3Ay Unlike nitrogen, which can be removed from the flow cycle via denitrification, phosphorus cannot readily be removed from the liquid cycle. Therefore, the re-release data of the Experimental digester could be, and was used, to develop a plausible quantitative relationship between methanol dosage and the increase in phosphorus re-release. On the assumption that there was no increase in phosphorus re-release with no methanol addition, Figure 22 indicates that approximately 1 litre of additional methanol added per day in this system would escalate the phosphorus re-release by 2 %. These results are very preliminary and it should be noted that they are not based on a detailed analysis and data acquisition that would be possible from more a controlled setting like bench-scale experiments. Figure 23 contrasts the daily, incoming, influent orthophosphate level with the orthophosphate content of the decant and filtrate recycle, for both the Control and Experimental digesters. Orthophosphate is the form of phosphorus that is monitored most closely in the wastewater treatment because of difficulty in treatment and removal, as well as its potential for causing eutrophication in the receiving waters. Results showed that the combined total mass of orthophosphate in the decant and filtrate ranged from 0.03 to 0.44 kg per day, while the influent sewage ranged from 1.1 to 2.5 kg daily. Over the duration of the experiment, the highest amount of orthophosphate being recycled back into the S B R (0.44 kg/day) was equivalent to 28% of the orthophosphate in the influent entering the S B R at the same time. Furthermore, considering that the lowest influent orthophosphate load was 1.1 kg per day, the worst-case scenario would be that the Experimental digester decant and filtrate recycle could increase the orthophosphate load into the S B R by 39%, depending on exact methanol dosages. 83 Such a significant concentration of orthophosphate re-entering the S B R could easily impose extra burden on the B N R capacity of the bioreactor. A likely reason for part of the scenario is the overnight settling period prior to decanting and wasting. The digester could have become anoxic, especially in Run 3, when the dissolved oxygen levels were low; consequently, the organic phosphorus would be released into the liquid phase and hydrolyzed into orthophosphate. In summary, the addition of methanol into the Experimental S B R brought about a significant increase in phosphorus loading into the digester and resulted in substantial increase in the phosphorus re-entering the S B R , via decant and filtrate, due to re-release emanating from the digestion process itself. However, i f the re-release could be eliminated, the apparent phosphorus removal in the S B R with methanol would be greater. 4.5 Sludge Stabili ty Disposal of sludge has recently been a focal point for new ideas and practices because of the large cost that is associated with transport and disposal of the treated sludge. Application of the treated sludge to agricultural has been looked upon favorably as a method to dispose of the sludge and also alleviate the overall cost of sludge disposal. The biosolids can be beneficial for agricultural use, since it usually contains both nutrients and organic matter. However, the biosolids are generally odorous, contain significant levels of pathogens and are attractive to vectors (Farrell et al., 1996). Due to the potential risk of public health and safety, the presence of pathogens and the potential for disease carrying vectors are unacceptable conditions that cannot be ignored. Therefore, these biosolids must be stabilized before they can be used for land application. Many extensive research studies point to the specific oxygen uptake rate (SOUR) as the main barometer for sludge stability for aerobic sludge digestion processes; the general 86 understanding is that the S O U R should not exceed 1.5 mgO2 /1000mgVSShr for the sludge to be considered biologically stable (Eikum et al, 1977). The US E P A ' s (1993) 40 C F R Part 503 standards for the management and disposal of sewage sludge guidelines suggest several options to stabilize the sludge, to satisfy the E P A ' s strict rules regarding the application of sewage sludge. These are summarized in Table 14 (US E P A , 1993). Table 14. US EPA Sludge Stabilization Options Option Description V S S Reduction Achieve minimum volatile suspended solids (VSS) reduction of 38% Specific Oxygen Uptake Rate ( S O U R ) In aerobic digestion processes only, the S O U R must be equal or less than 1.5 m g O 2 / 1 0 0 0 m g V S S h r at 20 ° C p H p H of sludge is raised to 12 or higher for 24 hours In this full-scale experiment, the S O U R and the V S S reduction were used to evaluate the sludge stabilizing performance of the Experimental and Control digesters. Since the final product of sludge digestion would be considered for land application or other purposes, the digester waste sludge was sampled and tested, as detailed in Section 3.2.3.6. Over the course of this study, the liquid temperature of the digesters rose steadily and performing the S O U R test on a sample at a consistent temperature became difficult; every week, the sample temperature was different from the previous week, unlike the controlled settings one might find in a bench-scale experiment. The samples were place in a basin of controlled temperature, but the sample temperature could not be adjusted in time while the sample was being aerated, in preparation for the test. Since the test involves measuring the biological activity of the sample, the test must be performed as early as possible from the time of sampling. Therefore, the constant temperature setting was "sacrificed" and the data were later standardized to a constant temperature of 20 °C. A modified version of the Arrhenius equation was used to adjust the S O U R values, based on numerous findings and recommendations (Eikum et al., 1977): 87 SOUR Tl = 0 TX-T2 SOUR Tl Where, S O U R T i = Specific oxygen uptake rate at T i ; S O U R T 2 = Specific oxygen uptake rate at T2; and 0 = Temperature sensitivity coefficient. A temperature coefficient, 0 = 1.05, for digestion of waste activated sludge, was utilized; this was recommended based on a review of the data of several investigators (Grady et al., 1980). Figure 24 shows the S O U R of the Experimental and Control digester waste sludges, adjusted to 20 °C by the means of the above equation. Figure 24 results showed that, at no time over the duration of the experiment, was the recommended S O U R of 1.5 mgO 2 /1000mgVSShr being met. Generally, the Experimental digester waste sludge had higher S O U R values (peaking at 7.8 mgO 2 /1000mgVSShr), than the Control side sample. This could be an effect of sludge age, since the Experimental side sludge ages were generally shorter, thus having less time to digest and stabilize the sludge. Upon examination of the sludge age (see Figure 6 in Section 4.3), it was determined that the S O U R trend mirrored (in reverse) that of the sludge age. Ahlberg et al. (1972) also determined that a correlation existed between the S O U R and the sludge age. Figure 25 and 26 plots the S O U R against the sludge age. According to the power model, from Figures 25 and 26, (determined using Microsoft Excel 2000 regression analyzer) the observed rate of S O U R decrease, with increasing sludge age, for Control and Experimental digesters, were the following: Control: y = 33.572 x y = 25.271x -0.5636 R 2 = 0.1521 Experimental: -0 .4947 R 2 = 0.5976 y = S O U R (mgO 2/1000mgVSShr) x = Sludge Age (days) 88 D OZ © (J«SSA8ui000l /ZO S u i ) HflOS o 1— o o s as so < v M T3 O CN OX) cs OX) O S3 O U © DO ^ 0* t3 j3 55 TS S3 « o o> s-S3 OX) D oz: © (jqssASuiooox/zo^ui) aaos A s the R squared value indicates, especially for the Control, the correlation between the model and the data of the actual S O U R and sludge ages, is low. A s such, these results should only serve as a rough estimate of the actual situation, given the many variables present in this experiment and the lack of data points. However, it is clear from these results that the longer the sludge age, the more stable the sludge w i l l become, confirming previously published data. A s mentioned earlier, a 38% V S S reduction would satisfy the qualifications of a stabilized sludge, according to the U S E P A guidelines. Figure 5, in Section 4.3, shows that during Run 1, 2nd month of Run 3, and Run 4, the Experimental digester satisfied this criteria, and similarly during Run 2, 2nd month of Run3, and Run 4 for the Control digester. Despite the fact that the corresponding S O U R values all exceeded the maximum allowable, the sludge samples that satisfied the V S S reduction can be classified as "stable", since one of the U S E P A criteria is met. A possible explanation for this is that, at 38% V S S reduction, much of the biodegradable material in the sludge is degraded to lower activity forms. The remaining biodegradable material degrades so slowly that vectors are not necessarily attracted to the sewage sludge (US E P A , 1989). However, it should be noted that there is a great source of error for this S O U R test, because of the variables that were present, such as weather, overnight settling characteristics, inconsistent D O in the digester during the experiment, and varying incoming S B R waste sludge characteristics. Furthermore, the S O U R results obtained were inconsistent and would require more detailed analysis in a "more-controlled" environment, to determine a definitive conclusion as to whether the sludges can be classified as stable or not. Wi th respect to the effect of methanol injection on sludge stability, the methanol addition to the Experimental S B R , causing shorter sludge ages in the Experimental digester than the Control digester, was found to decrease the stability of the resulting digested sludge. 92 4.6 Sludge Dewaterabili ty Handling and disposal of biosolids is a high cost operation. Wi th the reduction of water content in the sludge, the volume of sludge would be reduced and there would be great economic benefits. A l so on a practical level, dewatered sludges are easier to handle, than thickened or liquid sludge, as they may be shoveled, and transported by belt conveyors (Metcalf and Eddy, 1991). The District of Kent Wastewater Treatment Plant employs a belt filter press with polymer injection. The belt filter press is considered the most versatile mechanical dewatering mechanical unit, as it has been proven to be effective for almost all types of municipal wastewater sludge (Metcalf and Eddy, 1991). The capillary suction time (CST) test is being widely used to determine the optimum dose of polymer to be added during the dewatering process. Recent studies showed that the C S T test results could be used as a measure of sludge dewaterability (Vesilind, 1988). It is generally understood that, the lower the C S T , the better the sludge dewaterability (Triton Electronics Ltd. , 1998). Figure 27 plots the C S T test results for both the Experimental and Control digester waste sludges. In general, the C S T values decreased as the experiment progressed, and there was very little variance in results between the Experimental and Control digester waste sludges. This result suggested that the dewaterability of sludge was not affected by the methanol addition in the Experimental digester. Moreover, there was no observable trend that may be caused by sludge age difference, dissolved oxygen, or nutrient content. In order to discover the cause of the distinct trend of C S T or sludge dewaterability results, the equation that forms the basis of the C S T test was examined (Vesilind, 1988): 93 (D3S) 3UIJX H O i p n g ^JBHldB3 Where, % = Filterability constant; <|) = Instrument constant (does not vary with sludge sample); \i = Viscosity of filtrate (kgs 2 /m 2 ) ; C = Concentration of solids (mg/L); and t = Capillary suction time (CST) (seconds). According to the above equation, the variables that affect C S T are concentration of solids and viscosity. The increase in total suspended solids of the digester waste sludges was found to increase the C S T result, as indicated in Figures 28. However, based on this set of results, this influence on the C S T time seemed to be limited. A s the linear model indicates in Figure 28, the correlation between C S T and the total suspended solids was very low. Studies have shown that filtrate viscosity is about 10% higher than that of water. (Vesilind, 1988) Since viscosity decreases linearly with liquid temperature, it was both logical and practical, to compare the C S T result with the sample temperatures taken at the time of testing. In support of the use of temperature as a substitute variable for viscosity, Vesi l ind also studied the relationship of C S T and temperature in his study. Figures 29 and 30 show the plot of C S T against temperature, for both the Experimental and Control digester waste sludges. Based on the results shown on Figures 28 and 29, it was apparent that fluid viscosity, influenced by temperature, was the major variable that affected sludge dewaterability. The Experimental and Control digester waste sludges showed a decrease of 4.2 and 4.4 C S T seconds per degree of liquid temperature increase, respectively. 95 s © H U - D s-Q a w> 13 j 3 CO = o 0> OX) jg co #c 2 " © co T3 c a 3 co "<3 o H «t-i o «a S 3 W oo <u 9 CUD (T/8ui) spijos papuadsns psjox u es OJ a OX) j 3 35 S-cu cu OX) © © U © S3 SM £ cu Q S3 © eu S-S3 C3 S-CU a CU H © • r<\ CU SN 3 OX) (spuoDas) XSD 97 These values also were almost identical to the decrease of 4.3 C S T seconds per degree of liquid temperature that was found based in Vesil ind's (1988) study. Although the sludge characteristics of Vesil ind's study were no doubt different, the closeness of fit enhanced the validity of the results obtained in this experiment. In addition, the linear models for both Experimental and Control sides showed good correlation, further adding to the confidence in these results; it also demonstrated that the viscosity of filtrate appeared to have a greater influence on C S T than the total suspended solids of the sludge. There are actually many other important factors that influence sludge dewaterability, including particle size distribution, bound water content, fluid density, particle surface potential, particle surface area, sludge yield strength, and sludge permeability (Lotito et al, 1993). However, such variables are impractical and difficult to control in field studies, and as such, are not discussed in this work. In summary, the methanol injection into the Experimental S B R appeared to have no effect on the dewaterability of the digested sludge. However, it was found that the dewaterability was affected by both the total suspended solids concentration of the digested sludge, and by the temperature of the digested sludge (that controls the filtrate viscosity), with the latter found to be more influential. 4.7 General Performance Issues This section includes discussions on some of the performance issues regarding the sludge processing and handling units of the treatment plant, that were observed prior to and during the study; these include sludge bulking, possible V F A production, and seasonal variations. The District of Kent Wastewater Treatment Plant, which receives municipal wastewater from approximately 5000 people in Agassiz, B C , Canada, is expected to receive additional wastewater from regional population growth, the Kent federal 99 maximum security prison and a nearby First Nations community. In response to the expected future increase in flow, any recommendations that would be beneficial to the successful operation of the sludge processing and handling units are detailed in this section. 4.7.1 Sludge Bulking Sludge bulking is one of the most common problems that occurs in wastewater treatment plants. Poorly settling sludge and foaming cause operational problems in treatment plants. Often these problems result in the plant operators having to inspect and re-assess the current operating methods of the treatment plant and make minor adjustments in the operating system. Although the greatest impact of sludge bulking is felt in the aeration basins, the sludge digesters are equally, i f not more, distressed by the bulking sludge. Poor settling wi l l directly result in low solids concentration in the digester waste sludge, and excess solids are often carried back into the flow system via the supernatant decant recycle. Filamentous microorganisms have been identified as the cause of sludge bulking. Ironically, filamentous microorganisms are actually very beneficial to the sludge settling process, when they are " in balance" with the floc-forming microorganisms. The filamentous microorganisms provide a rigid support network upon which the floc-formers can attach and grow into a large size floe, with good settling characteristics. In addition, the outward branching of the filamentous microorganisms allows the floe to filter out fine particulates, allowing bridging or joining of floes, which further improves the decant quality. When the filamentous microorganisms begin to dominate and outgrow the floc-formers within the mixed liquor, they extend far beyond the boundaries of the floe and into the bulk liquid of the mixed liquor. A s a result, they interfere with the compaction of the floe, and this interference produces a sludge with a poor settling rate. The resulting floe or sludge is more buoyant and is termed filamentous bulking sludge (Water Pollution Control Federation, 1990). 100 Over the course of the experiment, sludge bulking was observed at various times, but was most prominent during the 2nd month of Run 3 for both the Control and Experimental aerobic sludge digesters. Table 15 lists the digester decant solids concentrations, which can be used as an indicator for sludge bulking. The discussion of specific causes and effect can be found later in this section. The Sludge Volume Index (SVI) test is probably a better indicator of sludge bulking, since it measures the settling of the sludge; however, since these samples were based on high solids digester sludges, the settling was very slow and nearly undetectable during the test duration specified in Standard Methods ( A P H A , 1996), in a column apparatus. Therefore, the S V I test was not undertaken. Table 15. Suspended Solids Levels of Digester Decant Run # TSS Control Decant (mg/L) TSS Experimental Decant (mg/L) Pre-Run 43 415 1 803 500 2 984 315 3 - 1 st month 653 545 3 - 2nd month 992 688 4 344 674 Filamentous microorganisms include a variety of bacteria, actinomycetes, and fungi that add up to twenty-nine commonly found microorganisms in activated-sludge processes (Water Pollution Control Federation, 1990). The digester sludges were sampled, fixed, Gram stained and examined with a microscope to identify which filamentous microorganism manifested themselves in the digester sludges. 101 Figure 31. lOx Photograph of Control Digester Sludge (June 16,1999 - Run 3) Figure 32. 40x Photograph of Control Digester Sludge (June 16,1999 - Run 3) 1 0 2 Figure 33. lOx Photograph of Experimental Digester Sludge (June 15,1999 - Run 3) The filamentous microorganism that was most abundant was identified as Nocardia spp., based on the microscopic observation of its Gram stain color, shape, length, and other characteristics. Figures 31 to 34 show microscopic photographs of the sludge samples taken during the 2nd month of Run 3 from both digesters, that demonstrated fdamentous microorganism manifestation. A s these figures indicate, the filamentous microorganisms dominate the floes, with branching and web-like structures readily observed. Many possible causes exist that could lead to blooming of filamentous microorganisms; wastewater characteristics, such as fluctuations in flow and strength, p H , temperature, staleness, nutrient content, and the nature of the waste components; design limitations such as air supply capacity, clarifier design, return sludge pumping capacity limitations, short circuiting, and poor mixing; operational causes such as low D O in the aeration basin, insufficient nutrients, widely varying organic waste loading, overaeration, low food-to-microorganism (F/M) ratio, and an insufficient soluble B O D 5 gradient (Metcalf and Eddy, 1991). In this study, where the process units of interest were the aerobic sludge digesters, the most likely causes of sludge bulking were suspected to be sludge age and D O level in the digester. Table 16 summarizes the average sludge age and D O levels of the two digesters for each run, along with the F / M ratios in the SBRs . 104 Table 16. Possible Causes of Sludge Bulking in Digesters Run # Control Experimental Sludge Age (days) Average Dissolved Oxygen (mg/L) F/M Ratio in SBR Sludge Age (days) Average Dissolved Oxygen (mg/L) F/M Ratio in SBR Pre-Run 31 1.2 0.21 NA 6.6 0.24 1 36 2.2 0.23 65 3.3 0.21 2 47 3.5 0.15 30 3.9 0.14 3 - 1 st month 23 1.5 0.32 15 1.1 0.28 3 - 2nd month 30 0.9 0.31 19 0.6 0.33 4 41 1.8 0.30 36 1.7 0.34 The excessive growth of filamentous microorganisms occurs when the growth rate of the filamentous microorganisms is higher than that of the floc-forming microorganisms (Sezgin, 1978). Normally, the floc-forming bacteria have the competitive advantage over the filamentous microorganisms, because they are able to rapidly absorb and store the food material or utilize the dissolved oxygen. At low concentrations of soluble substrates, nutrients or DO, however, filamentous microorganisms have an advantage, due to their high surface-to-volume ratios. Once the filaments have grown into the bulk liquid, beyond the boundaries of the floc-forming bacteria, they are exposed to an even higher substrate levels; as such, the filamentous microorganisms will grow even faster in the bulk liquid and result in bulking sludge in a very short time (Water Pollution Control Federation, 1990). From Tables 15 and 16, both the effect of long sludge ages and periodic low DO levels can be seen. For runs with low DO levels, sludge bulking was observed and resulted in a high solids level in the decant. In addition, long sludge ages coupled, with low F/M ratios, also resulted in poor quality digester decant. 105 Actions to counter the bulking of sludge, include increasing the D O level and reducing the sludge age. For this experiment, the low D O levels during Run 3 in the two digesters, as discussed in Section 4.3, were due to the fact that a high influx of S B R waste sludges existed, but the aerators were unable to raise the D O sufficiently to maintain an adequate level. Metcalf and Eddy (1991) suggested that the D O should be raised to a minimum of 2.0 mg/L, under normal loading conditions, or the installation of new aerators might be required. It is understood that the longer the sludge age, the lower the level of substrate available ( F / M ratio). In this study, where the process unit of interest was the aerobic digester, the feed was not the influent but rather the S B R waste sludge. This means that the available substrate is minimal, as the floc-forming bacteria are already in endogenous decay mode. In addition, the digesters carry a greater concentration of solids in the tank compared to the SBRs . Therefore, even though the F / M ratio of the SBRs was used to interpret the amount of substrate that the incoming S B R waste sludge had been exposed to, the F / M ratio in the digesters would be much lower and thus susceptible to sludge bulking. The sludge age can be reduced by either increasing the wasting rate of the digester sludge or reducing the operating volume of the digesters. However, it is important not to compromise the solids reduction performance of the digesters by shortening the sludge age excessively, since solids reduction is the paramount goal of sludge digesters. A balance of the sludge age and D O of the digesters that does not yield any bulking sludge, and yet achieve good solids reduction performance, is the major challenge that the District of Kent Wastewater Treatment Plant currently faces. 106 4.7.2 Volatile Fatty Acids (VFAs) Production For enhanced biological phosphorus removal (EBPR), the presence of volatile fatty acids, or VFAs, are beneficial as the VFAs are quickly consumed by the phosphorus removing bacteria prior to phosphorus uptake (Barnard et al., 1998). VFAs are fermentation products of normally-occurring, facultative organisms under anaerobic conditions. Therefore, many biological nutrient removal (BNR) wastewater treatment plants have an anaerobic zone, in order to promote the production of VFAs. The District of Kent Wastewater Treatment Plant is equipped with two SBRs that can be programmed to operate to suit the needs for BNR to occur. However, without a separate dedicated anaerobic zone, the prospect of EBPR is limited. The SBR becomes anoxic, then anaerobic, during the settling and wasting phase, but the duration is considered insufficient to produce any significant amount of VFAs. The two aerobic digesters at the District of Kent Wastewater Treatment Plant have a holding capacity of 200 m3 each and are aerated continuously. The aeration is stopped a day before the digester is scheduled to decant, waste and dewatered the next day. If the dissolved oxygen (DO) level in the digesters drops sufficiently to allow the digesters to become anaerobic, and given that the digesters are kept idle overnight to allow for sludge settling, it was anticipated that VFA production could indeed occur overnight in the digesters. The VFAs would then enter the SBRs via the recycled decant, for the use by phosphorus removing bacteria. Starting from June 2nd, the decant and filtrate from both digesters were sampled each time the digesters were kept idle overnight, for the scheduled decanting, wasting and dewatering. The samples were filtered and preserved as detailed in Section 3.2.3.5. The results of the VFA detection test for acetic, propionic, and butyric acid, along with the dissolved oxygen level of the digester following overnight settling, are summarized in Table 17. 107 Table 17. Volatile Fatty Acid Concentration in Digester Decant and Filtrate Control Experimental Date Dissolved Oxygen (mg/L) Decant (mg/L) Filtrate (mg/L) Dissolved Oxygen (mg/L) Decant (mg/L) Filtrate (mg/L) 06/03/99 0.0 1 20 (Acetic Acid) No sample 0.0 N / D 2 No sample 06/09/99 1.19 N/D N/D 0.52 N/D N/D 06/16/99 2.83 N/D N/D 0.00 N/D N/D 06/23/99 0.28 N/D N/D 0.40 N/D N/D 06/30/99 0.66 N/D N/D 0.00 N/D N/D 07/06/99 0.11 N/D N/D 0.36 N/D N/D 07/14/99 0.04 N/D N/D 0.11 N/D N/D Dissolved oxygen level in digester after overnight settling 2 N / D - None detected. Detection limit is 1 ppm The results overwhelmingly indicated that the digesters were not suitable and could not be relied upon for V F A production under the existing mode of operation, since all of the samples contained minimal V F A s , except for one sample. The most likely reason for this was that the length of time under anaerobic conditions was insufficient for the fermentation and accumulation of V F A s in the digester. The air was usually turned off at 4:30pm and the decanting started at 8:30am the next morning; this means that the microorganisms had approximately sixteen hours to initially, wait until the digester became anaerobic, then start producing V F A s . It was obvious from the results that this time was insufficient for any significant amount of V F A s to accumulate in the digester; in some cases, the digester never did become anaerobic. Moreover, the low ambient temperature at night, although likely minor, could also negatively affect the fermentation process. The one exception, sample of Control digester decant taken on June 3 r d 1999, contained acetic acid, since there had been a previous wasting just two days earlier on 108 June 1 s t. This operational rarity resulted in three consecutive days of no D O in the digester, thus providing the necessary detention time for the V F A s to be produced and accumulate in the digester. In summary, the overnight idle period of the aerobic digesters did not prove to be suitable for any significant V F A production. A separate tank, under anaerobic conditions, would be required to produce the V F A s necessary to achieve E B P R , unless sufficient V F A s are present in the raw sewage itself, o a sustained basis. 4.7.3 Seasonal Variations The District of Kent Wastewater Treatment Plant in Agassiz, B C , Canada, is located in the upper Fraser Val ley and closely adjacent to the Fraser River into which the final effluent is discharged. There are four distinct seasons of spring, summer, fall and winter. The two most influential variables to the aerobic digester, that would vary seasonally, would be digester temperature and incoming influent flow, with temperature, probably being the most influential on biological activity. The flow affects all units of the treatment plant, and an increase in influent flow would also increase the S B R wasting to the digesters; therefore, the sludge age of the digesters would be affected. A s mentioned in Section 3.2.2, on-line monitoring of the digester temperature and p H was established and the data was recorded by on-site, data logger over the duration of the experiment. Figure 35 plots the average daily temperatures of the Experimental and Control digesters that were obtained via the on-line monitoring system. Figure 36 plots the average weekly p H of the digesters and they are compared with the sludge age of the digesters. Figure 35 was plotted to verify the claim that depending on the buffering capacity of the system, the p H of aerobic digesters w i l l drop to a low value (-5.5) at long detention times (Metcalf and Eddy, 1991). 109 ON ON So O l oo •tf c 3 D C Mi CO c tr CM C D tr c tr c tr & i V 4* ft^ 4i 41 ON —. oo O N 3 \ ON ON —. N O ON CTN O C N O O q o d (3) duiax CA S-I CU -t— CA CU © U i 3 i -cu a s cu H *e3 Q cu 01) CS u cu >• ^ iri CU u 3 DX) 110 (step) 3§v 3 § P n I S Figure 35 shows that both digesters experienced similar temperature values as expected. Moreover, a gradual increase in temperature was observed typical for a large body of water, caused by the gradual increase of ambient temperature as the experiment progresses from spring (March 1999) into summer (August 1999). Figure 36 indicates that, as the p H fluctuated in a trend analogous to the sludge age, that is, an inverse relationship seemed to exist, with p H dropping as sludge age increased. The reason cited for this includes the increased presence of nitrate ions in solution (hence increased acid production), and the lowered buffering capacity, due to air stripping (Metcalf and Eddy, 1991). A s noted in Section 4.3, the influx volume of S B R waste sludge is a critical parameter for the solids reduction in the digesters. The S B R wasting is manually controlled by the plant operators, but ultimately, it is a function of the incoming influent flow rate. Historical data from 1997, plotted in Figure 37, shows the seasonal trends of the influent flow. Since the District of Kent Wastewater Treatment Plant services both the stormwater and municipal wastewater from Agassiz, B C , a clear trend of high precipitation in the summer and winter months was observed in the influent flow. The sludge digesters would be most strained during the summer months; however, the higher operating temperatures in the summer may alleviate the pressure somewhat. A n examination of the 1997 data shows summer influent flows that are well over double that of 1999, a sign of extremely heavy rain or local flooding, due to the excessive runoff from the melting snowpack. In these cases of extremely high flow, the digesters theoretically would be under enormous pressure to perform. However, traditionally, high influent flow results in weak sewage, as the domestic wastewater flow is diluted by the incoming stormwater. Therefore, most of the flow could be removed as effluent from the S B R and the increase in sludge wasting could be manageable for the aerobic digesters. This, however, is purely hypothetical, especially since the situation would be different every year. 112 (XBp/£Ul) M O J J X]UJlIOJ\[ 38BJ3AV 4.7.4 Scale- Up Issues The District of Kent Wastewater Treatment Plant was originally designed for a maximum flow of 5000 m 3/day. A s Figure 37 showed, the incoming flow, with the exception of the summer 1997, did not exceed 2600 m 3 per day in over three years. The Kent federal maximum-security prison uses a wastewater treatment facility that includes a comminutor (muffin monster), an extended aeration basin (oxidation ditch design), and a clarifier. Currently, the treatment system is extremely over loaded and the treatment facilities at the site have proved to be inadequate. However, due to its status as a federal operation, the plant is not subject to the same strict regulations as other regional wastewater treatment plants, such as the District of Kent Wastewater Treatment Plant. With the number of inmates increasing, the federal government has approached the District of Kent to service the prison wastewater at the District of Kent Wastewater Treatment Plant. In addition to the prison wastewater, the District of Kent is also negotiating the servicing of wastewater from the nearby Sea Bi rd Island First Nations community, as well as the Village of Harrison Hot Springs (as a future possibility). Table 18 shows the characteristics of the Kent federal maximum-security prison wastewater from November 1998 to January 1999. Table 18 indicates that the influent characteristics do not differ greatly with what is considered medium-strength municipal wastewater. Traditionally, prison wastewater is stronger than municipal wastewater, due to the lower water consumption rate of the inmates; since these results were only from a three-month window, it is not entirely representative of the expected characteristics of the wastewater. The Kent prison average influent flow rate for the three-months was 533 m 3/day, and over the same three-months, the District of Kent Wastewater Treatment Plant received 1481 m 3/day in 1997 to 1999. This would result in a 36% increase of incoming influent flow, and this figure does not include the additional source of wastewater from the First Nations community; thus, the actual flow in the future could be even higher. Even 114 with these preliminary results, there is no question that the future incoming flow into the treatment plant is anticipated to be of higher daily volume and strength. Table 18. Kent Prison Influent Profile (Nov 98 - Jan 99) Parameter Influent Flow (m3/day) 533 Temperature (°C) 19.5 pH 6.6 TSS (mg/L) 242 BOD (mg/L) 271 COD (mg/L) 646 Ammonia NH 3 (mg/L) 21.1 Phosphorus P (mg/L) 9.9 TKN (mg/L) 36 With the expected increase in incoming flow, the sludge processing and disposal units at the treatment plant will have to be ready to handle the increased load. In this study, the methanol injection into the Experimental SBR increased the incoming waste sludge into the Experimental digester considerably, compared to the Control digester. Although the difference would be less than that compared to the anticipated increase of sludge due to the additional influent, the observations made in this full-scale experiment helped identify some areas in the sludge processing units that would need to be improved. The most problematic area was the dissolved oxygen level in the digesters. The sludge digesters were not operated at the minimum recommended DO level of 1.0 mg/L during Run 3 of this experiment, which produced the most influx of SBR waste sludge. The result 115 was poor solids reduction, poor sludge stability, and sludge bulking. It was suspected that the dissolved oxygen was depleted faster than the aerators could supply. Considering that solids reduction is the most important performance parameter of sludge digestion units, it is paramount that proper measures be taken to ensure that the dissolved oxygen level is not compromised. The existing aerators for the sludge digesters at the District of Kent Wastewater Treatment Plant are coarse-bubble diffusers. The coarse bubble diffusers produce larger bubbles of air, compared to the fine bubble diffusers, and consequently have a lower oxygen transfer efficiency due to the limited total surface area of the bubbles. Indeed, the coarse bubble diffusers are cheaper and require less maintenance, but the cost savings from improved solids reduction could alleviate or even exceed the incurred cost of fine bubble diffusers. Moreover, the current aeration system at the treatment plant is such that a single air pump source supplies air to 3 sinks; the two aerobic digesters, the headworks at the front end of the plant, and the septage holding tank. Depending on the usage of the other two facilities, the air transferred to the digesters w i l l vary even i f the air valves at the digesters have not been adjusted. More studies should be done to determine the feasibility of upgrading the coarse bubble diffusers. The sludge dewatering unit, which was the belt filter press, w i l l most likely have to be scaled-up as well . The speed at which the belt filter press is operated, is especially important, since it ultimately decides how much sludge can be wasted from the digester. During the experiment, dewatering was usually done in one day, since the consequences of having the air off in the digester for more than one day was too costly; time was the limiting factor in how much sludge was wasted and dewatered. Wi th the increase in sludge influx into the digester, the amount of digested sludge, in turn, was also increased, since the liquid holding capacity of the digester was limited. If, in the future, the digesters are required to waste considerably more than what was observed during the experiment, the belt filter press may need to operate at a higher speed to dewater more, in a limited time. The added polymer dose would also have to be optimized, to ensure the best possible 116 results. The polymer dose was last determined in M a y 1998, by the polymer supplier (Nuttall, 1999). However, considering the fact that at different sludge dewaterability results (CST test), different optimal doses of polymer exist, the polymer dose would have to be updated from time to time to make certain that the dewatering unit is operating at optimum conditions. A well-performing, sludge dewatering unit w i l l reduce the cost of transporting the sludge cake significantly. The sludge cake is conveyed via a screw conveyor into a sludge storage building. This warehouse has an open-ended wall for the sludge transport truck to enter and operate. The environment inside the warehouse was humid and odorous due to the presence of the sludge cake. If the roof of the warehouse can be replaced by a transparent roof, perhaps made of material such as a thick slab of plexiglass, the incoming sunlight could pay surprising dividends by increasing the evaporation of water in the stored sludge cake. Although this would be of limited usefulness during the winter, in the summer when Agassiz, B C , Canada, consistently experiences average temperatures of over 25 degrees, this effect could be significant. The drier sludge cake would be less odorous but more importantly, it would be lighter. The potential cost savings could render the improvement in the warehouse as beneficial, even i f it is only useful for the summer months. The District of Kent Wastewater Treatment Plant is facing a potential scale-up situation due to the expected increase in flow. The SBRs are designed to handle up to 5000 m 3 per day of incoming flow, but observations made during this full-scale experiment have cast doubts as to whether the sludge digestion and dewatering units are up to the challenge. Several improvements in equipment, and operating strategy could prepare the sludge handling units for the future. St i l l , further inquiries into the feasibility of such improvements should be carried out before any action is taken. 117 5 CONCLUSIONS AND RECOMMENDATIONS 5.1 Summary and Conclusions The primary goal of this research was to determine whether the use of methanol in the Sequencing Batch Reactors (SBRs) affects the aerobic digester performance. This was achieved by monitoring four key performance areas of sludge digestion: solids reduction performance; nutrient balance - influx, accumulation, and re-release; stability of digested sludge; and dewaterability of digested sludge. The secondary objective of this goal was to investigate general performance issues surrounding the biosolids handling units of the District of Kent Wastewater Treatment Plant in response to the expected increase in incoming flow in the future. This section summarizes the findings and the conclusions that were made based on the findings. 5.7./ Effect of Methanol in SBR on Aerobic Digester Performance The methanol injection into the Experimental S B R for biological nutrient removal, resulted in a significant increase in the solids level in the SBRs , and the sludge wasting to the Experimental digester was increased accordingly, by the plant operators. As the methanol dosage increased from 27 L/day to 54 L/day and finally to 81 L/day, the sludge wasting rate escalated as well . Based on the observed increase in wasting to the Experimental digester, compared to the Control digester, the following relationship was determined: 1 L/day of Methanol added = 0.39 kg of additional solids produced in the S B R This, however, was not based on a detailed analysis, and it should only serve as a preliminary guideline. 118 A s the flow increased, the so did the wasting into both digesters, with a subsequent reduction in sludge age. A s the wasting increased, so did the strain on the coarse bubble diffusers to supply sufficient oxygen; and in Run 3, where the methanol dosage was at 81 L/day, the dissolved oxygen level in the digesters became inadequate for proper aerobic digestion. The resulting volatile suspended solids reduction was also poor, at 22 and 20% for the Control and Experimental digesters, respectively. Since the addition of methanol increases the incoming sludge volume into the digester, methanol negatively affected the solids reduction performance of the aerobic digester. The total nitrogen influx into the Experimental digester paralleled that of the volume of S B R waste sludge, since the nitrogen content in the biomass of the two S B R waste sludges was very similar, except when methanol was first introduced and when it was discontinued. The nitrogen accumulation in digesters was not a concern, since the total nitrogen present in the digesters was largely controllable by the adjusting the wasting of digester sludge. The increased solids resulting in short sludge ages and inducing low dissolved oxygen levels, was found to limit hydrolysis of organic nitrogen into soluble forms. Although methanol addition in the Experimental S B R brought about several characteristic changes to the digester and to the digester sludge, the impact of the nitrogen re-release was found to be negligible, compared to the nitrogen in the influent flow. The Experimental S B R waste sludges contained much more organic phosphorus in the biomass than the Control S B R waste sludge, but contained less orthophosphate in the bulk liquid. Overall, the total phosphorus influx was greater in the Experimental digester, due to the increased wasting from the Experimental S B R . The accumulation of total phosphorus was also controlled largely by adjusting the digester sludge wasting. The increase in incoming total phosphorus in the Experimental digester, resulted in significant amounts of total phosphorus being re-released and re-entering the S B R , via the decant and filtrate. The amount of re-release was high enough to raise the incoming total phosphorus load into the S B R by up to 39%. Based on the results of the phosphorus re-release, the following relationship was found: 119 1 L/day of Methanol added = 2% increase of Total Phosphorus re-release The addition of methanol into the Experimental S B R brought about a significant increase in phosphorus loading into the digester and resulted in substantial increase in the phosphorus re-entering the S B R ; thereby, increasing the burden on the S B R for phosphorus removal, and possibly affecting the overall plant phosphorus removal. The specific oxygen uptake rate (SOUR) and volatile suspended solids reduction was used to determine sludge stability. The specific oxygen uptake rate (SOUR) showed that it was affected by the sludge age of the samples. N o sludge sample taken during the course of the experiment met the required S O U R value of 1.5 mgO 2 /1000mgVSShr, as recommended by E P A guidelines (1993). In regards to the effect of methanol injection on the sludge stability, the methanol addition to the Experimental S B R (causing shorter sludge ages in the Experimental digester than the Control digester) was found to worsen the stability of the resulting digested sludge. The capillary suction time (CST) test was used exclusively to determine sludge dewaterability. The methanol injection into the Experimental S B R appeared to have no effect on the dewaterability of the digested sludge, as samples from both sides had very similar C S T results; there was no observable trend that could be linked to the methanol addition. It was found that the dewaterability was affected by the total suspended solids concentration of the digested sludge, but affected even more by the temperature of the digested sludge, that controls the filtrate viscosity. The following relationship between C S T and temperature was found: 1 °C increase in sludge temperature = 4.2 seconds decrease in C S T (Experimental) In summary, methanol addition in the S B R bioreactor resulted in an increase in solids production in the S B R and, therefore, increased the sludge wasting to the aerobic digester. There were no direct effects of the methanol addition to the sludge digestion process, but the increase in sludge wasting to the digester resulted in a detrimental affect on all parts of 120 the digester performance that were monitored, except nitrogen re-release and sludge dewaterability. 5.1.2 General Performance Issues Filamentous microorganisms caused sludge bulking in both digesters during the experiment. Nocardia spp, was identified as the filamentous microorganism that manifested itself in the digester sludge. Although there are many factors that could cause the excessive growth of filamentous microorganisms, three were identified: low dissolved oxygen in digester; long sludge age; and low food-to-microorganism ratio ( F / M ratio). Although controlling sludge bulking is important, changing operations to reduce bulking at the expense of solids reduction is not recommended. The challenge to plant operators is to fine tune the mode of operation to achieve both goals. The overnight settling period, before digester sludge is wasted, was initially anticipated to be suitable for volatile fatty acids ( V F A s ) production in the aerobic digester. The results overwhelmingly indicated that the digesters were not suitable and could not be relied upon for V F A production under the existing mode of operation, as all the samples (except one) contained minimal V F A s . The fermentative organisms came under seriously detrimental conditions during the digesters' aeration cycle, and the non-aeration period was not long enough for a viable colony of fermentative organisms to be re-established. A separate tank under anaerobic conditions would be required to produce the V F A s necessary to achieve E B P R . The sludge digester, as noted in this experiment, was greatly affected by the incoming solids loading. Moreover, the S B R wasting was dependant on the incoming influent flow. Therefore, the summer and winter months, where the flows are at highest, would present the greatest challenge for the plant operators to maintain an adequate level of digester performance. 121 The District of Kent Wastewater Treatment Plant could receive additional influent from the Kent federal maximum security prison and the nearby Sea B i r d Island First Nations community in the near future. The future incoming flow into the treatment plant is anticipated to be of higher daily volume and strength. The coarse bubble diffuser installed in both digesters may not be adequate to handle the increased flow, based on the findings of this research. The belt filter press may have to operate at higher speeds in order to waste more digested sludge to accommodate the expected increase in S B R waste sludge. A plan to utilize the heat and sunlight to evaporate the water content in the sludge cake may be of benefit in terms of cost savings. 122 5.2 Recommendations for Further Research Continued research is recommended on the following topics: • Study of the oxygen transfer efficiency of the existing coarse bubble aerators -determine whether it is adequate or is in need of upgrade to fine bubble aerators; • Study of the kinetic rates, including the endogenous decay coefficient and temperature sensitivity coefficient - to assist in future retrofit design of digesters; • Study of the sludge bulking in the digesters - a more extensive study into the exact causes and fine tuning the operation, as to not sacrifice solids reduction while still achieving minimal sludge bulking; • Study of the V S S reduction performance in the digesters at lower sludge ages -most of this experiment was done with the digester sludge ages between 20 to 60 days; Wi th the expected increase in influent flow, the digesters may not be able to hold such lengthy sludge ages in the future; • Study of the optimum digester operating characteristics to minimize phosphorus re-release - optimum D O , and optimum sludge age, to minimize the re-release while maintaining decent V S S reduction • Study of the optimum polymer dosage required for sludge dewatering at cold and warm temperatures - polymer dosage should be adjusted to produce the best quality sludge cake, since a correlation between C S T and sample temperature was found. This research is the biosolids management component of an on-going study at the University of British Columbia to explore the potential of methanol for Biological Nutrient Removal (BNR) purposes. Additional research projects are now underway at this plant to further understand the relationship between methanol, biological nutrient removal, and other areas of wastewater treatment. 123 6 REFERENCES Agerbaek, M . L . , and Keiding, K . , (1993) On the origin of specific resistance to filtration, Water Science and Technology, 28:1:159-168 Ahlberg, N .R . , and Boyko, B.I . (1972) Evaluation and design of aerobic digesters, Journal of Water Pollution Control Federation, 44:634 Alleman, J.E. 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C C C c c CO CO CO 3 3 3 3 3 -p ~p -p -p -p -p CM CO * oo CO CO no oo CO CO d CO A" Z ' " _ CM CO z CM IS- IS- IS- Is- O CO 3 < tf-CD oo o> OJ 00 op op op op Op 3 3 3 3 3 ~3 —> ~3 "3 uo ri tf- •A 00 CM CM OO CD O) CD 00 Cp op Op Op OO 3 3 3 3 3 -3 -3 ~3 -p -p uo ri tf- CO T— CM CM CO APPENDIX A-5: Temperature and Dissolved Oxygen in Digesters PRE-RUN Date 5- Jan-99 6- Jan-99 7- Jan-99 8- Jan-99 9- Jan-99 10- Jan-99 11- Jan-99 12- Jan-99 13- Jan-99 14- Jan-99 15- Jan-99 16- Jan-99 17- Jan-99 18- Jan-99 19- Jan-99 20- Jan-99 AVERAGE Temperature (C) Cont Exp 11.6 13.6 13.4 12.9 12.7 12.6 13.8 13.4 14.2 14.6 13.6 13.3 (page 1 DO (mg/L) Cont Exp 0.3 0.9 Of 2) PRE-RUN 11.2 11.9 11.5 12.3 12.3 13.2 10.1 10.5 13.5 13.2 13.3 12.1 0.5 0.1 0.5 1.1 1.4 0.9 3.0 0.3 1.8 0.8 1.9 1.0 0.0 1.9 2.2 0.4 0.2 4.3 8.3 4.1 0.9 0.5 0.3 2.0 Temperature (C) Cont Exp 24- Feb-99 25- Feb-99 26- Feb-99 27- Feb-99 28- Feb-99 1- Mar-99 2- Mar-99 3- Mar-99 4- Mar-99 5- Mar-99 6- Mar-99 7- Mar-99 8- Mar-99 9- Mar-99 10- Mar-99 11- Mar-99 AVERAGE 13.4 14.1 13.6 13.8 11.4 13.2 13.5 12.6 13.2 13.7 14.7 13.4 11.4 13.2 12.3 11.9 11.3 11.8 12.2 11.5 10.8 12.7 11.4 11.9 DO (mg/L) Cont 1.7 0.9 0.8 0.2 3.1 0.2 1.2 1.2 2.2 0.4 1.4 1.2 Exp 3.4 12.6 8.7 11.7 13.9 1.1 6.1 5.2 2.9 7.1 0.3 6.6 RUN1 11-Mar-99 14.7 11.4 1.4 0.3 12-Mar-99 12.5 12.3 0.3 5.6 13-Mar-99 14-Mar-99 15-Mar-99 16-Mar-99 13.6 13.1 0.9 5.6 17-Mar-99 13.1 12.9 0.5 6.1 18-Mar-99 13.0 12.9 6.3 0.5 19-Mar-99 13.0 11.6 4.5 2.4 20-Mar-99 21-Mar-99 22-Mar-99 14.1 14.0 0.7 1.0 23-Mar-99 24-Mar-99 15.3 13.8 1.5 0.7 25-Mar-99 15.6 15.1 0.2 8.4 26-Mar-99 14.4 14.5 0.4 0.6 27-Mar-99 28-Mar-99 29-Mar-99 30-Mar-99 13.5 14.0 2.1 8.4 31-Mar-99 14.1 13.6 3.6 0.5 1-Apr-99 13.1 13.9 5.2 1.0 2-Apr-99 3- Apr-99 4- Apr-99 5- Apr-99 6-Apr-99 14.9 14.6 1.0 6.0 7-Apr-99 14.8 13.6 4.8 2.2 AVERAGE 14.0 13.4 2.2 3.3 147 RUN2 8-Apr-99 9-Apr-99 14.7 14.8 0.0 10.3 10-Apr-99 11-Apr-99 12-Apr-99 15.7 15.4 1.9 5.0 13-Apr-99 15.0 15.2 0.6 4.7 14-Apr-99 15.2 14.5 5.8 1.9 15-Apr-99 15.4 15.4 1.3 1.1 16-Apr-99 17-Apr-99 18-Apr-99 19-Apr-99 17.0 16.3 0.5 2.9 20-Apr-99 17.8 15.1 5.5 1.9 21-Apr-99 16.6 16.2 1.4 7.3 22-Apr-99 16.4 15.3 6.7 0.3 23-Apr-99 16.9 4.4 24-Apr-99 25-Apr-99 26-Apr-99 17.2 16.1 3.8 0.1 27-Apr-99 17.0 7.9 28-Apr-99 29-Apr-99 16.3 15.3 7.3 0.9 30-Apr-99 16.7 17.0 1.8 6.3 1-May-99 2-May-99 3-May-99 16.9 16.9 0.3 6.2 4-May-99 16.6 17.0 9.1 1.6 5-May-99 17.2 16.9 8.9 0.6 6-May-99 15.5 16.0 0.6 6.3 AVERAGE 16.3 16.0 3.5 3.9 A P P E N D I X A -5: T e m p e r a t u r e a n d D i s s o l v e d O x y g e n in D i g e s t e r s (page 2 of 2) RUN3-1 RUN3-2 7-May-99 16.0 16.0 1.5 7.4 6-Jun-99 8-May-99 7-Jun-99 17.9 18.2 0.5 0.0 9-May-99 8-Jun-99 18.6 18.5 2.9 0.5 10-May-99 16.1 16.0 7.0 2.6 9-Jun-99 19.2 18.1 1.2 0.1 11-May-99 15.7 15.7 0.1 8.0 10-Jun-99 19.1 18.5 2.1 0.0 12-May-99 16.3 16.4 4.9 0.1 11-Jun-99 18.6 18.4 2.1 0.6 13-May-99 17.3 17.0 5.5 0.1 12-Jun-99 14-May-99 17.3 16.3 0.2 1.6 13-Jun-99 15-May-99 14-Jun-99 19.8 19.9 0.0 0.0 16-May-99 15-Jun-99 20.6 20.3 0.7 0.0 17-May-99 18.0 18.3 5.2 0.0 16-Jun-99 19.9 19.7 2.8 0.6 18-May-99 17.1 18.5 0.0 0.0 17-Jun-99 20.3 19.3 0.1 0.0 19-May-99 17.7 17.3 3.2 0.3 18-Jun-99 19.4 20.1 0.4 0.0 20-May-99 17.3 17.0 0.4 0.0 19-Jun-99 21-May-99 17.3 16.9 1.0 1.0 20-Jun-99 1 st15days 2.6 1.9 1st15days 1.3 0.2 22-May-99 21-Jun-99 19.5 19.2 0.0 1.6 23-May-99 22-Jun-99 19.4 18.9 0.9 0.4 24-May-99 23-Jun-99 20.1 19.2 0.3 0.2 25-May-99 17.9 17.2 0.2 0.4 24-Jun-99 19.6 18.0 0.0 0.0 26-May-99 16.7 17.2 1.9 0.2 25-Jun-99 19.2 18.7 0.0 0.0 27-May-99 18.1 17.7 0.0 0.1 26-Jun-99 28-May-99 17.9 17.7 0.1 0.3 27-Jun-99 29-May-99 28-Jun-99 19.2 18.7 0.0 0.0 30-May-99 29-Jun-99 18.5 18.0 1.6 0.0 31-May-99 18.5 18.0 0.4 1.3 30-Jun-99 19.5 18.9 0.7 4.6 1-Jun-99 17.8 17.7 0.0 0.0 1-Jul-99 2-Jun-99 18.8 17.7 0.0 0.0 2-Jul-99 3-Jun-99 17.7 17.5 0.0 0.0 3-Jul-99 4-Jun-99 18.1 17.5 0.7 0.0 4-Jul-99 5-Jun-99 5-Jul-99 18.4 18.1 1.5 2nd 15days 0.4 0.3 2nd 15days 0.4 0.9 1 month 17.4 17.2 1.5 1.1 1 month 19.3 18.9 0.9 0.6 RUN4 RUN4 6-Jul-99 19.3 18.9 3.1 0.4 21-Jul-99 7-Jul-99 18.6 18.5 0.1 4.2 22-Jul-99 20.9 20.2 2.7 0.0 8-Jul-99 19.0 17.4 0.0 1.0 23-Jul-99 9-Jul-99 19.9 19.4 0.3 0.0 24-Jul-99 IO-Jul-99 25-Jul-99 11-Jul-99 26-Jul-99 19.0 19.2 1.9 2.6 12-Jul-99 20.8 20.4 4.7 0.6 27-Jul-99 19.4 19.1 4.4 0.7 13-Jul-99 20.8 20.2 1.7 0.0 28-Jul-99 19.7 19.2 0.9 3.3 14-Jul-99 20.7 19.8 0.0 0.1 29-Jul-99 20.1 19.2 0.1 2.7 15-JU.I-99 20.1 19.3 2.4 0.0 30-Jul-99 19.4 19.6 0.5 3.1 16-Jul-99 20.4 19.9 0.6 0.5 31-Jul-99 17-Jul-99 1-Aug-99 18-Jul-99 2-Aug-99 19-Jul-99 20.5 20.3 0.7 4.0 3-Aug-99 21.2 20.4 3.0 1.0 20-Jul-99 20.9 20.0 4.6 1.9 4-Aug-99 21.0 21.1 1.6 3.9 1st15days 1.7 1.1 2nd 15days 1.9 2.2 1 month 20.1 19.6 1.8 1.7 148 CO CD c CA CO CC CO (/) c (A tz "5 .2 •<-.fc CD ~T CO Q-m x o z LU 0 . 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