The Pennsylvania State University. The Graduate School. Department of Agricultural and Biological Engineering

Transcription

1 The Pennsylvania State University The Graduate School Department of Agricultural and Biological Engineering COPROCESSING WATER TREATMENT RESIDUALS AND BIOSOLIDS FOR PHOSPHORUS MANAGEMENT IN WASTEWATER TREATMENT A Dissertation in Agricultural and Biological Engineering by Malcolm Taylor Submitted in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy May 2013

2 i The dissertation of Malcolm Taylor was reviewed and approved* by the following: Herschel A. Elliott Professor of Agricultural Engineering Dissertation Advisor Chair of Committee Brian Dempsey Professor of Civil Engineering James M. Hamlett Associate Professor of Agricultural Engineering Robert D. Shannon Associate Professor of Agricultural Engineering Paul Heinemann Professor of Agricultural Engineering Head of the Department of Agricultural and Biological Engineering *Signatures are on file in the Graduate School

3 iii ABSTRACT Treatment and disposal of water treatment residuals (WTR) continues to be an area of increased focus in the drinking water industry due to increasingly stringent local, state, and federal regulations. One area with great potential to reduce operating costs is the beneficial use of WTR, defined here as an alternative to conventional disposal methods (e.g., landfilling, discharge to sanitary sewers). WTR have been shown to reduce the soluble P content in amended soils and reduce dissolved P in runoff water and leachate. Whereas the majority of research investigating the P-binding capacity of WTR has focused on the co-application of WTR and biosolids, limited research has been completed on the effects of blending and conditioning WTR and biosolids prior to dewatering (hereafter referred to as coprocessing). Moreover, of the coprocessing studies that have been completed, minimal attention has been given to the effects of coprocessing on the soluble-p content of the resulting biosolids. The purpose of this research was therefore to investigate coprocessing as an additional strategy for managing residuals produced in water and wastewater treatment under evolving nutrient management policies. The effects on sludge dewaterability and P concentrations in water generated during dewatering (reject water) were investigated. A unique aspect of this effort was a quantitative evaluation of how coprocessing influences the environmental lability of P in biosolids land application. Laboratory experiments were conducted to evaluate the effects of coprocessing on: (1) biosolids dewaterability and polymer dose, (2) the water extractable P (WEP) concentration in dewatered biosolids, and (3) the total P concentration in reject water from dewatering operations. Results were then validated through a case study involving full-scale water and wastewater treatment facilities.

4 iv Laboratory studies investigated dewaterability (using capillary suction time, CST) of combined alum residuals (Al-WTR) and anaerobically digested biosolids at various blending ratios (BR), defined as the mass ratio of WTR to biosolids on a dry solids basis. Without polymer addition, the CST was 160 s for a BR of 0.75 compared to 355 s for the biosolids alone. The operational polymer dose (OPD), defined as the polymer dose yielding CST of 20 s, was reduced from 20.6 g kg -1 dry solids for the biosolids alone to 16.3 and 12.6 g kg -1 when BR was 0.75 and 1.5, respectively. Precipitated Al hydrous oxides in the WTR likely caused flocculation of the biosolids particles through heterocoagulation or charge neutralization. The solids contents of the blended materials and biosolids at their respective OPDs were not statistically different (α = 0.05) following dewatering by a belt-filter press (BFP). It was concluded that coprocessing with Al- WTR improved biosolids dewaterability and reduced polymer dosage. The impact of coprocessing on the partitioning of P during dewatering and the environmental lability of biosolids-p as measured by water extractable P (WEP) were also evaluated in lab-scale studies. The reject water total P (TP) content from dewatering biosolids alone (250 mg L -1 ) was progressively reduced with increasing BR reaching 60 mg L -1 for a BR = 1.5. The dewatered cake (~20% solids) WEP varied inversely with BR, dropping from 7.6 g kg -1 (biosolids only) to ~0.2 g kg -1 for BR = 1.5. Polymer addition resulted in lower reject liquid TP for all conditions tested, indicating the cationic polyelectrolyte contributed to P binding. A case study was completed at the Media, PA, wastewater treatment plant in an attempt to validate findings from the laboratory experiments. Al-WTR were coprocessed with digested anaerobic biosolids at BR = 1.3. The solids content of the BFP cake was increased from 19.4% to 22.0% and the polymer dosage was reduced from 19.3 to 14.1 g kg -1. Biosolids WEP was reduced from 0.1 to 0.01 mg kg -1, illustrating the great potential of coprocessing to reduce WEP in biosolids. Total P in the BFP filtrate was reduced from to 6.34 mg L -1. Because the

5 v filtrate is recycled to the headworks, coprocessing results in a net reduction in P loading to the WWTP. Phytoavailable P (measured as Mehlich-3 P) was not reduced by application of the coprocessed material. Trace element content in the WTR is comparable to, or less than, the trace element content in the biosolids, therefore coprocessing will not prohibit land application or beneficial use with respect to trace metal concentrations. Aluminum (Al) in coprocessed cake was shown to be highly insoluble, therefore land application should not induce Al phytotoxicity or enhance Al movement to surface or ground waters. For this facility, coprocessing was established as a viable method for processing WTR and biosolids with several distinct benefits. The potential for using WTR for P control applications appears promising and the demand for a product with this capability is steadily increasing as concerns about surface water degradation and P-based nutrient management continue to receive intense focus. In practice, the extent of these benefits may be limited by factors such as: variability in the chemical and physical properties of WTR and biosolids, (2) infrastructure and economic considerations, and (3) quantities of WTR produced relative to the amount of biosolids generated by a municipality. Nonetheless the results of this study indicate that coprocessing provides an additional strategy for managing residuals produced in water and wastewater treatment under evolving nutrient management policies.

6 vi TABLE OF CONTENTS LIST OF FIGURES... ix LIST OF TABLES... xi ACKNOWLEDGEMENTS... xii Chapter 1 Introduction Background and Justification Research Goals and Objectives Dissertation Structure Data Management Definitions Factors for Conversion Statistical Methods... 7 Chapter 2 Literature Review Introduction Significance of Phosphorus Pollution The Role of Phosphorus in Eutrophication Sources of Phosphorus Mechanisms of Phosphorus Loss Phosphorus Management and Risk Assessment Control of Soluble Phosphorus The Phosphorus Index Phosphorus Source Coefficients Water Treatment Residuals Disposal of WTR Land Application of WTR Influence of WTR on Soil Phosphorus Solubility and Phytoavailability Factors Influencing the P-binding Capacity of WTR Blending of WTR with P Sources Coprocessing WTR and Biosolids Effect of Coprocessing on Dewaterability and Polymer Dose Effect of Coprocessing on P Control in Wastewater Treatment Identification of Future Research Needs Chapter 3 Participating Utilities Facility Descriptions and Solids Characterization The City of York; Wastewater Treatment Plant Sludge Processing and Biosolids Generation The York Water Company... 32

7 vii Generation of WTR Materials and Methods Sample Collection Al ox Determination Evaluation of Al ox in WTR Over Time Dewatering and Drying of WTR Samples Results and Discussion Biosolids Characterization WTR Characterization Ammonium Oxalate Extractable Al (Al ox ) in WTR Al ox Concentration in WTR from Three Stages of Processing Measured Over Time Effect of Drying on Al ox Concentration in WTR Summary and Conclusions Chapter 4 Influence of Water Treatment Residuals on Dewaterability of Wastewater Biosolids Introduction Capillary Suction Time Test as an Indicator of Sludge Dewaterability Methods Results and Discussion Characteristics of Biosolids and WTR Influence of Coprocessing on Dewaterability Influence of Coprocessing on OPD and Cake Solids Conclusions Chapter 5 Coprocessing WTR and Biosolids: Impact on Phosphorus Partitioning and Biosolids Recycling Introduction Methods and Materials Materials Coprocessing Procedure Phosphorus Analysis Results and Discussion Prescreening of WTR Samples with Varying Al ox Content Effect of Coprocessing on Total P in Reject Water from Dewatering Effect of Coprocessing on WEP in Dewatered Biosolids Effect of Coprocessing on Phosphorus Partitioning Influence of Nutrient Levels and WEP on Land Application Rates and P- Index Scores Summary and Conclusions Chapter 6 Coprocessing Al-WTR and Anaerobically Digested Sludge for Control of Water Soluble Phosphorus at Media, PA: A Case Study Introduction Facilities Descriptions

8 6.2.1 Ridley Creek Water Treatment Plant Media Wastewater Treatment Plant Materials and Methods Sample Collection and Analysis Belt Filter Press Filtrate Processing and Analysis WEP Testing of Belt-filter Press Cake Samples Results and Discussion Characterization of Liquid WTR and Biosolids Solids Content of Belt Filter Press Cake Phosphorus Content in BFP Filtrate Water Extractable Phosphorus (WEP) in Belt Filter Press Cake Implications of Coprocessing on Land Application Nutrient Based Land Application Rates Phytoavailability of Phosphorus Trace Element Concentrations Aluminum Content Summary and Conclusions Chapter 7 Major Findings; Engineering Implications and Future Research Areas Major Findings Influence of Coprocessing on Dewaterability of Biosolids Influence of Coprocessing on Biosolids WEP and Phosphorus in Reject Water from Dewatering Validation of Results with a Full-Scale Case Study Engineering Implications Maximum Practical Blending Ratios Operational and Economic Considerations Permitting and Regulatory Considerations Potential for Future Research Effects of Aluminum Reactivity on the Dewaterability of Biosolids Integrating Coprocessing and Solids Management Effect of Coprocessing on Wastewater Treatment and Effluent Phosphorus Effect of Coprocessing on Nitrogen in Reject Water Odor Reduction viii

9 ix LIST OF FIGURES Figure 3.1. City of York WWTP; simplified sludge processing schematic Figure 3.2. York water WTP process flow diagram Figure 3.3. Al ox concentration in WTR from three stages of processing Figure 3.4. Effect of dewatering and drying on the Al ox in WTR samples of equal age Figure 4.1. Schematic diagram of the CST apparatus Figure 4.2. Typical results showing the effect of chemical conditioning on biosolids CST Figure 4.3. Capillary suction time for the materials Figure 4.4. Percent solids following dewatering with bench-top belt filter press for 3 BRs conditioned with polymer at the operational polymer dose Figure 5.1. Comparison of dissolved and total P in BFP filtrate samples Figure 5.2. Water extractable phosphorus for a range of coprocessed WTR / biosolids blends Figure 5.3. Total P in reject water for range of coprocessed WTR / biosolids blends dewatered to 10% solids Figure 5.4. Total P in reject water for range of coprocessed WTR / biosolids blends dewatered to 20% solids Figure 5.5. Water extractable phosphorus in dewatered biosolids for a range of coprocessed WTR / biosolids blends dewatered to 20% solids Figure 5.6. Phosphorus mass balance for coprocessing WTR and biosolids at BR = Figure 5.7. Partitioning of Phosphorus for a range of coprocessed WTR / Biosolids Blends with polymer added at the optimal polymer dose Figure 5.8. P-Index score as a function of soil test M 3 P for biosolids and a WTR: biosolids blend Figure 5.9. Influence of WEP on P-index prescribed management for 95-fields in Pennsylvania with active biosolids land application programs Figure 6.1. Ridley Creek Water Treatment Plant flow schematic

10 x Figure 6.2. Media Wastewater Treatment Plant flow schematic Figure 6.3. Ridley Creek WTP and Media WWTP biosolids & residuals handling flow schematic Figure 6.4. BFP cake dryness (% solids) for liquid biosolids feed and WTR / biosolids blend feed Figure 6.5. Effect of coprocessing on total phosphorus in BFP filtrate Figure 6.6. Effect of coprocessing on the water extractable phosphorus in biosolids and WTR / biosolids blend Figure 6.7. Trace element content of biosolids and coprocessed blend as a percentage of PA DEP limits for land application of sewage sludge

11 xi LIST OF TABLES Table 2.1. Pennsylvania P Index Table 3.1. Selected properties for City of York WWTP sludge & biosolids Table 3.2. WTR characteristics for the York Water Company s Grantley Road WTP Table 4.1. Characteristics of biosolids and Al-WTR Table 5.1. Ammonium oxalate extractable aluminum in WTR at various stages of processing Table 5.2. Influence of Nutrient Content and WEP on Land Application Rates and P- Index Values Table 5.3. Pennsylvania P-Index input conditions found in the study area Table 6.1. Dewatering system operating parameters: Media WWTP case study Table 6.2. Characterization of the Ridley Creek WTP and Media WWTP biosolids Table 6.3. Nutrient data and N-based land application rates for biosolids and coprocessed cake Table 6.4. Mehlich-3 P and WEP in soil amended with biosolids and coprocessed cake Table 6.5. Trace elements in biosolids and coprocessed cake for land application Table 6.6. Al concentrations in biosolids and coprocessed material from the Media WWTP and Ridley Creek WTP Table 6.7. Al stress test for soil amended with biosolids and coprocessed cake

12 xii ACKNOWLEDGEMENTS First and foremost I must acknowledge the great influence Dr. Herschel Elliott has been in my life. Without his guidance, patience and understanding this project would not be possible. I have become a dedicated environmental scientist with a strong moral compass, a loving father and proud uncle because of him. I remain truly grateful and humble for his spiritual guidance. I must also thank my committee members Dr. Brian Dempsey, Dr. Jim Hamlet and Dr. Robert Shannon. Also a great thanks to Dr. Robin Brandt and Jim DeWolfe. Thank you for shaping me throughout my career. I am glad that I had the chance to know each of you as a friend as well as a teacher and mentor. You have all influenced me profoundly, and for that I thank you and promise to make you proud. My son Evan has been a great support all along and he continues to be my inspiration to promote an acute awareness of the importance of environmental stewardship for future generations. I appreciate the editorial assistance provided by Emily Haner, who no doubt saved me many a headache. And to my dearest Cindy, I thank you for your patience, support and understanding. Without your support, I am not sure if I would have ever climbed this mountain. The ABE department and staff also receive my greatest appreciation. I remain forever grateful for the fellowship that made this all possible. The administrative staff has been exemplary and for that, I owe a share of my future success. In closing I acknowledge the great contribution of my parents Calum and Catherine Taylor. There is no question that it is by the will of these great people that I have become the man I am. I dedicate this work in remembrance of their great spirit.

13 1 Chapter 1 Introduction 1.1 Background and Justification Water treatment residuals (WTR) are the primary byproduct of the drinking water treatment industry. Processing, handling, dewatering and disposal of WTR constitute a sizable portion of the operating costs of drinking water production. Historically, WTR have been managed as a waste product and were commonly released back into the source water or disposed in a landfill. Years of research, fueled by a desire to minimize byproducts and their impact on the environment, as well as reduce disposal costs, have led to the identification of ways to beneficially use WTR. These include use as soil amendments, additives in brick manufacturing and road subgrades, and most recently, and perhaps most significantly, phosphorus (P) control from non-point sources of pollution. Controlling the loss of P from agricultural soils to surface waters is essential to protecting surface water quality, as excessive concentration of soluble P is the most common source of eutrophication in freshwater systems (Correll, 1998). The substantial capacity of WTR to bind soluble P has been well documented and is attributed to the reactive aluminum (Al) and iron (Fe) hydrous oxides originating from the coagulants used in the drinking water process (Elliott et al., 1990). In the past, the ability of WTR to bind soluble soil nutrients was viewed as a disincentive for agricultural applications (Cornwell et al., 2000). Recently increased concern about surface water degradation from nutrient laden runoff has led to research focusing on using this unique property of WTR to control the loss of P from nutrient rich environments.

14 2 The wastewater industry is also faced with the challenge of managing the byproduct of wastewater treatment (biosolids) in an environmentally suitable and cost effective manner. Approximately 7.1 million dry tons of biosolids are generated each year at more than 16,000 municipal wastewater treatment plants in the U.S. (U.S. EPA, 2009). Biosolids are a valuable source of nitrogen (N) and P for agriculture. Approximately 50% of the biosolids produced are land applied as a source of nutrients and organic matter to the soil (Brandt and Elliott, 2009). Biosolids are typically applied at rates to supply the N requirements of crops. These application rates usually result in excessive P additions due to the imbalance of N and P in biosolids relative to crop needs (Elliott et al., 2005). Many strategies are being used to address this problem and reduce P transport to surface water, including consideration of new best management practices (BMPs). One possible BMP is to remove dissolved P from runoff water and leachate by the land application of P-sorbing materials, such as WTR (Dayton and Basta, 2005). A number of studies have established WTR application as an effective method for reducing P loss from biosolids-amended soils (Elliott et al., 2002; Gallimore et al., 1999). To reduce P loss, WTR have been spread on buffer strips, mixed into the soil and co-applied with the biosolids. Results have shown that the P-binding capacity of WTR varies considerably and is dependent on the chemical and physical characteristics of the WTR. Coagulant type, source water characteristics, dewatering methods and aging have all been shown to influence the P- binding capacity of WTR (DeWolfe, 2006). Previous research has focused on the co-application of WTR and biosolids to offset the soluble P availability from biosolids. However, the impact of blending WTR with biosolids prior to conditioning and dewatering (coprocessing) on water extractable P (WEP) levels in biosolids, dewatering performance and P content in water generated during dewatering (reject water) has not been comprehensively evaluated.

15 3 The potential use of WTR for nutrient control appears promising, and the demand for a product with this capability is steadily increasing as concerns about surface water degradation from nutrient-rich runoff rises. The purpose of this study was therefore to build upon the existing body of knowledge regarding the use of WTR to control P loss from biosolids-amended soils by investigating coprocessing WTR with biosolids. Results of this research will help to further identify conditions that can result in an endproduct that meets agronomic needs while minimizing the amount of P loss to aquatic systems. 1.2 Research Goals and Objectives The primary goal of this study was therefore to investigate coprocessing of WTR and biosolids as an effective solids management option and a potential BMP for the control of water soluble P in wastewater treatment and biosolids management. To accomplish this goal, the following objectives were identified: Conduct a comprehensive review of literature pertaining to the capacity of WTR to bind soluble forms of P and reduce potential for P runoff to surface waters. Examine coprocessing as a method of reducing the water soluble fraction of P in processed WTR / biosolids blends. Investigate the effects of coprocessing on dewatering process performance and associated chemical dosages. Examine the influence of coprocessing on the P concentration in recycle flows from dewatering operations. Validate bench-scale results by coprocessing WTR and biosolids at a full-scale water and wastewater facility.

16 4 1.3 Dissertation Structure To systematically accomplish the identified research objectives, important considerations and implications associated with coprocessing of WTR and biosolids are addressed in separate chapters. The chapters on the experimental work have been developed as stand alone documents, each containing an introductory narrative, materials and methods, results and discussion, and summary and conclusions. Chapter 2 provides a literature review on processing and disposal options for WTR with a focus on the utilization of the material for the control of soluble-p. Characteristics contributing to the capacity of WTR to bind soluble-p are discussed. Management techniques for land-based disposal of biosolids and WTR, and the implications of coprocessing on land application, are also presented to provide a background and justification for this research. To generate experimental data for this research, the City of York wastewater treatment plant (WWTP) and the York Water Company (YWC) graciously provided samples of biosolids and WTR. Chapter 3 provides a description of these facilities including treatment methods and major unit processes. A detailed characterization of the biosolids and WTR samples used in experiments is also presented. Also included in Chapter 3 is characterization of WTR samples from three stages of WTR processing to examine the variation in Al ox content. Lab-scale experiments were then conducted to further quantify the effects of aging and drying on Al ox content in WTR. Chapter 4 addresses the effects of coprocessing on dewatering performance and required conditioning chemical (polymer) dosages through a series of bench-top experiments. The capillary suction time (CST) test and a bench-top belt filter press (BFP) were used as indicators of dewaterability for a range of WTR / biosolids blend ratios. Building upon these results, the polymer dose required to meet dewaterability targets (as measured by % solids) was determined.

17 5 Finally, chemical and physical mechanisms describing the role of WTR in biosolids dewatering are proposed. In Chapter 5 the results from a series of lab-scale experiments are presented to demonstrate the effects of coprocessing on the WEP in processed biosolids and the P concentration in reject water from mechanical dewatering processes. A mass balance is presented to illustrate the phase transitions of P throughout the dewatering process. The effects of polymer addition during dewatering on the resulting biosolids WEP and reject water P is also examined. Implications of coprocessing on the chemical characteristics of the blended materials and the resulting impacts on land application efforts are then discussed. To verify experimentally derived results on a full scale, Chapter 6 presents a case study conducted at the Aqua America WWTP in Media, PA. At this facility, biosolids are coprocessed with WTR from the nearby Ridley Creek WTP and mechanically dewatered with a BFP to produce a blended WTR / biosolids material. The effects of coprocessing on the biosolids WEP content and P concentration in the reject water are presented. Characterization of the dewatered biosolids with and without coprocessing and the resulting impacts on land application efforts are discussed. Chapter 7 discusses implications of this research for the water and wastewater industries and the field of environmental engineering in general. Suggestions for continued and collaborative research in this area are also provided.

18 Data Management Due to the expansive amount of raw data collected throughout this study, only summarized data are provided in tables and figures. A copy of all data can be made available upon request by contacting the author, Malcolm Taylor, at mxt121@psu.edu Definitions The following terms used throughout this dissertation are defined as follows: Sludge: Liquid biosolids: Liquid WTR: Liquid blend: Biosolids cake: Filtrate: Coprocessing: Coprocessed cake: BR: Reject Water: Wastewater solids prior to digestion Wastewater solids post-digestion but pre-dewatering Water treatment residuals prior to dewatering Combined WTR and biosolids streams prior to dewatering Dewatered biosolids Liquid fraction separated during dewatering with a belt filter press Mixing, conditioning and dewatering of a blend of liquid WTR and liquid biosolids Combined WTR and biosolids streams after dewatering [Finished material for land application] WTR / biosolids blend ratio (mass / mass basis) e.g.: A blend of 75 g of WTR and 100 g of biosolids results in a BR value of Water separated and removed from solids during dewatering operations (centrifugation; belt-filter press etc.) and returned to the head of the treatment process.

19 Factors for Conversion Units: Throughout this dissertation metric units are used for all units of measure unless otherwise indicated. For instances in which industrial standards are commonly expressed in English units, these industry standards are used. The following conversion factors to metric units are provide to facilitate interpretation of values. English Multiply by: Metric Equivalent gallons per minute (gpm) liters per minute (L/min) gallons per hour (gph) liters per hour (L/hr) million gallons per day (mgd) x 10 3 cubic meters per day (m 3 /d) pounds per ton (lbs / ton) 0.5 kilogram per metric ton (kg/tonne) pounds per acre (lbs/ac) kilograms per hectare (kg/ha) tons per acre (tons / ac) metric tonnes per hectare (tonne/ha) Pounds per square inch (psi) Pascal (Pa) Statistical Methods Statistical analysis conducted in this study was completed using Minitab statistical software version Unless otherwise indicated, data was evaluated for statistical significance using analysis of variance (ANOVA) procedures at a 95% confidence level. At this level the test statistic alpha (α) equals For WTR and biosolids characterization, samples were collected from the participating utilities as described in Chapter 3. Chemical and physical characteristics of WTR and biosolids can vary widely over time depending on treatment conditions, inflow characteristics and seasonal variations (AWWA, 1990). Consequently, samples were collected on three separate dates to provide a general representation of the sample

20 8 characteristics over the course of the study. Results from the three sampling events were averaged (n = 3) and presented in Table 3.1. Because individual samples are susceptible to biological degradation, experiments were conducted within a predetermined amount of time following sample collection as described in the respective chapters. For dewaterability (Chapter 4) and P partitioning (Chapter 5) experiments and the case study (Chapter 6), subsamples were randomly taken from a composite and independently processed and analyzed as described in the respective methods sections. Variability among subsamples was tested for statistical significance for the various treatments tested. The number of subsamples used in ANOVA analysis (n) is indicated in for each individual experiment. The number of subsamples used is not the same for all experiments due to variables such as available sample volume, time required to conduct a round of experiments and analytical cost.

21 9 References AWWA, Water Quality and Treatment; 4 th Edition. McGraw Hill Inc. New York, New York. Brandt, R.C., and H.A. Elliott Sustaining biosolids recycling under phosphorusbased nutrient management. Water Practice, 3(1):1-14. Cornwell, D.A., R.N. Mutter, and C. Vandermeyden Commercial Application and Marketing of Water Plant Residuals. AWWA Research Foundation, Denver, CO. Correll, D.L The role of phosphorus in eutrophication of receiving waters: A review. J. Environ. Qual., 27: Dayton, E.A., and N.T. Basta Use of drinking water treatment residuals as a potential best management practice to reduce phosphorus risk index scores. J. Environ. Qual., 34: DeWolfe, J.R Water Residuals to Reduce Soil Phosphorus. AWWA Research Foundation, Denver, CO. Elliott, H.A., B.A. Dempsey, D.W. Hamilton and J.R. DeWolfe Land application of water treatment sludges; impact and management. AWWA Research Foundation, Denver, CO. Elliott, H.A., G.A. O Connor, P.Lu, and S. Brinton Influence of water treatment residuals of phosphorus solubility and leaching. J. Environ. Qual., 31: Elliott, H.A., R.C. Brandt, and G.A. O Connor Runoff phosphorus losses from surface applied biosolids. J. Environ. Qual., 34: Gallimore, L.E., N.T. Basta, D.E. Storm, M.E. Payton, R.H. Huhnke, and M.D. Smolen Water treatment residual to reduce nutrients in surface runoff from agricultural land. J. Environ. Qual., 28: U.S. EPA Biosolids: Targeted National Sewage Sludge Survey Report. EPA/822/R- 08/014. Office of Research and Development, Washington, DC.

22 10 Chapter 2 Literature Review 2.1 Introduction Treatment and disposal of water treatment residuals (WTR) continues to be an area of increased focus in the drinking water industry due to increasingly stringent local, state, and federal regulations concerning WTR handling and disposal. Additionally, increasing operating costs and the need to invest a higher portion of revenues into an aging infrastructure have forced the drinking water industry to thoroughly examine cost-saving measures. One area with great potential to reduce operating costs is the beneficial use of WTR. Beneficial use in this context is defined as an alternative to conventional disposal methods (such as landfilling, direct stream discharge, or discharge to the sanitary sewer) that does not cause harm to the environment or threaten human health (Cornwell et al., 2000). One beneficial use of WTR that has received recent attention relies on the capacity for WTR to chemically bind or fix soluble forms of phosphorus (P). Specifically, the targeted beneficial use is improved P management of agricultural and urban lands by controlling the release of P to surface and ground waters. This literature review therefore provides background information on the significance of P pollution including sources of P, mechanisms of loss, impacts of P on aquatic environments and a summary of P risk assessment and management strategies. Also included in this review is a description of WTR characteristics and methods of handling and disposal. A summary of studies examining the P-binding capacity of WTR and the potential for beneficial use of WTR as P management strategy is presented. Finally, the various techniques in which WTR has been used in P-binding studies are summarized along with a description of how the current research

23 11 provides an additional strategy for managing residuals produced in water and wastewater treatment under evolving nutrient management policies. 2.2 Significance of Phosphorus Pollution Controlling the loss of P from agricultural soils and urban areas such as golf courses and lawns is essential to protecting surface water quality, as excessive concentrations of soluble P are the most common source of eutrophication in freshwater aquatic systems (Correll, 1998). In fact, in a 1996 report on water quality in the U.S., eutrophication was identified as the main cause of impaired surface water quality (U.S. EPA, 1996). Eutrophication restricts water use for fisheries, recreation, industry and drinking because of increased growth of undesirable algae and aquatic weeds and the oxygen shortages caused by their death and decomposition. In dramatic incidences, proliferation of toxic bacteria in drinking water supplies can pose serious health risks to animals and humans. Outbreaks of the highly toxic dinoflagellate, Phiesteria piscicida, in the eastern United States and Chesapeake Bay tributaries in particular, have been linked to excess nutrients in affected waters, increasing the public awareness of eutrophication and the need for a solution (Sharpley et al., 2003). 2.3 The Role of Phosphorus in Eutrophication Eutrophication of most fresh water around the world is accelerated by P inputs (Schindler, 1977; Sharpley and Smith, 1994). Although nitrogen (N) and carbon (C) are also essential to the growth of aquatic biota, most attention has focused on P inputs because of the difficulty in controlling the exchange of N and C between the atmosphere and water and the fixation of atmospheric N by some blue-green algae. Therefore, P is often the limiting element,

24 12 and its control is of prime importance in reducing the accelerated eutrophication of fresh waters (Sharpley et al., 2003). Phosphorus in runoff from agricultural land is an important component of non-point source pollution and can accelerate eutrophication of lakes and streams. Long-term land application of P as fertilizer and animal wastes has resulted in elevated levels of soil P in many locations in the U.S. (Daniel et al., 1998). 2.4 Sources of Phosphorus A variety of nutrient sources are added to agricultural and urban lands for plant fertilization including commercial fertilizers, animal manures and processed wastewater sludge (biosolids). Excess P is common because nutrients are generally added to meet plant nitrogen needs often resulting in an overabundance of P. Repeated applications of fertilizers or manures to meet crop N needs leads to a buildup of excess P that potentially becomes susceptible to offsite loss via runoff and leaching (Elliott and O Connor, 2007). 2.5 Mechanisms of Phosphorus Loss Through the mechanisms of rainfall, irrigation, erosion, and subsequent runoff, P can be transported to nearby surface water bodies such as rivers, lakes and reservoirs. When soil P levels are not excessive, up to 90% of the P transported from cropland is bound to soil particles (Sharpley and Beegle, 1999). Under these circumstances conventional erosion control measures can be very effective in reducing the potential for offsite P movement. However, in situations where soil P levels are far in excess of background soil levels and plant needs, offsite transport of soluble P in dissolved forms can contribute a significant portion of the P transported offsite. Although leaching of P to groundwater is generally not a major transport process, areas

25 13 characterized by sandy soils, high permeability and shallow ground water can be particularly susceptible to deep leaching of P (Elliott et at., 2002). A great many studies have shown that soluble P is readily available to algae and thus particularly relevant to water quality degradation (Schindler 1977; Sharpley and Smith, 1994; Elliott et al., 2002; Sharpley et al., 2003). 2.6 Phosphorus Management and Risk Assessment Control of Soluble Phosphorus Recognizing this potential threat to surface water quality, both the federal government and a number of states have developed guidance to control the non-point introduction of P into surface waters. In the early 1990s, a USDA-sponsored program emerged to develop assessment tools for areas with water quality problems, termed the Southern Extension Research Activity Information exchange group 17 (SERA-17). SERA-17 participants recognized that while soil test P is related to P concentrations of runoff, different amounts of P can be lost from sites with similar soil test P results (DeWolfe, 2006). Prior studies showed that site characteristics such as slope and vegetative cover can have an overriding effect on P loss in runoff and erosion. Thus, P loss from fields with similar soil test P values can vary significantly depending on soil management practices and climatic, topographic and agronomic factors that affect runoff and erosion. The need to better account for these interrelated factors in nutrient management plans led to the development of the P-Index.

26 The Phosphorus Index The P-Index is a tool developed to help farmers and land managers to identify sites that have a high risk of P loss to aquatic systems. The index uses source factors and transport factors to estimate the potential for P to be lost offsite as a result of runoff or leaching. Since its introduction, the P-Index has evolved into many versions, but the basic concept of assigning values to source factors and transport factors to quantify the risk of offsite P movement is integral to virtually all versions. Table 2.1 shows the various source and transport factors used in the PA P-Index (Beegle et al., 2006). Input values typical of biosolids land application in PA (Brandt and Elliott, 2009) are inserted into Table 2.1 to illustrate the use of the Index. As the focus on control of non-point sources of P continues to increase, the P-Index has become an important assessment tool and is used in most state-supported nutrient management strategies. Brandt and Elliott (2005) report that 49 states currently use, or are in the processes of developing, P-indices. Although efforts so far have mainly targeted manures and commercial fertilizers, P-based management will likely apply to all land-applied residuals including biosolids. Relatively few states have explicitly addressed biosolids, despite the prevalence of land application for biosolids disposal (Elliott et al., 2005).

27 15 Table 2.1. Pennsylvania P Index (Beegle et al., 2006) SOURCE FACTORS SOIL TEST Mehlich-3 Soil Test P (ppm P) 168 Soil Test Rating = 0.20* Mehlich-3 Soil Test P (ppm P) 34 FERTILIZER-P RATE Fertilizer-P (lb P 2 O 5 /acre) 0 FERTILIZER APPLICATION METHOD 0.2 Placed or injected 2" more deep or 0.4 Incorporated <1 week following application 0.6 Incorporated > 1 week or not incorporated following application in April - October 0.8 Incorporated >1 week or not incorporated following application in Nov. - March 1.0 Surface applied to frozen or snow covered soil 0.2 Fertilizer Rating = Fertilizer Rate x Fertilizer Application Method 0 BIOSOLIDS-P RATE Biosolids P (lb P 2 O 5 /acre) 767 BIOSOLIDS APPLICATION METHOD 0.2 Placed or injected 2" more deep or 0.4 Incorporated <1 week following application 0.6 Incorporated > 1 week or not incorporated following application in April - October 0.8 Incorporated >1 week or not incorporated following application in Nov. - March 1.0 Surface applied to frozen or snow covered soil 0.4 P SOURCE COEFFICIENT 0.4 TRANSPORT FACTORS Biosolids-P Rating = Biosolids-P Rate x Biosolids Application Method x Biosolids-P Availability 123 EROSION Soil Loss (ton/a/yr) 3 RUNOFF POTENTIAL SUBSURFACE DRAINAGE CONTRIBUTING DISTANCE MODIFIED CONNECTIVITY 0 Excessively 0 None 0 > 500 ft ft. Riparian Buffer APPLIES TO DIST < 100 FT BPR biosolids (0.8) and all other biosolids (0.4)...OR determined using Water Extraction Testing 2 Somewhat Excessively 4 Well/Moderately Well 1 Random 6 Somewhat Poorly to 199 ft. OR 350 to 500 ft. 200 to 349 ft. < 100 ft. with 35 ft. buffer Transport Sum = Erosion+Runoff Potential+Subsurface Drainage+Contributing Distance 1.0 Grassed Waterway or None Source Factor Sum 8 Poorly/Very Poorly 2 * Patterened 9 < 100 ft. 1.1 Direct Connection APPLIES TO DIST > 100 FT * OR rapid permeability soil near a stream Transport Sum x Modified Connectivity/ "9" factor does not apply to fields with a 35 ft. buffer receiving manure. P Index Value = 2 x Source x Transport

28 Phosphorus Source Coefficients Biosolids generally exhibit a high total P/N ratio relative to crop requirements. Application rates that are based on plant N needs result in soil total phosphorus (STP) levels beyond optimal for plant growth. However, such soil P enrichment may result in over-prediction of the potential for soluble P unless the solubility of P is considered. State regulations are turning towards P-based nutrient management plans. If the relative solubility of P is not considered when evaluating acceptable application rates, the practice of land applying biosolids could be severely restricted (Elliott and O Conner, 2007). Phosphorus source coefficients (PSC) are used in some P-indices in the calculation of the source factor. The PSC assigns a numerical value to the P availability of a nutrient source compared to inorganic P fertilizer (USDA ARS, 2006). In the Pennsylvania (PA) P-Index, values range from 1.0 for nutrient sources with a high level of P availability (swine manure or commercial fertilizers) to 0.4 for sources that are known to contain relatively less soluble P (biosolids or alum-treated manure). In the PA P-Index, table values or analytical characterization of the water extractable P (WEP) can be used to determine an appropriate PSC value (USDA ARS, 2006). Work by Brandt and Elliott (2005) found that the table PSC value for biosolids of 0.4 frequently resulted in artificially high P-Index scores, because measured PSC values for many types of biosolids were significantly below 0.4. Using a standardized table value for the PSC for biosolids, in many instances, does not accurately reflect the actual soluble-p availability because the solubility, bioavailability, and transport potential of P vary significantly among biosolids types (Brandt et al., 2004). Updating biosolids source coefficients using PSC values based on the actual biosolids applied will usually lower P-Index scores for biosolids-amended sites.

29 Water Treatment Residuals Water treatment residuals (WTR) is the general term used to describe the solids generated during the purification of drinking water. Although the chemical and physical composition of the residuals varies depending on the source water and the treatment processes used, WTR generally contain the solids removed from the source water and the chemical coagulants used in the drinking water treatment process. In some instances, other solids-producing chemicals such as powdered activated carbon (PAC), polymers, or lime may be used in the treatment process and will contribute to the WTR produced. Coagulation of surface waters is by far the most widely used water supply treatment technology and is commonly referred to as conventional water treatment. Residuals produced by coagulation processes make up the majority of the WTR produced by the water industry (Cornwell et al., 1987). In conventional treatment the coagulation process itself generates most of the WTR produced. The most commonly used coagulants are aluminum (Al) and iron (Fe) salts. When added to water, these coagulants form amorphous hydroxides (e.g. Al(OH) 3 in the case of Al) which then precipitate and contribute significantly to the chemical and physical characteristics of the WTR. 2.8 Disposal of WTR Treatment and disposal of WTR continue to receive attention in the drinking water industry due to more stringent local, state and federal regulations concerning residual handling and disposal practices. The discharge of untreated residuals is all but prohibited in most instances under the National Pollutant Discharge Elimination System (NPDES) of the Clean Water Act. Direct discharge to sanitary sewers is also becoming more restrictive due to tougher wastewater

30 18 pretreatment standards and the limited solids-handling capacity of many wastewater treatment plants (WWTP). Disposal in landfills is becoming increasingly expensive and the WTR occupy valuable space that should be reserved for materials that cannot be beneficially reused. As a result, alternative use programs for the beneficial use of WTR are increasingly being investigated throughout the drinking water industry. A variety of beneficial use alternatives for WTR have been investigated and implemented. These include cement and brick manufacturing, turf farming, composting, silvicultural application and landfill cover, among others (ASCE, 1996). One emerging beneficial use of WTR involves the capacity for WTR to be used to chemically bind soluble forms of P. Specifically, the targeted beneficial use is improved P management of agricultural and urban lands by controlling the release of P to surface waters. 2.9 Land Application of WTR When land application was originally investigated as a potential reuse option for WTR, it was thought that the lack of organic components and potentially high metals content in the WTR would inhibit water industry acceptance. A 1996 study of land application of WTR was conducted to establish a scientific basis for land application of WTR. This study examined data regarding the chemical composition and fractionation of WTR, the phytotoxicity of WTR, and the effects of WTR on the physical and nutrient status of soil (Elliott et al., 1990). The results of this study concluded that WTR, by their nature, are usually low-value fertilizers. However, it was found that moderate applications of WTR may improve the physical and chemical condition of soil through several mechanisms including increased organic content, increased moisture retention capacity and increased cation exchange capacity (CEC). Subsequent studies also concluded that processed WTR have physical and chemical characteristics similar to fine-textured

31 19 soils (DeWolfe, 1993), with levels of plant micro- and macronutrients comparable with soil (Elliott and Dempsey, 1991). To address concerns over elevated metals in WTR, Elliott et al. (1990) assessed potential metal mobility under field conditions. Results of this study showed that most heavy metals in WTR are bound in forms not readily released into solution due to their strong adsorption and coprecipitation by Al and Fe oxides originating from the coagulants used during treatment of water. At neutral or alkaline soil ph, less than 6 percent of the total concentration of heavy metals tested was in the exchangeable fraction. Most commercial crops require near-neutral soil ph conditions, which further limits the bioavailability of trace metals Influence of WTR on Soil Phosphorus Solubility and Phytoavailability Phosphorus is a major element in plant nutrition and much research has been conducted on the availability of P in soils following the application of WTR. The addition of WTR to manure or fertilizer-treated soils can effectively reduce both runoff and soil extractable P concentrations because of the relatively high amounts of hydrous oxides of Al and Fe contained in WTR. In soils, these compounds provide a reactive surface area with considerable capacity to fix environmentally labile P (Hamad et al., 1992; Peters and Basta, 1996; Elliott et al., 2002). This characteristic of WTR has been effectively demonstrated in lab and field studies to reduce P bioavailability and runoff potential in soils (Peters and Basta, 1996; Cox et al., 1997; Gallimore et al., 1999). Studies have examined the many factors influencing the P-binding capacity of WTR including the chemical composition of WTR, soil characteristics, application rates and methods, equilibration time, ph, particle size and surface area. Ippolito et al. (2011) provide a detailed review of the many studies examining this property of WTR.

32 20 Historically, P retention was the predominant factor limiting land application of WTR given the great importance of P to optimum plant growth. Greenhouse studies found P deficiencies in grasses were possible when application rates of Al or Fe-based WTR exceeded 10g per kg of soil (g kg -1 ) (Heil and Barbarick, 1989). Cox et al. (1997) demonstrated that surface application of WTR to an acidic soil (ph = 4.4) reduced dry matter yields, tissue P concentrations and P uptake of wheat. Elliott and Singer (1988) also reported P deficiencies in tomato plants with WTR loading at 60 g kg -1. In general, these studies demonstrated that excessive WTR application may result in reduced P concentrations in plant tissues but does not appear to induce other nutrient deficiencies or toxicities (Basta et al., 2000). Whereas in the past, the ability of WTR to bind soluble soil nutrients was viewed as a potential limiting factor for agricultural applications, increased concern about surface water degradation from nutrient laden runoff has led to research focusing on using this unique property of WTR to beneficially control the loss of P from nutrient-rich environments (Cornwell et al., 2000) Factors Influencing the P-binding Capacity of WTR While it has been well established that WTR can be used to bind P and reduce the risk of incidental loss of soluble P to water bodies, different WTR have varying degrees of impact on soil test P and soluble P. Both Dayton et al. (2003) and Novak and Watts (2004) reported that WTR can differ substantially in P-binding capacity because of variations in their oxalateextractable Al and Fe concentrations. In an effort to explain these variations in P-sorption capacities, Dayton et al. (2003) examined the WTR components thought to contribute to P- sorption (reported as P max ). Dayton and Basta (2005) found a significant (r 2 = 0.69, P <0.01) relationship between P max and acid ammonium oxalate-extractable aluminum (Al ox ). Dayton and Basta (2005) proposed an improved method using Al ox to predict WTR P max using a modified acid

33 21 ammonium oxalate-extraction method for Al ox. The improved linear relationship (r 2 = 0.91, P <0.001) provides a useful tool for determining WTR P max without the development of a complete isotherm (Dayton and Basta, 2005). A study completed by DeWolfe (2006) found that the extent of dewatering and the age of the WTR material also impacted the P-binding capacity of the WTR. In general, the reactivity of WTR decreased with increasing WTR age (DeWolfe, 2006). Agyin-Birikorang and O Connor (2009) also reported that freshly-generated Al-WTR samples were potentially more reactive than dewatered WTR samples stockpiled for 6 months or longer Blending of WTR with P Sources Studies that have examined the capacity for WTR to bind soluble P in soils have used a variety of methods to apply the WTR to the soil or test plot. Dayton and Basta (2005) conducted a series of studies in which WTR was applied in buffer strips at the edge of test plots, incorporated into the soil, and co-blended with organic materials (manures and biosolids). Addition of WTR as a buffer strip greatly reduced dissolved runoff P (DRP) and the reduction was related to the P max of the applied material. Similarly, soil incorporation reduced both soluble and Mehlich 3 extractable-p (M 3 P) levels, thereby reducing the potential for offsite migration of DRP. Co-application of WTR with a high P organic material reduced the solubility of P in the organic amendment before land application. In a similar study evaluating the effects of coapplication, Elliott et al. (2002) found that co-blending WTR with manure or biosolids before land application reduced soluble P. DeWolfe (2006) reported that blending WTR with high P soils (> 300 mg kg -1 M 3 P) significantly reduced soluble P at WTR loading rates as low as 10 g kg - 1.

34 22 In each of the studies examining the effects of co-blending on soluble P (Codling et al., 2000; Elliott et al., 2002; Ippolito et al., 1999), WTR was blended with manures, biosolids, or high P solids using dried WTR samples. The effects of percent solids and age of the dried WTR were generally not considered in the evaluations of P-sorbing capacity. Because solids content and age of WTR have been shown to have an effect on the sorption capacity of WTR (DeWolfe, 2006; Agyin-Birikorang and O Connor, 2009) evaluation of the influence of these parameters on the P-fixing capacity of WTR warrants further investigation and study Coprocessing WTR and Biosolids Whereas the majority of research investigating the P-binding capacity of WTR has focused on the co-application of WTR and biosolids, limited research has been completed on the effects of blending WTR with biosolids prior to conditioning and dewatering (referred to herein as coprocessing). Coprocessing is differentiated from discharge of WTR into sanitary sewers in that coprocessing refers to blending of WTR and biosolids post-digestion but prior to dewatering. This has the advantage of centralizing dewatering operations while avoiding increased solids loading and impacts on the activated sludge process associated with WTR discharge to sanitary sewers. The few coprocessing studies that have been conducted have generally focused on one of two areas: (1) Effects of coprocessing on sludge dewaterability and polymer dose and (2) Impacts of coprocessing on P concentrations in reject water from dewatering processes Effect of Coprocessing on Dewaterability and Polymer Dose Several studies have investigated coprocessing and dewatering behaviors of WTR-sludge blends. In 1995, the City of Allentown, PA conducted full-scale pilot studies where alum-based

35 23 WTR were trucked to the WWTP and blended with anaerobically digested solids prior to polymer addition and belt-filter press (BFP) dewatering. Using a 50:50 (mass basis) WTR / biosolids blend, the solids content of the BFP cake was increased to 25.6% compared to 20.8% solids for WWTP sludge alone. In addition, the polymer dose required for dewatering the blend was half that needed for the biosolids alone. Using fixed costs for chemicals, transportation and disposal, it was concluded that coprocessing could reduce processing and disposal costs by approximately 20% (Koplish et al., 1995). Lai and Liu (2004) investigated the coprocessing of alum sludge and waste activated sludge and found that dewaterability, as measured by capillary suction time (CST) and specific resistance to filtration (SRF), improved with increasing fraction of alum sludge. Although the exact mechanism is uncertain, they proposed that the alum sludge acted as a skeleton builder and rendered the mixture more incompressible which, in turn, enhanced dewaterability (Lai and Liu, 2004). Using SRF as a measure of dewaterability, Yang et al. (2007) also reported an improvement in the dewatering characteristics of the coprocessed mixture relative to the digested sludge alone. Results of these studies confirmed previous findings (Hsu and Pipes, 1973) that aluminum hydroxide improved the dewaterability of biological sludge. While aluminum salts have long been used to condition wastewater solids (U.S. EPA, 1987), the fact that Al-WTR have a similar effect on municipal sludges is a more recent discovery. Although studies demonstrate that the optimal WTR-sludge blend ratio and the required polymer dose varies widely depending on the chemical and physical characteristics of the materials and dewatering method, other studies have similarly demonstrated improved dewaterability and reduced polymer demand with an increased fraction of Al-WTR (Koplish et al., 1995; Chen et al., 2007).

36 Effect of Coprocessing on P Control in Wastewater Treatment Several studies have examined the use of WTR to immobilize P in wastewater treatment processes. In a lab-scale equilibration study, Mortula and Gagnon (2007) observed P removal in wastewater effluent of 91-98% using oven dried alum concentrations of 4 to 16 g L -1. In another study in which a sludge bed of alum-based WTR was used in a continuous flow column, ortho-p (PO 3-4 ) was reduced by over 80% in a 30-day period. Even with an extremely high P loading rate (210.5 g PO 3-4 /m 2 d) and a hydraulic loading rate of 2.8 m 3 /m 2 d, the alum sludge bed remained stable and did not reach saturation for over 60 d, confirming the great potential of WTR for immobilization of P in wastewater treatment (Razali et al., 2006; Yang et al., 2006). To examine the role of WTR in immobilizing P in reject water from dewatering processes, several studies have investigated coprocessing as a P management technique. In general, results suggest that coprocessing WTR with biosolids can be highly effective in immobilizing soluble P in reject streams. Yang et al., (2007) reported that using mix ratio of approximately 1:1 (anaerobic digested sludge: alum sludge; mass basis) resulted in a 99% reduction in dissolved P in the reject water through the adsorption of P by alum in the WTR. This finding is particularly important given that recycling of reject water with a high P concentration can lead to reduced P removal efficiency in wastewater treatment, especially those practicing biological P removal (BPR), where the P in reject water can be considerably higher than that of conventional activated sludge systems (Tchobanoglous et al., 2003) Identification of Future Research Needs A great deal of research has been conducted on the P-binding capacity of WTR. The amorphous Al and Fe salts used in the coagulation process retain a fraction of their adsorption

37 25 capacity and therefore provide a nutrient sink for readily-soluble elements such as P. Numerous studies have demonstrated, both in lab and field experiments, that WTR can be a useful agent to reduce the potential for soluble forms of P to be available for transport to surface water bodies. P-binding studies have focused on the application of WTR in buffer strips, soil incorporation, and co-blending with P sources. In all instances, WTR have been shown to reduce the soluble P content in soil extractable-p and dissolved forms of P contained in runoff. Whereas the majority of research investigating the P-binding capacity of WTR has focused on the co-application of WTR and biosolids, limited research has been completed on the effects of coprocessing WTR and biosolids. Moreover, of the few coprocessing studies that have been completed, minimal attention has been given to the effects of coprocessing on the soluble-p content of the resulting biosolids. The potential for using WTR for nutrient control applications appears promising and the demand for a product with this capability is steadily increasing as concerns about surface water degradation and P-based nutrient management continue to receive intense focus. The purpose of this research was therefore to further investigate coprocessing of WTR and biosolids, specifically as it relates to P management strategies. As with previous coprocessing studies, effects on sludge dewaterability and P concentrations in reject water are investigated. This study is differentiated from the existing literature however, in that coprocessing as a technique for reducing the environmental lability of P in land application is investigated.

38 26 References Agyin-Birikorang, S., and G.A. O Connor Aging effects on reactivity of an aluminumbased drinking-water treatment residual as a soil amendment. Science of the Total Environment, 407: ASCE, and AWWA Management of Water Treatment Plant Residuals. New York, New York.: American Society of Civil Engineers and American Water Works Association. Basta, N.T., R.J. Zupancic, and E.A. Dayton Evaluating soil tests to predict bermudagrass growth in drinking water treatment residuals with phosphorus fertilizer. J. Environ. Qual., 29: Beegle, D., R. Bryant, W. Gburek, P. Kleinman, A. Sharpley, and J. Weld The Pennsylvania phosphours index, Version 2. Penn State University: University Park, Pa. Brandt, R.C., H.A. Elliott, and G.A. O Connor Water-extractable phosphorus in biosolids: Implications for land-based recycling. Water Res., 67: Brandt, R.C., and H.A. Elliott Sensitivity analysis of the Pennsylvania phosphorus index for agricultural recycling of municipal biosolids. J. of Soil and Water Conservation, 60(4): Brandt, R.C., and H.A. Elliott Sustaining biosolids recycling under phosphorusbased nutrient management. Water Practice; Water Env. Fed., 3(1):1-14. Chen, Y.C., G. Adams, Z. Erdal, R. Forbes Jr., J.R. Hargreaves, M. Higgins, S. Murthy, J. Novak, J. Witherspoon, W. Toffey The effect of aluminum sulfate addition on production of volatile organic sulfur compounds from anaerobically digested biosolids. Water Practice; Water Env. Fed., 1:1-13. Codling, E.E., R.L. Chaney, and C.L. Mulchi Use of aluminum and iron-rich residues to immobilize phosphorus in poultry litter and litter-amended soils. J. Environ. Qual., 29: Correll, D.L The role of phosphorus in the eutrophication of receiving waters: A review. J. Environ. Qual., 27: Cornwell, D.A., M. Bishop, R. Gould, and C. Vandermeyden Handbook of Practice; Water Treatment Plant Waste Management. Denver, CO.: Awwa Research Foundation. Cornwell, D.A., R.N. Mutter, and C. Vandermeyden Commercial Application and Marketing of Water Plant Residuals. Denver, CO.: Awwa Research Foundation. Cox, A.E., J.J. Camberato, and B.R. Smith Phosphate availability and inorganic transformation in an alum sludge-affected soil. J. Environ. Qual., 26: Daniel, T.C., A.N. Sharpley, and J.L. Lemunyon Agricultural phosphorus and eutrophication: A symposium overview. J. Environ. Qual., 27(2):

39 27 Dayton, E.A., N.T. Basta, C.A. Jakober, and J.A. Hattey Using treatment residuals to reduce phosphorus in agricultural runoff. J. Am. Water Works Assoc., 95: Dayton, E.A., and N.T. Basta Use of drinking water treatment residuals as a potential best management practice to reduce phosphorus risk index scores. J. Environ. Qual., 34: DeWolfe, J.R Analysis, bioavailability and mineralization of nutrients on soils receiving WTR. AWWA/WEF Joint Residuals Management Conference., Phoenix, AZ. 5-8 Dec American Water Works Assoc., Denver, CO. DeWolfe, J.R Water Residuals to Reduce Soil Phosphorus. AWWA Research Foundation, Denver, CO. Elliott, H.A., and L.M. Singer Effect of water treatment sludge on growth and elemental composition of tomato (Lycopersicon esculentum) shoots. Commun. Soil Science Plant Anal., 19: Elliott, H.A., B.A. Dempsey, D.W. Hamilton, and J.R. DeWolfe Land application of water treatment sludges; impact and management. Am. Water Works Assoc. Res. Foundation, Denver, CO. Elliott, H.A., and B.A. Dempsey Agronomic effects of land application of water treatment sludges. J. Am. Water Works Assoc., 84: Elliott, H.A., G.A. O Connor, P.Lu, and S. Brinton Influence of water treatment residuals of phosphorus solubility and leaching. J. Environ. Qual., 31: Elliott, H.A., R.C. Brandt, and G.A. O Connor Runoff phosphorus losses from surface applied biosolids. J. Environ. Qual., 34: Elliott, H.A., and G.A. O Connor Phosphorus management for sustainable biosolids recycling in the United States. Soil Biology & Biochemistry, 39: Gallimore, L.E., N.T. Basta, D.E. Storm, M.E. Payton, R.H. Huhnke, and M.D. Smolen Water treatment residual to reduce nutrients in surface runoff from agricultural land. J. Environ. Qual., 28: Hamad, M.E., D.L. Rimmer, and J.K. Syers Effect of iron oxide on phosphate retention by calcite and calcareous soils. J. Soil Sci., 43: Heil, D.M., and K.A. Barbarick Water treatment sludge influence on the growth of sorghum-sudangrass. J. Environ. Qual., 18: Hsu, D.Y. and W.O. Pipes Aluminum hydroxide effects on wastewater treatment processes. J. Water Poll. Cont. Fed., 45:

40 Ippolito, J.A., K.A. Barbarick, and E.F. Rendte Co-application of water treatment residuals and biosolids on two range grasses. J. Environ. Qual., 28: Ippolito, J.A., K.A. Barbarick, and H.A. Elliott Drinking water treatment residuals: A review of recent uses. J. Environ. Qual., 40:1-12. Koplish, D., J. McMahon, and J. Valek Efficiencies of centralized residuals dewatering and co-disposal. P 8-21 to In Proceedings: 1995 Biosolids and Residuals Management. Water Environment Federation / American Water Works Association Joint Specialty Conference, Kansas City, MO. July 23-26, WEF, Alexandria, VA. Lai Y.Y. and J.C. Liu Co-conditioning and dewatering of alum sludge and waste activated sludge. Water Sci. and Tech., 50(9): Mortula, MD.M., and G.A. Gagnon Phosphorus treatment of a secondary municipal effluent using oven-dried alum residual. J. of Env. Sci. and Health Part A, 42: Novak, J.M., and D.W. Watts Increasing the phosphorus sorption capacity of southeastern Coastal Plain soils using water treatment residuals. Soil Science, 169: Peters, J.M., and N.T. Basta Reduction of excessive bioavailable phosphorus in soils by using municipal and industrial wastes. J. Environ. Qual., 25: Razali, M., Y.Q. Zhao, and M. Bruen Effectiveness of a drinking-water treatment sludge in removing different phosphorus species from aqueous solution. Separation and Purification Tech., 55: Schindler, D.W The evolution of phosphorus limitation in lakes. Science (Washington DC), 195: Sharpley, A.N., and S.J. Smith Wheat tillage and water quality in the Southern Plains. Soil Tillage Research, 30: Sharpley, A.N., and D. Beegle Managing phosphorus for agriculture and the environment. Coop. Ext. Serv. Pennsylvania State University, University Park, PA. Sharpley, A.N., T. Daniel, T. Sims, J. Lemunyon, R. Stevens, and R. Parry Agricultural phosphorus and eutrophication. Second Edition. USDA Agricultural Research Service. ARS-149. Tchobanoglous, G., L.F. Burton, and H.D. Stensel Wastewater Engineering: Treatment and Reuse, 4 th ed.; Metcalf & Eddy: New York. U.S. Department of Agriculture Agricultural Resource Service The Pennsylvania Phosphorus Index: Version 2. University Park, PA.: The Pennsylvania State University. 28

41 29 U.S. EPA Dewatering Municipal Wastewater Sludge. EPA/625/1-87/014. Office of research and development, Washington, DC. U.S. EPA Environmental indicators of water quality in the United States. EPA/841/R- 96/002. Office of Research and Development, Washington, DC. Yang Y., D. Tomlinson, S. Kennedy, and Y.Q. Zhao Dewatered alum sludge: a potential adsorbent for phosphorus removal. Water Sci. & Tech., 54: Yang Y., Zhao Y.Q., A.O. Babatunde A.O. and Kearney P Co-conditioning of the anaerobic digested sludge of a municipal wastewater treatment plant with alum sludge: Benefit of phosphorus reduction in reject water. Water Environ. Res., 79(13):

42 30 Chapter 3 Participating Utilities Facility Descriptions and Solids Characterization 3.1 The City of York; Wastewater Treatment Plant Biosolids used in laboratory experiments were generated at the City of York Wastewater treatment plant (WWTP), located in York County, Pennsylvania. The City of York WWTP has a design capacity of 26 mgd and serves a population of approximately 75,000. Typical flows in 2008 averaged around 9.3 mgd. The York WWTP uses a patented Anaerobic/Oxic (A/O) process for biological phosphorus removal (BPR) preceded by primary settling tanks. Following the A/O process, the mixture of treated wastewater and activated sludge flows to the secondary clarifiers for solids / liquid separation. Sludge collected in the clarifiers is returned to the head of the process as part of the return activated sludge process with a portion of the sludge wasted to the thickeners as needed. Effluent from the clarifiers flows to a sand filtration system for polishing. After filtration, the effluent flows though an ultraviolet light contact tank for final disinfection and then is discharged to the Codorus Creek. Using this BPR process, phosphorus (P) is reduced from an average influent value of 4.0 mg/l to below 1.0 mg/l in the final effluent Sludge Processing and Biosolids Generation The sludge processing system used at the City of York WWTP is shown in Figure 3.1. Sludge generated during the biological treatment process is wasted from the system as required and pumped to a floatation thickener.

43 31 Return Flows to headworks Sludge / Solids Flows Return / Recycle Flows Waste Activated Sludge from BPR Treatment Trains Floatation Thickeners Anaerobic Digesters Sludge Holding Tanks Sampling Location Primary Clarifiers Polymer Feed Point Inline Static Mixer Solids Truck Loading Facilities Centrifuges Centrate Recycle to headworks Return Sump DEWATERING BUILDING Figure 3.1. City of York WWTP; simplified sludge processing schematic. After thickening, the sludge is then pumped to a series of anaerobic sludge digesters and stabilized for d. Sludge from the primary clarifiers is pumped directly to the anaerobic digesters. From the digesters, stabilized sludge (referred to hereafter as biosolids) is transferred to holding tanks two times daily (one time from each of the two primary digesters). From the holding tanks the material is dewatered using two centrifuges, each with a capacity of 100 gpm. Centrate from the centrifuges is recycled back to the headworks. Prior to centrifuging, the biosolids are conditioned with organic polymers. The WWTP currently uses a proprietary branched-chain, cationic polymer. Typical dosages range from lbs (45% active) per dry ton of biosolids. This relatively high polymer dose is required

44 because the anaerobically digested sludge from the BPR process mixed with primary sludge has been historically difficult to dewater. 32 The current dewatering operation is scheduled in ten 8-h shifts, five days a week. When dewatering, the centrate flow rate (192,000 gpd) represents approximately 3% of the average daily flow. Under normal operating conditions, this dewatering system produces a biosolids cake with an average solids content of 17-18%. In 2008 the City of York WWTP generated approximately 10,500 wet tons (average 18% solids) of biosolids, all of which were land applied through an arrangement with a private contractor. Analysis data for the biosolids are presented later in this Chapter. 3.2 The York Water Company The water treatment residuals (WTR) samples used in the research were collected at the York Water Company s (YWC) Grantley Road Water Treatment Plant (WTP), located in York County, Pennsylvania. The YWC supplies an average of 20 million gallons of water per day to approximately 60,000 residential, commercial and industrial customers. WTR are generated during the process of conventional potable water treatment (e.g. coagulation, sedimentation and filtration) at the Grantley Road WTP. A schematic of the treatment process is provided in Figure 3.2. The source of raw water is the East Branch of the Codorus Creek. A consistent base flow at the intake is maintained through the management of two upstream reservoirs, Lake Williams and Lake Redman (Figure 3.2). In addition to providing storage for drought conditions, these reservoirs act as retention basins during high flow periods. As a consequence, the sediment load or turbidity in the source water is relatively consistent with an average value between NTU. The sediment in the source water

45 consists primarily of soil and sediment generated within the watershed with little input from industrial sources. 33 The primary treatment chemical used is alum for coagulation. Additional chemicals used include potassium permanganate for oxidation and, occasionally, powdered activated carbon (PAC) for taste and odor control. Prior to discharge to the finished water reservoir disinfection is provided using sodium hypochlorite and ammonia (chloramines formation) followed by lime for ph adjustment. The WTR generated at the Grantley Road WTP was selected for this study due to the consistent source water quality and the relative simplicity of the chemical treatment processes employed at the WTP. In addition, the YWC has historically pursued beneficial reuse alternatives for WTR disposal. As a consequence, considerable data exist demonstrating that this material does not contain significant levels of contaminants and is therefore well suited for beneficial reuse. Under typical conditions, the alum dose ranges between mg L -1, resulting in a total suspended solids (TSS) / alum ratio near 1:1. Physical and chemical characteristics of the WTR collected throughout this study are presented later in this chapter.

46 34 Figure 3.2. York water WTP process flow diagram Generation of WTR WTR material generated at the Grantley Road WTP consists of suspended solids from the source water and chemical precipitates created during the treatment process. These materials can be generally classified into two categories: (1) Aluminum sulfate (alum) coagulant sludge and (2) solids in filter backwash water. As illustrated in Figure 3.2, alum sludge is collected from the bottom of the inclined plate sedimentation basins and periodically pumped to a thickener. Solids captured during filtration are collected during filter backwashing and transferred to a washwater clarifier before being pumped to the thickener and mixed with the sedimentation basin sludge.

47 35 In the thickener, WTR are concentrated to approximately 5% solids. From this stage, WTR are conditioned with a cationic polymer and pumped to a dehydrator for mechanical dewatering. Solids produced from the dehydrator are in the range of 20 25% solids. All the WTR processed at the facility are hauled by a private contractor and used as a soil amendment in a beneficial reuse program. On average, the Grantley Road Plant produces between 700 to 800 dry tons of WTR annually. Prior to the installation of the mechanical dewatering system in 2007, the YWC used a series of lagoons to dewater WTR. With this system WTR were periodically pumped from the settling basins to one of several lagoons (Figure 3.2). Decant was removed from the lagoons as needed and discharged back to the East Branch of the Codorus Creek. WTR were then left to dry until they could be handled as a solid material (target of > 20% solids). WTR were then removed from the basins with front-end loaders and trucks and placed in stockpiles for further drying and storage. Dried WTR was then hauled by a private contractor and used as a soil amendment in a beneficial reuse program. WTR used in dewatering experiments (Chapter 4) and P-binding experiments (Chapter 5) were collected from the mechanical dewatering system after the thickeners and prior to the addition of conditioning polymer (sampling location D; Figure 3.2). Samples were between 2.5 and 5% solids with an average solids content of 3.77%. Additional WTR were collected from the bottom of the sedimentation basins, the lagoons and stockpiles (sample locations A, B & C; Figure 3.2) to evaluate the effect of aging and drying on the reactive Al fraction in the WTR.

48 Materials and Methods Sample Collection WTR samples used for Al ox determination were collected from three locations representing WTR at various stages of processing. Samples were collected from the bottom of the sedimentation basins, storage lagoons and stock piles (sampling locations A, B & C respectively, Figure 3.2). Further descriptions of the sampling protocol are: (A) Sed. basin samples (2.4% solids) were collected by syphoning WTR from the bottom of the basins into three 5-gallon buckets. Samples were mixed thoroughly while a 1 L subsample was collected from each bucket. These sub-samples were then mixed to form a single composite for analysis. (B) Lagoon samples (21.9%) were collected from a storage lagoon that had been left to dry for a minimum 6 months. Twelve core samples were collected randomly across the lagoon at depths ranging from 3to 12 in. The cores were then mixed together to form a single composite for analysis. (C) Stockpile samples (63.9%) were collected from 4 stockpiles that had been previously removed from one of the lagoons and left to dry for a minimum of one year. Sampling areas were cleared of any vegetation or roots that had grown on the surface of the stock pile. Three core samples were collected from each of the stockpiles at depth ranging from 3 inches to 2 feet. The cores were then mixed together to form a single composite for analysis.

49 37 Samples were analyzed for total solids in accordance with standard procedures (APHA et al., 1998) and stored in airtight containers for the duration of the study to maintain the initial moisture content and physical structure of the samples Al ox Determination Al ox was determined using a modified version of an acid ammonium oxalate extraction (McKeague and Day, 1993). The modification, based on the work of Dayton and Basta (2005) increases the dilution ratio from original of 40:1 (10 ml of extractant / 0.25 g of sample) to 100:1 (25 ml of extractant / 0.25 g of sample) and reduced the maximum particle size from 2 mm to <150 µm. These revisions were made to account for the relatively high Al ox found in WTR compared to that of soils for which the original McKeague and Day method was developed. Dayton and Basta (2005) reported a significant increase in the Al ox extracted using the 100:1 dilution for samples containing > 50 mg Al kg -1. They also reported a 2.46-fold increase in Al ox extraction when the particle size was reduced from 2 mm to <150 µm. Given the variations in Al ox extraction methods and the potentially profound effects on the results, it is important to ensure consistency in analytical methods when comparing results from different studies. In accordance with the modifications proposed by Dayton and Basta (2005) extractions were performed with a buffered solution of 0.2 M ammonium oxalate [(NH 4 ) 2 C 2 O 4 ] and 0.2 M oxalic acid (H 2 C 2 O 4 ), adjusted to ph 3. All WTR samples were passed through a 150 µm sieve. As required, dried samples were crushed with a mortar and pestle to enable passing through the sieve. This was particularly important for the dry samples (> 60% solids) as they formed into very hard rock-like fragments, that when remained uncrushed were not dissolved in the extraction process. Samples were then analyzed for Al by inductively coupled plasma atomic emission spectroscopy (ICP-AES) following digestion using U.S. EPA method 3051 (U.S. EPA, 1994).

50 Evaluation of Al ox in WTR Over Time Al ox in WTR samples from the sampling location A,B & C (Figure 3.2) was determined after 10, 150 and 300 d following sample collection. For the duration of the experiment samples were stored in airtight containers to maintain the initial moisture content of the samples. The solids content of the samples was verified in accordance with standard procedures (APHA et al., 1998) prior to each evaluation. For all samples, Al ox was extracted and analyzed as described in the previous section Dewatering and Drying of WTR Samples Mechanical Dewatering & Drying A sample of freshly precipitated WTR (4.8% solids) was collected after the thickeners (sampling location D; Figure 3.2). The wet WTR sample was preserved in an airtight container. Within 48 h of sample collection, a portion of the sample was mechanically dewatered in the lab to evaluate the effect of dewatering on the Al ox concentration. Samples were centrifuged for 20 min at 3,500 rpm to thicken the solids fraction to approximately 5 to 10%. Centrifuged solids were then dewatered using a bench-top filter press or oven drying, depending on the experiment. Oven dried samples were dried at C for 4 h and analyzed for solids content. Pressed samples were dewatered using a bench-top Crown Press Belt Press (BFP) Simulator. Samples were pressed for 4 min at 200 psi to generate a BFP cake solids content between 18 and 25% solids. To facilitate removal of the pressed WTR from the bench-top press and avoid fouling of the belt, filter paper (17 µm nominal pore size) was used to line the belt. This filter paper absorbed the water squeezed from the samples, thereby facilitating handling and

51 39 removal of the cake from the press simulator. BFP cake samples were then stored in air-tight containers to maintain consistent moisture content prior to analysis of water extractable phosphorus (WEP) and solids content. Al ox was measured as described in the previous section. Lagoon Drying A second portion of the freshly precipitated wet sample (4.8% solids) was air-dried in the lab in a simulated lagoon. Wet WTR was poured into a 9 x 12 plastic tub to a depth of ¼. On a 24-h cycle an additional ¼ of WTR was added for 3 more days to provide a total depth of 1. The WTR were then left to air dry in the lab with an average temperature of 72 o F. After one week; 1/8 of distilled water was added to the tub to simulate wetting / drying cycles common to uncovered lagoon systems. This wetting cycle was repeated again after another 3 d once the initial 1/8 of water had evaporated. After 14 d of drying the lagoon dried WTR was removed from the tub and prepared for Al ox extraction and analysis as described in the previous section.

52 Results and Discussion Biosolids Characterization The sludge samples used for this study were collected from the sludge holding tank effluent line at the location indicated in Fig Samples were collected prior to the addition of polymer. Biosolids samples collected at this location were consistently between 1 and 2% solids. The physical and chemical characteristics of the biosolids, as collected and following processing, are summarized in Table 3.1. Table 3.1. Selected properties for City of York WWTP sludge & biosolids. Digested Sludge (1) Processed Biosolids (2) Parameter Average (n = 3) ph % solids 1.32% 23.2% Phosphorus Parameters Total Phosphorus (g kg -1 ) :100 WEP (g kg -1 ) (3) PWEP (% of P T ) % Nitrogen Parameters Total Nitrogen (N) (g kg -1 ) Ammonium N (NH 4 -N) (g kg -1 ) Calculated Organic N (g kg -1 ) Total Elemental Contents Total Calcium (g kg -1 ) Total Magnesium (g kg -1 ) Total Iron (g kg -1 ) Total Sodium (g kg -1 ) Total Aluminum (g kg -1 ) Results expressed on dry weight basis unless otherwise noted (1) Collected prior to polymer addition and dewatering (see Fig. 3.1) (2) Centrifuged and dewatered with bench-scale belt filter press (3) Measured WEP in sample with polymer added at optimal polymer dose

53 41 Total N (68.8 g kg -1 ) and P (33.8 g kg -1 ) concentrations are representative of biosolids produced nationally (U.S. EPA, 1985). Total P and WEP (2.36 g kg -1 ) are also typical of anaerobically digested biosolids from BPR processes (Brandt et al., 2004). Compared to biosolids WEP values from conventional activated sludge processes ( g kg -1 ) (Brandt et al., 2004), the relatively high WEP content of these samples make them well suited for this study examining the effects of WTR addition on WEP concentrations in biosolids. Total elemental concentrations are also representative of biosolids produced nationally (U.S. EPA, 1994) and reflect the fact that the chemical precipitants (i.e. alum or FeCl 3 ) are not used at this BPR facility. Total Ca, Fe and Al concentrations were slightly higher in the processed biosolids (23.2% solids) than in the liquid samples (1.32%). It is suggested that this result is a combination of experimental error and variability in the chemical composition of the samples.

54 WTR Characterization Properties of the WTR are shown in Table 3.2. The characteristics of the York WTR are typical for alum-based WTR (ASCE et al., 1996; Agyin-Birikorang et al., 2009; Ippolito et al., 2011). Table 3.2. WTR characteristics for the York Water Company s Grantley Road WTP. Parameter Thickened WTR (1) Dewatered WTR (2) Average (n = 3) ph % solids 3.77% 19.7% Phosphorus Parameters Total Phosphorus (g kg -1 ) :100 WEP (g kg -1 ) (3) PWEP (% of P T ) % Nitrogen Parameters Total Nitrogen (N) (g kg -1 ) Ammonium N (NH 4 -N) (g kg -1 ) Calculated Organic N (g kg -1 ) Total Elemental Contents Total Calcium (g kg -1 ) Total Magnesium (g kg -1 ) Total Iron (g kg -1 ) Total Sodium (g kg -1 ) Total Aluminum (g kg -1 ) Results expressed on dry weight basis unless otherwise noted (1) Collected from thickeners prior to polymer addition and dewatering (2) Centrifuged and dewatered with bench-scale belt filter press (3) Measured WEP in dewatered WTR sample (no polymer)

55 43 The nutrient content of the WTR (Total P = 0.2%, Total N = 0.6%) is also typical for alum-based WTR (Ippolito et al., 2011) and is reflective of the relatively low nutrient content of WTR when compared to that of biosolids (Total P = 3.3%, Total N = 5.0) (U.S. EPA, 1995). Thus, coprocessing of WTR and biosolids is expected to have the effect of diluting the nutrient value of the biosolids. Effects of coprocessing on nutrient concentrations and implications on land-based application are examined in Chapter 5. The total Al (Al T ) concentration constitutes approximately 13% of the total dry weight, attributed primarily to the use of alum in the treatment process. This value is on the high end of the range ( %) of typical values given by Ippolito et al. (2011) and reflects the sweep flocculation coagulation strategy used at the Grantley Road WTP (Raw water TSS / alum dose ratio is near 1:1). Notably, the Al T drops to 8.8% when the WTR was dewatered in the lab. (A description of the dewatering procedure used is provided in materials and methods section). Centrate generated during the initial stage of dewatering was analyzed and found to contain essentially no dissolved Al; indicating that the Al is predominantly contained in the solids cake (10% solids). During the next stage of the bench-top dewatering process, the cake was pressed with the belt press to 19.7% solids. During this procedure, the water squeezed from the cake was absorbed entirely by the filter paper used to line the belt. For this reason, this liquid fraction could not be readily analyzed. It is suggested that during this dewatering process, a fraction of the Al is lost to the liquid phase as fine solids; possibly colloidal Al hydrous oxide particles become enmeshed within the filter paper pores, thereby resulting in a lower Al T concentration in the pressed cake. This appears to be a potential limitation of the bench-top dewatering process used and may therefore warrant further investigation in future studies.

56 Ammonium Oxalate Extractable Al (Al ox ) in WTR Because Al constitutes such a sizable portion of the total WTR mass, virtually all studies investigating alum-based WTR for P reduction report the Al T content. Total Al values in WTR typically range from 9.4 to 14.2% of the total dry mass (Ippolito et al., 2011). Studies have shown, however, that a measure of the reactive fraction of the Al T is a better indicator of the P- sorbing potential of WTR. WTR generated from surface water supplies treated with aluminum sulfate (alum) contain reactive hydrous oxides of Al with substantial P-fixing capacity (Elliott et al., 1990). Previous studies have established a strong relationship between P-sorbing capacity and the acid NH 4 -oxalate extractable aluminum (Al ox ) content in WTR (Dayton et al., 2003; Dayton and Basta, 2005; Gallimore et al., 1999). The acid NH 4 -oxalate method extracts the amorphous forms of Al and Fe is therefore a good measure of the noncrystalline and poorly ordered Al and Fe (McKeague and Day, 1996). The relative effectiveness of WTR in fixing soluble-p therefore depends largely on the Al ox content of the WTR. The Al ox content in WTR varies widely and depends on a multitude of factors including, but not limited to; source water characteristics, treatment operational methods, coagulant dosing regimes, and storage conditions prior to land application. In particular, studies have shown that dewatering and aging (age after dewatering) can decrease the P binding capacity of WTR. Researchers suggest that this is primarily due to a reduction in the amorphous nature of the material and the mineralization of aluminum hydroxides to less reactive crystalline forms (DeWolfe, 2006). An evaluation of the Al ox content of WTR generated at the Grantley Road WTP was therefore completed to characterize the P-binding capacity of the WTR from various stages of processing. The primary objectives of this evaluation were to: quantify the Al ox in the WTR

57 45 relative to previously reported values, compare Al ox in WTR from various stages of processing and evaluate the effects of aging and drying on Al ox in WTR Al ox Concentration in WTR from Three Stages of Processing Measured Over Time Figure 3.3 shows the Al ox content of WTR samples of varying age and stages of processing. The Al ox concentration ranged from 107 g kg -1 in freshly precipitated material (sedimentation basin), to 91 g kg -1 in WTR partially dried and aged for approximately 6 months (lagoon) to 58 g kg -1 in dried WTR aged for over one year (stockpile).

58 Al ox (g/kg) cd A c c ab B ac ad Same letter indicates no significant differnece at α = Lower case letters pertain to individual samples only, and upper case letters pertain to composite average. e be e C 0 Sed Basin (2.4%) Lagoon (21.9 %) Stock Pile (63.9 %) Sampling Location (% Solids) 10 days 150 days 300 days Composite Average Figure 3.3. Al ox concentration in WTR from three stages of processing. [Bars indicate 95% C.I. around one standard error for individual samples (n=2)]. [Average is composite of samples from given location (n = 6)]. Although it is tempting to conclude that aging and drying are responsible for the reduction in Al ox, it must be noted that the exact age and treatment conditions (i.e. coagulant dose and raw water characteristics) under which the lagoon and stockpiled materials were generated cannot be precisely defined. In surface water treatment, the optimum coagulant dose relative to the amount of sediment being removed from the water changes in response to changes in raw water quality. The inherent variability in the chemical composition of WTR means that comparisons of materials of varying age must be interpreted with caution. A second factor that may potentially change the chemical composition of the WTR is the loss of fine particles to the liquid stream during decanting of the lagoons or during precipitation events.

59 47 To evaluate the effect of aging, independent of drying conditions, samples were stored in air tight containers and sampled for Al ox after 150 d and again at 300 d (Figure 3.3). The moisture content of the samples was measured prior to each event and was found to remain constant over the 300 d period. The results reveal that for all three samples there was no significant change (α = 0.05) in the Al ox concentration over 300 d. These results suggest that time alone (i.e. assuming a constant moisture content) does not cause a reduction in Al ox through conversion to a less reactive crystalline form of Al.

60 Effect of Drying on Al ox Concentration in WTR A second set of experiments was conducted on a new batch of freshly precipitated WTR (Al ox = 78 g kg -1 : Al T = 114 g kg -1 ) to evaluate the effect of mechanical dewatering and drying on the Al ox concentration (Figure 3.4). The WTR samples used for this experiment were collected almost a year after the samples used in the aging experiment were collected (Al ox = 107 g kg -1 : Al T = 133 g kg -1 ). The difference in the Al ox concentrations for these samples collected from the various sample locations demonstrates the variability in the chemical composition of the WTR. Same letter indicates no significant difference α = a a a Figure 3.4. Effect of dewatering and drying on the Al ox in WTR samples of equal age. [Bars indicate 95% C.I. around one standard error (n=4)]. The initial Al ox concentration was reduced from 78 g kg -1 in the wet sample to 71 g kg -1 in the oven-dried sample and 70 g kg -1 in the lagoon dried sample. The reduction in Al ox after

61 49 dewatering and drying was very slight and was not statistically significant (α = 0.05). These results suggest that neither mechanical dewatering nor drying alone cause a significant decrease in the Al ox concentration. An alternative explanation as to why reductions are not observed is provided in a study by Agyin-Birikorang and O Connor (2009). In this study, the researchers reported no reduction in Al ox concentration for WTR samples incubated at 52 o C for 24 weeks. They proposed that 200 mm oxalate extraction (same extractant strength used in this study) was too strong and was actually extracting all of the Al. When they reduced the extractant strength to 5 mm, a significant reduction in Al ox was reported. Although this is worth consideration, it is noted that the material used in their study was generated from a source water with low turbidity and high color and organic content. For this reason, it is likely that essentially all of the Al in the sample was hydrous oxides of alum from the alum used in the treatment process. By contrast, samples of sediment from the raw water intake at the Grantley Road WTR was found to contain on average 116 g kg -1 Al T. In the study reported herein a 20% difference in Al T and Al ox (78 g kg -1 ) indicates that the 200 mm extractant was not extracting all of the Al. This finding underscores the difficulty of making generalizations about the behavior and characteristics of WTR, given the vast range of variability from plant to plant.

62 Summary and Conclusions The York Water WTR and York biosolids are fairly typical for material produced nationally. Due to the consistency of the source water quality and relative simplicity of the chemical treatment process, the York WTP provided a reliable source of WTR. Additionally, because BPR facilities such as the York WWTP generally have a higher biosolids WEP, efforts to sustain land application under proposed P-based nutrient magagement will be greatest for BPR facilities. For these reasons the samples provided by the participating facilities are well suited for puposes of this study. As a measure of P-sorbing capaciity, Al ox was measured in WTR samples of varying age and stages of processing. The Al ox ranged from 107 g kg -1 in freshly precipated material to 58 g kg -1 in WTR stockpiles and aged for over a year. A second set of experiments evaluated the Al ox in freshly precipitated WTR samples using different dewatering and drying methods. The initial Al ox was reduced from 78 g kg -1 to 71 g kg -1 in the oven-dried sample and 70 g kg -1 in the lagoon dried sample. Observed reductions in Al ox were not significantly different (α = 0.05), indicating that mechanical dewatering or drying alone did not cause a significant decrease in the Al ox concentration. Results from Al ox experiments suggest that a combination of aging and drying (or thermal stabilization) maybe required to reduce the Al ox concentration through mineralization to less reactive forms as suggested in previous studies (DeWolfe, 2006). Further studies of broader scope (i.e. different WTR samples and more measurements) are recommended to better define the conditions under which Al transformations in WTR occur.

63 51 Regardless of the reasons for the decrease in Al ox, the P-binding capacity as measured by Al ox concentration appears to be highest for the freshly precipitated material followed by the lagoon and stockpiled materials, respectively. This trend supports the concept of blending and dewatering (coprocessing) freshly-generated WTR and biosolids to maximize the P-sorbing potential of the WTR. This theory is explored in detail in Chapter 5. In practical application these results also suggest that measurement of Al ox in addition to Al T may provide a valuable management tool for predicting the capacity for WTR coprocessing to reduce the environmental lability of biosolids.

64 References Agyin-Birikorang S., and G. O Connor Aging effects on reactivity of an aluminum-based drinking-water treatment residual as a soil amendment. Science of the Total Environment, 407: Agyin-Birikorang S., G. O Connor, and T. Obreza Drinking water treatment residuals to control phosphorus in soils. University of Florida IFAS Extension. SL 300. American Public Health Association; American Water Works Association; Water Environment Federation Standard Methods for the Examination of Water and Wastewater, 20 th Edition.; Washington, D.C. American Society of Civil Engineers, American Water Works Association, and the United States Environmental Protection Agency Technology transfer handbook: Management of water treatment plant residuals. ASCE and AWWA. New York, New York and Denver, CO. Brandt R.C., H.A. Elliott, and G.A. O Connor Water-extractable phosphorus in biosolids: Implications for land-based recycling. Water Env. Research, 76(2): Dayton, E.A., N.T. Basta, C.A. Jakober, and J.A. Hattey Using treatment residuals to reduce phosphorus in agricultural runoff. J. Am. Water Works Assoc., 95(4): Dayton, E.A., and N.T. Basta A method for determining the phosphorus sorption capacity and amorphous aluminum of Al-based drinking water treatment residuals. J. Environ. Qual., 34: DeWolfe, J.R Water residuals to reduce soil phosphorus. AWWA Research Foundation, Denver, CO. Gallimore, L.E., N.T. Basta, D.E. Storm, M.E. Payton, R.H. Huhnke, and M.D. Smolen Water treatment residual to reduce nutrients in surface runoff from agricultural land. J. Environ. Qual., 28: Ippolito, J.A., K.A. Barbarick, and H.A. Elliott Drinking water treatment residuals; A review of recent uses. J. Environ. Qual., 40:1-12. McKeague, J.A., and J.H. Day Ammonium oxalate extraction of amorphous iron and aluminum. P In M.R. Carter (ed.) Soil sampling methods of analysis. Lewis Publ., Boca Raton, FL. U.S. EPA Microwave assisted acid digestion of sediments, sludges, soils, and oils Method In Test Methods of Evaluating Solid Waste, Physical / Chemical Methods, 3 rd ed.; SW-846. U.S. EPA Process design manual: Land application of sewage sludge and domestic sewage. EPA/625/R-95/001. Office of Res. And Development, Cincinnati, OH. 52

65 53 Chapter 4 Influence of Water Treatment Residuals on Dewaterability of Wastewater Biosolids 4.1 Introduction Drinking water treatment processes generate a variety of residual products depending on characteristics of the source water and types of unit operations used. In conventional coagulationflocculation treatment, suspended solids and organic matter in the source water are removed by the addition of aluminum or iron salts as coagulants. While water treatment residuals (WTR) were once primarily discharged back to surface waters without further treatment (Elliott et al., 1990), regulatory constraints on this practice have increased efforts to find alternative and beneficial options for dealing with WTR. A recent review (Ippolito et al., 2011) focused on use of WTR as a sorbent for P and other constituents. Discharge of WTR into sanitary sewers has long been practiced as a convenient and cost effective method for municipalities to handle the residuals produced in water purification (U.S. EPA, 1996). Some issues are increased solids loadings, potential negative impacts on the activated sludge process, and changes in digester operation at the wastewater treatment plant (WWTP). More recently, there has been interest in blending of the WTR post-digestion but prior to dewatering. This has the advantage of centralizing dewatering operations for both WTR and wastewater solids but has the potential to change the dewaterability and handling characteristics of the solids produced. A few studies have investigated the blending and dewatering of WTR and wastewater solids simultaneously (hereafter referred to as coprocessing). In 1995, the City of Allentown, PA, conducted full-scale pilot studies where alum-based WTR were trucked to the WWTP and blended with anaerobically-digested solids prior to polymer addition and belt-filter press (BFP)

66 54 dewatering. Using a 50:50 (mass basis) WTR / biosolids blend, the solids content of the BFP cake was increased to 25.6% compared to 20.8% solids for WWTP sludge alone. In addition, the polymer dose required for dewatering the blend was half that needed for the biosolids alone. Using fixed costs for chemicals, transportation and disposal, it was concluded that coprocessing could reduce processing and disposal costs by approximately 20% (Koplish et al., 1995). Lai and Liu (2004) investigated the coprocessing of alum sludge and waste-activated sludge and found that dewaterability, as measured by capillary suction time (CST) and specific resistance to filtration (SRF), improved with increasing fraction of alum sludge. They proposed that the alum sludge acted as a skeleton builder and rendered the mixture more incompressible. Using SRF as a measure of dewaterability, Yang et al. (2007) also reported an improvement in the dewatering characteristics of the coprocessed mixture relative to the digested sludge alone. With the recent trend toward privatization of both water and wastewater treatment operations, there is renewed interest in the savings associated with eliminating redundant operations. Dewatering, a unit process common to water and wastewater treatment, is one operation which can potentially be centralized. This chapter addresses the influence of added WTR on the dewaterability of biosolids. Specifically, the aim was to assess how blending of WTR and biosolids influenced the required polymer dosage and resulting cake solids concentration following belt-filter pressing. Additionally, insight is provided into the interaction of WTR and wastewater solids as they influence dewatering and towards identifing conditions under which coprocessing of WTR and biosolids would be practical for a municipality. 4.2 Capillary Suction Time Test as an Indicator of Sludge Dewaterability The capillary suction time (CST) device, invented by Baskerville and Gale (1968) provided the first simple, rapid and inexpensive means for measuring the dewatering potential of

67 55 wastewater sludges. Since the 1970s, the CST has been used widely in wastewater and water treatment plants and research applications to quantify changes in sludge dewaterability (Vesilind, 1988). The CST device measures the time to move a volume of filtrate over a specified distance as a result of the capillary suction pressure of dry filter paper. The CST device consists of two plastic blocks, a stainless steel collar, a piece of filter paper, three electrical contacts that are attached to the upper block and an electrical timer (Figure 4.1). The equipment is assembled with the filter paper placed between the two plastic blocks and the collar placed in the opening of the upper block. Initially, a volume of sludge is introduced into the collar. Water is drawn from the base of the collar by the capillary suction pressure of the filter paper forming a wetting front on the filter paper which moves radially outward. As the wetting front advances outward, it reaches the first of two contacts (sensor 1) and an electrical signal starts the timer. When the front reaches the outer contact (sensor 2) the timer is stopped. Figure 4.1. Schematic diagram of the CST apparatus

68 CST (sec) 56 The CST value (time in s) is then read directly from the timer. The CST is a measure of the ease with which water is released from the biosolids slurry. A low CST time therefore indicates relative ease in separating the water from the solids and implies the sludge has good dewaterability. Based on extensive testing of the CST device and past operating experience, a CST below 20 s is generally required to achieve good dewaterability with a belt-filter press (Triton Electronics Ltd., 1998; U.S. EPA, 1987). In practice, the filterability of wastewater solids is improved with the addition of a conditioning agent such as a cationic polymer. To illustrate, Figure 4.2 shows CST values for anaerobically-digested biosolids collected from the Penn State WWTP (1.6% solids initially) as a function of cationic polymer dose. From this type of experiment, the amount of conditioning polymer required for favorable dewaterability can be determined Polymer Dosage (mg L -1 ) Figure 4.2. Typical results showing the effect of chemical conditioning on biosolids CST. [Anaerobically digested biosolids (1.6% solids) collected from the Penn State WWTP and conditioned with cationic polymer].

69 57 Due to the empirical nature of the CST test, the results are unique to a given sludge and the instrument being used. For this reason, it is generally not meaningful to compare CST values from one sludge to another (Vesilind and Davis, 1988). Reported CST values for unconditioned wastewater sludge range widely from less than 100 s (Chen et al., 1996), to over 1,200 s (Lai and Liu, 2004). The U.S. EPA design manual for dewatering municipal wastewater sludge (U.S. EPA, 1987) suggests typical CST values for unconditioned sludges range between 100 and 300 s. Similar variability is reported for WTR, with CST values ranging from 45 s (Lai and Liu, 2004) to over 190 s (Yang et al., 2007). Key factors affecting the CST value of a given slurry include the initial solids content and the viscosity of the filtrate being withdrawn (Triton Electronics Ltd., 1998). The extent to which CST is affected by the initial solids content is dependent on the characteristics of the material being tested, however in general, the CST will increase with increasing solids content of the sample (Vesilind, 1988; Triton Electronics Ltd., 1998). Secondary factors include homogeneity of the sludge, type of digestion, inorganic content and sludge age (Triton Electronics Ltd., 1998; U.S. EPA, 1987). The CST test is therefore most useful for comparing required dosages of various conditioning agents or selecting the optimum dosage for a given sludge. Variability in CST testing has also been attributed to the empirical nature of the test apparatus itself. Initial efforts to model radial water flow for the CST test assumed constant porosity, capillary suction pressure and hydraulic permeability of the filter paper (Lee and Hsu, 1992). However, Meeten and Smeulders (1995) reported that when examining radial flow in the filter paper, water saturation varies with distance away from the center of the sample collar and hence neither the capillary suction pressure nor the effective water porosity are constant. As such, a theoretically mathematical description of the CST operation is difficult (Scholz and Tapp, 2006; Smiles, 1998). A second limitation with the conventional CST device is the inability to account

70 58 for sedimentation of particles during the test. During the course of the test, sediments such as suspended solids and heavy flocs can accumulate on top of the filter paper, hindering water flow that may subsequently lead to an overestimation of the cake resistance to dewatering (Scholz and Tapp, 2006). Efforts to reduce the variability of CST measurements and allow for comparison of CST values for various sludges have lead to the development of revised test procedures and proposed modifications to the CST apparatus (Scholz and Tapp, 2006). A study by Sawalha and Scholz (2007) found that using a rectangular funnel in place of a standard circular funnel significantly improved test repeatability through reduced variability in CST measurements. In addition, stirring the sludge samples during the CST test resulted in a small reduction of the sedimentation effect (Sawalha and Scholz, 2007). Despite its limitations, the CST test is inexpensive, simple to conduct, and can rapidly provide data on the effect of chemical conditioners on sludge dewaterability (Vesilind, 1988). The CST test therefore remains a useful empirical tool for practicing engineers and operators (Smiles, 1998). 4.3 Methods The wastewater solids were collected at the City of York WWTP located in York, PA. This facility uses biological phosphorus removal (BPR) for nutrient control. Primary and waste activated sludge are combined and anaerobically digested. Digested sludge samples (~ 1.3% solids) were collected from the sludge holding tanks prior to polymer addition and dewatering by centrifugation. Because the sludge at the point of collection meets the U.S. EPA pathogen and pollutant requirements for land application, hereafter the material is referred to as biosolids. The alum-based WTR (Al-WTR) was collected from the York Water Company s Grantley Road

71 59 facility, located in York County, PA, which uses aluminum sulfate as the primary coagulant. The Al-WTR samples were collected from a thickening tank prior to chemical addition and dewatering. The Al-WTR (~ 4 5% solids) is a composite of solids from the sedimentation basins and filter backwashing operations. Elemental content of the biosolids and Al-WTR were determined by acid digestion followed by analysis via inductively coupled plasma atomic emission spectroscopy using U.S. EPA methods 3051 and 6010B. Nitrogen was determined by Kjeldahl analysis. Solids and ph (1:1 solids-to-distilled water) were determined by standard procedures (APHA et al., 1998). Specific gravity was also measured by standard procedures (APHA et al., 1998) to facilitate volume-to-mass conversions. As a measure of the reactivity of Al in the Al-WTR, the oxalate-extractable Al content (Al ox ) was determined by the method of McKeague and Day (1993) and revised for Al ox determination in WTR by Dayton and Basta (2005a). The polyelectrolyte used by the York WWTP, Pollu-treat CL-455 from (Pollu-tech Inc.), was also used in the laboratory experiments. Pollu-treat CL-455 is a branched chain, cationic product that is commonly used for hard to dewater sludges. For laboratory experiments, a polymer solution was prepared by diluting concentrated polymer with distilled water to a concentration of 0.2% (1 ml polymer / 499 ml water). The diluted solution was slowly mixed for 45 min until no visible clumps of concentrated polymer remained. Polymer solution was used within 48 h of preparation. Volumes of Al-WTR were mixed with 200 ml of biosolids to give various blending ratios (BR), defined as the mass ratio of Al-WTR to biosolids on a dry weight basis. The blended samples were mixed for 1 h on a reciprocating shaker at 200 evolutions per min (EPM). Mixed samples were removed from the mixer and placed in 500-mL jar-testing beakers. For coprocessing experiments in which polymer was added, a 0.2% polymer solution was injected into the WTR/ biosolids blends. Samples were placed on the jar tester at 200 rpm and polymer

72 60 solution was injected with a syringe into the vortex of the mixing zone approximately 2 cm below the liquid surface. Mixing continued at 200 rpm for 20 s and then was reduced to 60 rpm for an additional 90 s. Capillary suction time (CST) was used to evaluate the dewaterability characteristics for Al-WTR, biosolids, and blended materials. The CST was measured using a Triton CST Apparatus (Triton Electronics Ltd., Type 319, Dunmow Essex, United Kingdom) fitted with a 1.8-cm diameter cylinder. In accordance with manufacturer s recommendations, Whatman No. 17 chromatographic paper (7 cm x 9 cm) was used for filters. A 6 ml portion of the blended sample was poured into the funnel, and the CST was recorded a minimum of 6 times for each sample. Sampling technique was developed such that sample variance was less than 15%, as recommended by the equipment manufacturer. The operational polymer dose (OPD) was defined as the minimum polymer dose consistently resulting in a CST time of 20 s (U.S. EPA, 1987), reported as the average of 6 CST measurements with a maximum coefficient of variation of 15%. The OPD is interpreted in this study as an operationally defined value indicative of good sludge dewaterability. In practice, the optimal or most cost efficient polymer dose may vary from the OPD depending on polymer costs, dewatering methods and targeted solids content. Nonetheless, the OPD serves as a useful operational tool to evaluate the action of chemical conditioners and WTR on sludge dewaterability. For the dewatering experiments, samples were centrifuged for 20 min at 3,500 rpm to thicken the solids fraction to approximately 5 to 10%. Solids were further dewatered using a bench-top Crown Press Belt Filter Press (BFP) Simulator. Samples were pressed for 4 min at 200 psi and then the solids content was determined (APHA et al., 1998). Statistical analysis of data was conducted by evaluating the analysis of variance (ANOVA) of the means of various treatments. Prior to analysis, data sets were evaluated to verify that the normality and

73 61 homogeneity of variance (α = 0.05) requirements for parametric ANOVA evaluation were satisfied. 4.4 Results and Discussion Characteristics of Biosolids and WTR The characteristics of the biosolids and Al-WTR are presented in Table 4.1. The biosolids had high total N and P levels characteristic of biological nutrient removal processes (Brandt et al., 2004). The Al-WTR total N content (12.5 g kg -1 ) is higher than typical (Ippolito et al., 2011) possibly due to filter backwashing with water treated with chloramines. The total P content is typical for WTR (Ippolito et al., 2011). The biosolids major elemental concentrations (Ca, Mg, Fe, Na, Al) are typical of materials produced nationally and reflect wastewater and sludge treatment processing. The high Al content of the WTR (132.9 g kg -1 ) reflects the use of alum as primary coagulant and is approximately 80% amorphous as reflected by extraction via ammonium oxalate. The CST of the biosolids (355 s) was within the range reported in the literature. Reported CST values for unconditioned wastewater sludge range widely from less than 100 s (Chen et al., 1996) to over 1,200 s (Lai and Liu, 2004). The U.S. EPA design manual for dewatering municipal wastewater sludge (U.S. EPA, 1987) reported typical CST values for unconditioned sludges range between 100 and 300 s. The CST of Al-WTR is generally reported to be lower than that of wastewater solids. Although Yang et al. (2007) reported CST of alum sludge to range from s, most cited values are lower and similar to the value (25.5 s) found in this study. Reported values for Al- WTR are 45 s (Lai and Liu, 2004) and 64 s (Ghebremichael and Hultman, 2004).

74 Papavasilopoulos and Bache (1998) found that CST varied with solids content and ranged from 15 to 53 s for solids concentrations of approximately1150 to 6300 mg L -1, respectively. 62 Table 4.1. Characteristics of biosolids and Al-WTR. Parameter Biosolids (1) Al-WTR (2) Average (n = 3) ph % solids 1.32% 3.77% Capillary Suction Time (s) Total Nitrogen (N) (g kg -1 ) (3) Ammonium N (NH 4 -N) (g kg -1 ) Calculated Organic N (g kg -1 ) Total Phosphorus (g kg -1 ) Total Calcium (g kg -1 ) Total Magnesium (g kg -1 ) Total Iron (g kg -1 ) Total Sodium (g kg -1 ) Total Aluminum (g kg -1 ) Oxalate-Extractable Al (g kg -1 ) (1) Collected, post-digestion, from thickeners prior to polymer addition and dewatering (2) Collected from thickeners prior to polymer addition and dewatering (3) Results expressed on dry weight basis Influence of Coprocessing on Dewaterability Figure 4.3 shows the dewatering behavior as measured by CST of the biosolids and Al- WTR alone and at two blend ratios as a function of polymer dose. It is evident that increasing the polymer dose improved dewatering capacity, although the extent of the effect varied among the materials. For each material, experiments were terminated when the CST fell below 20 s, considered the threshold for good dewaterability using a belt filter press (Triton Electronics Ltd., 1998). For a given polymer dose, the CST generally decreased as the amount of Al-WTR in the material increased.

75 63 Because the initial solids content of the WTR (3.77%) is higher than that of the biosolids (1.32%), increasing the BR will result in an increase in the initial solids content of the blends (BR = 0.75: 2.37%; BR = 1.5: 2.79%). Previous studies have shown that CST generally increases with increasing initial solids content (Vesilind, 1988; Triton Electronics Ltd., 1998). In this study, the effect of increasing BR on the CST appears to offset any influence of initial solids content as evidenced by a decreasing CST value with increasing BR. In fact, if the CST values had been corrected for initial solids content, the effect of increasing BR on the CST value may be more pronounced that shown in Figure 4.3. In this study, CST values were not corrected for initial solids content given the relative similarity in solids content (2.4 % difference) of the two materials. Future research on coprocessing using materials with more pronounced differences in solids contents should account for the potential influence of initial solids content on CST values. To determine if the effect of Al-WTR addition on CST was statistically significant, ANOVA analysis was conducted on the mean CST values at a common polymer dosage (14.4 g kg -1 ). A polymer dose of 14.4 g kg -1 was selected as a point of comparison because it was the was the lowest common polymer dose for which the OPD had been reached for any of the 3 blend ratios. ANOVA analysis yielded a p value < 0.001, demonstrating a statistically significant difference in the mean CST values (shown in Figure 4.3) at this polymer dosage. The reduction in CST of the biosolids with blending implies that the addition of WTR changes the properties of the biosolids structure to allow water to be more readily released. The exact nature of this process is complex and likely involves multiple effects depending on the characteristics of the biosolids and WTR particles, environmental factors, and mixing conditions.

76 CST (seconds) CST values at common polymer dose of 14.4 g kg (a) Biosolids BR = 0.75 BR = 1.5 WTR OPD = 20 sec 40.7 (b) (c) (14.4) Polymer dose (g kg -1 dry solids) Figure 4.3. Capillary suction time (CST) for the materials. Operational polymer dose (OPD) designated as polymer dose producing mean CST value of 20 s. Values are average of independent samples (n = 6). BR = WTR / biosolids blend ratio: dry mass basis. At a common polymer dose of 14.4 g kg -1, mean CST values followed by different letters are statistically different at α = It has been suggested that the Al-WTR acts as a skeleton builder, transforming the biosolids into a more rigid, less compressible structure (Lai and Liu, 2004). Here the Al-WTR is viewed as a physical conditioner, similar to the use of fly ash, cement kiln dust, gypsum, and other inert admixtures in sludge dewatering (Zhao, 2006). If the blending process is considered as the addition of a more easily dewatered material (e.g., WTR) to a poorly dewaterable amorphous biosolids mass, the mixture should more readily release water as the proportion of the WTR increases. In the present study, because of the enhanced dewaterability of the Al-WTR

77 65 (CST = 25.5 s) compared to the biosolids (CST = 355 s), a physical conditioning mechanism cannot be discounted. However, an argument can be made that electrochemical interactions between the Al- WTR and biosolids provide for aggregation mechanisms distinctly different from a physical skeleton-building phenomena. An interesting feature of the data in Figure 4.3 is the identical CST for the two blended samples (BR = 0.75 and 1.5) in the absence of polymer addition. This implies there is a limit to which the addition of WTR will improve sludge dewaterability in the absence of polymers. If the WTR was acting solely as a skeleton builder, doubling the proportion of Al-WTR in the blend should further enhance dewaterability as reflected in a lower CST. If enhancement of the dewatering characteristics can be viewed as agglomeration or aggregation of the biosolids particles to permit release of the entrained or bound water, it is possible that the added WTR served as a conditioner by promoting coagulation-flocculation of the biosolids particles. One potential mechanism of particle destabilization is heterocoagulation, or coagulation initiated by colloids carrying the opposite surface charge under the given conditions. Sewage sludge particles tend to be negatively charged at near-neutral ph conditions. The WTR particles can carry a net negative or positive charge, depending on the nature of the treatment process in which they are generated. If the WTR solids are largely composed of the inorganic silicate-rich sediment particles removed from the source water, it is likely that the WTR particles carry a net negative charge at near-neutral ph, which is typical of pure silica and hydrous oxides dominated by silica (Elliott and Huang, 1981). Researchers have indeed found that WTR particles have a negative zeta potential at circumneutral ph (Huang et al., 2002; Lai and Liu, 2004). However, Yang et al. (2007) reported alum sludge had a positive zeta potential. This WTR had a high total Al content (19.5% dry basis), and thus the WTR particles likely behave more like colloidal Al oxides or hydroxides, which are positively charged at circumneutral ph conditions (Parks, 1965).

78 66 A different mechanism that could be involved is that the Al-WTR caused coagulationflocculation of the sludge solids through adsorption-charge neutralization by polymerized hydrolysis products (e.g., Al(H 2 O) 5 OH 2+ ) remaining in solution from the alum addition in the water purification process. This mechanism is probably more prevalent for WTR generated where treatment involves higher alum doses and a sweep floc mode of operation. In the absence of zeta potential measurements to establish the electrokinetic properties of the WTR particles or soluble Al content in the interstitial water of the WTR, it may be difficult to make definitive statements about the mechanisms of enhanced dewatering of biosolids in response to Al-WTR additions. However, the total Al content (Al T ) of the WTR can be used in a general way to infer the nature of the interaction between WTR and biosolids particles. The Al T of WTR is highly variable and has been reported to range from ~3% (Dayton et al., 2003) to as high as 19.5% (Yang et al., 2004), with mean values reported at 11.9% (Ippolito et al., 2011). It seems plausible that WTR with relatively high Al T are more likely to have the capacity to flocculate biosolids particles via the mechanisms previously described. Whereas Al T is likely to be a reasonably good predictive tool for evaluating the ability of WTR to cause particle aggregation in biosolids-wtr blends, some measure of the reactivity of the Al in the WTR would be superior. In quantifying the phosphorus-fixing capacity of WTR, the reactive Al content as measured by oxalate extraction (Al ox ) has been found to be a useful parameter (Elliott et al., 2002; Dayton et al., 2003; Dayton and Basta, 2005b). Oxalate extraction is believed to measure the non-crystalline, x-ray amorphous fraction of Al and Fe oxides in soils, sediments, and WTR (Dayton and Basta, 2005a). Freshly precipitated Al hydroxide produced in coagulation basins should be predominantly amorphous in nature and have high Al ox as percentage of Al T. It is proposed that WTR with a high percentage of Al ox should also be more effective at conditioning the biosolids particles prior to coprocessing of WTR-biosolids blends. In the present study, roughly 80% of

79 67 the Al T was extractable by ammonium oxalate (Table 4.3). Since the Al ox / Al T ratio can vary widely for WTR, their ability to serve as a biosolids conditioning agent may also be quite variable. Indeed, McTigue et al. (1989) found that the addition of WTR had little influence on the drainage characteristics of the biosolids, whereas van Nieuwenhuyze et al. (1990) reported that in almost every case studied, the dewaterability of biosolids was shown to improve with the addition of WTR. The foregoing mechanistic interpretation is based on the formation of a WTR-biosolids structure amenable to dewatering via belt pressing. For centrifuge dewatering, the water removal process involves significant shear of the particles. High-sheer dewatering could result in dispersion of both the biosolids and WTRs, thereby negating any dewatering enhancement associated with coprocessing. The applicability of these results to centrifugation dewatering requires further experimentation Influence of Coprocessing on OPD and Cake Solids The OPD can be determined from Figure 4.3 where the data intersect the horizontal line for CST of 20 s. The OPD for the biosolids alone (20.6 g kg -1 ) was reduced to 16.3 and 12.6 g kg -1 for BR values of 0.75 and 1.5, respectively. To verify that the experimentally determined OPD did indeed result in good dewaterability, subsamples of the WTR / biosolids blends used in the CST experiments were dewatered with a bench top BFP as described in the methodology section. Percent solids data after dewatering are shown in Figure 4.4. As illustrated, when conditioned with polymer at the predetermined OPD, the BFP cake was above 20% for each of the 3 blend ratios tested. Results from an ANOVA test (α = 0.05) of this data revealed no significant difference in percent solids

80 % Solids among the three ratios tested. This conclusion confirms that the experimentally determined OPD was a suitable indicator of the polymer dose required to obtain good dewaterability % 22.5% Bars indicate ± one standard error 22.0% 21.5% 21.0% 21.2 (a) 20.9 (a) 21.4 (a) 20.5% 20.0% 19.5% 19.0% OPD = 20.6 OPD = 16.3 OPD = % 18.0% BR Ratios Figure 4.4. Percent solids following dewatering with bench-top belt filter press for 3 BRs conditioned with polymer at the operational polymer dose (OPD, g kg -1 ). [Values are average of independent samples (n = 3). Reported mean % solids values followed by the same letter are not significantly different (α= 0.05)].

81 Conclusions This study demonstrated that dewaterability (as assessed by CST) of anaerobically digested biosolids was enhanced by blending with Al-WTR with the effect generally being more pronounced as the fraction of Al-WTR increased. Moreover, the ability of the Al-WTR to destabilize the biosolids particles caused a substantial reduction in the OPD of the blends compared to the biosolids alone. For example, the OPD when BR was 1.5 was about 60% of the value for the biosolids alone. Since conditioning chemicals can represent half of the overall sludge management and handling costs (Christensen and Stulc, 1979), coprocessing may have a significant economic advantage. The extent to which WTR are useful in dewatering biosolids is determined to a large extent on the amount of Al hydroxide precipitate present in the WTR which, in turn, depends on the characteristics of the source water and mode of particle destabilization in the water treatment process. For example, WTR derived from highly colored, low turbidity water dosed within the sweep floc region typically contain a large amount of Al hydroxide precipitate (Papavasilopoulos and Bache, 1998). These WTR would tend to have higher total Al content and when fresh, a high Al ox as percentage of total Al. Such WTR should also have a high capacity to condition the biosolids and reduce the OPD. The research studies conducted at the bench-scale allowed a wide range of blending ratios to be investigated. In practice, however, the available quantities of WTR and biosolids are fixed and depend on many factors like the volumes and quality of treated water and wastewater, chemical dosages, and the nature of the treatment processes. For a conventional coagulationfiltration plant using alum as primary coagulant, kg of WTR will be produced per 1000 m 3 of water treated (Crittenden et al., 2005). For a conventional activated sludge plant with primary sedimentation, 240 kg of dry solids are produced per 1,000 m 3 of wastewater treated (Metcalf &

82 70 Eddy, 2003). Thus for a municipality where equal volumes of water and wastewater are treated by conventional means, the maximum BR would be Thus, the benefits of enhanced dewaterability or reduced polymer dosage afforded by coprocessing may be more limited than what is theoretically possible with higher blending ratios. Moreover, if the quantity of WTR is limited, those possessing a larger amount of reactive Al, as measured by Al ox, may be more effective in improving the structure of the wastewater solids to allow release of entrained water.

83 71 References APHA Standard Methods for the Examination of Water and Wastewater, 20 th edition. American Public Health Association / American Water Works Association / Water Environment Federation, Washington DC, USA. Baskerville, R.C., and R.S. Gale A simple automatic instrument for determining the filterability of sewage sludge. J. Inst. of Water Poll. Control., 1(3). Brandt, R.C., H.A. Elliott and G.A. O Connor Water-extractable phosphorus in biosolids: Implications for land-based recycling. Water Environment Research, 76(2): Christensen, G. L. and D.A. Stulc Chemical reactions affecting filterability in iron-lime sludge conditioning. J. Water Pollution Control Fed., 51: Chen, G.W., W.W. Lin, and D.J. Lee Capillary suction time (CST) as a measure of sludge dewaterability. Water Sci. and Tech., 34(3): Crittenden, J.C., R.R. Trussell, D.W. Hand, K.J. Howe and G. Tchobanoglous Water Treatment Principles and Design. John Wiley & Sons, Hoboken, NJ. Dayton, E.A., N.T. Basta, C.A. Jakober and J.A. Hattey Using treatment residuals to reduce phosphorus in agricultural runoff. J. of the American Water Works Association, 95: Dayton, E.A. and N.T. Basta. 2005a. A method for determining the phosphorus sorption capacity and amorphous aluminum of aluminum-based drink water treatment residuals. J. Environ. Qual., 34: Dayton, E.A. and N.T. Basta. 2005b. Use of drinking water treatment residuals as a potential best management practice to reduce phosphorus risk index scores. J. Environ. Qual., 34: Elliott, H.A. and C. P. Huang Adsorption characteristics of some Cu(II) complexes on aluminosilicates. Water Res., 15: Elliott, H.A., B.A. Dempsey, D.W. Hamilton and J.R. DeWolfe Land application of water treatment sludges; impact and management. AWWA Research Foundation, Denver, CO. Elliott, H.A., G.A. O Connor, P. Lu and S Brinton Influence of water treatment residual on phosphorus solubility and leaching. J. Environ. Qual., 31: Ghebremichael, K.A. and B. Hultman Alum sludge dewatering using Moringa oleifera as a conditioner. Water, Air & Soil Pollution, 158(1): Huang, C., J.R. Pan, C.G. Fu and C.C. Wu Effect of surfactant addition on dewatering of alum sludges. J. Env. Engng., 128:

84 72 Ippolito, J.A., K.A. Barbarick and H.A. Elliott Drinking water treatment residuals; A review of recent uses. J. Environ. Qual., 40:1-12. Koplish, D., J. McMahon and J. Valek Efficiencies of centralized residuals dewatering and co-disposal. Proceedings: 1995 Biosolids and Residuals Management. Water Environment Federation / American Water Works Association Joint Specialty Conference, Kansas City, MO. Water Environment Federation, Alexandria, VA. Lai, Y.Y. and J.C. Liu Co-conditioning and dewatering of alum sludge and waste activated sludge. Water Sci. and Tech., 50(9): Lee, D.J. and Y.H. Hsu Fluid flow in capillary suction apparatus. Ind. & Engng. Chem. Res., 31: Meeten, G.H. and J.B.A.F. Smeulders Interpretation of filterability measured by the capillary suction time method. Chem. Engng. Sci., 50: McKeague, J.A., and J.H. Day Ammonium oxalate extraction of amorphous iron and aluminum. P In Soil sampling methods of analysis, Carter M.R. (ed.), Lewis Publ., Boca Raton, FL. McTigue, N.E., D.A. Cornwell, and A.T. Rolan Impact of water plant waste discharge on wastewater plants. Proceedings: 1989 AWWA/WPCF Joint Residuals Management Conference, San Diego, CA. August Metcalf & Eddy, Inc Wastewater Engineering Treatment and Reuse, 4 th edition. McGraw Hill, New York. Parks, G.A The isoelectric points of solids oxides, solid hydroxides, and aqueous hydroxo complex systems. Chem. Rev. 65: Papavasilopoulos, E.N. and D.H. Bache On the role of aluminum hydroxide in the conditioning of alum sludge. Water Sci. and Tech., 38(2): Sawalha, O. and M. Scholz Assessment of capillary suction time (CST) test methodologies. Env. Tech., 28: Scholz, M. and J. Tapp Development of a revised capillary suction time test. Water Conditioning & Purification., 48: Smiles, D.E Water flow in filter paper and capillary suction time. Chem. Engng. Sci., 53: Triton Electronics Ltd CST Equipment Manual. Triton Electronics Ltd. Bigods Hall, Bigods Lane. Dunmow. Essex. CM6 3BE. U.S. Environmental Protection Agency Dewatering Municipal Wastewater Sludge. EPA/625/1-87/014. Office of research and development, Washington, DC.

85 73 U.S. Environmental Protection Agency Management of water treatment residuals. EPA/625/R-95/008. Office of research and development, Washington, DC. van Nieuwenhuyze, R.F., N.E. McTigue, and R.G. Lee Beneficial applications and innovative sludge disposal methods. In: Cornwell D.A. and Koppers H.M.M. eds. Slib, schlamm, sludge. Denver, CO: American Water Works Association Research Foundation: Vesilind, P.A. and H.A. Davis Using the capillary suction time device for characterizing sludge dewaterability. Water Sci. and Tech., 20: Vesilind, P.A Capillary suction time as a fundamental measure of sludge dewaterability. J. Water Poll. Cont. Fed., 60: Yang, Y., Y.Q. Zhao, A.O. Babatunde, and P. Kearney Co-conditioning of the anaerobic digested sludge of a municipal wastewater treatment plant with alum sludge: Benefit of phosphorus reduction in reject water. Water Environ. Res., 79(13): Zhao, Y. Q Involvement of gypsum (CaSO 4-2H 2 O) in water treatment sludge dewatering: A potential benefit in disposal and reuse. Separation Sci. Tech., 41:

86 74 Chapter 5 Coprocessing WTR and Biosolids: Impact on Phosphorus Partitioning and Biosolids Recycling 5.1 Introduction Previous studies have established a strong relationship between phosphorus (P)-sorbing capacity and the ammonium oxalate extractable aluminum (Al ox ) content in water treatment residuals (WTR) (Dayton and Basta, 2005). WTR generated from surface water supplies treated with alum (aluminum sulfate) contain a relatively high concentration of reactive hydrous oxides of Al with substantial P-fixing capacity (Elliott et al., 1990). The relative effectiveness of WTR in fixing soluble-p therefore depends largely on the Al ox content of the WTR. Because the total Al (Al T ) in WTR varies widely depending on a variety of factors such as source water characteristics, treatment operational methods and coagulant dosages (Ippolito et al., 2011; ASCE, 1996), the resulting Al ox concentration in WTR also varies widely. In addition, studies have shown that Al ox concentrations can vary depending on dewatering methods, storage conditions and sample age and can decrease over time due to a reduction in the amorphous nature of the material and the mineralization of aluminum hydroxides to less reactive crystalline forms (DeWolfe, 2006; Hyde and Morris, 2000). The Al T and Al ox concentration are therefore important parameters in assessing the P-sorbing capacity of WTR. Whereas previously conducted research has focused on the co-application of WTR to offset the soluble-p availability from biosolids using surface application of WTR (Dayton and Basta, 2005b; Gallimore et al., 1999; Peters and Basta, 1996), little research has been done on the

87 75 effects of mixing WTR with biosolids prior to processing and dewatering at the wastewater treatment plant (referred herein as coprocessing). To examine the potential benefits of coprocessing, the following experimental objectives were identified. (1) Measure the Al ox content of three WTR samples from various stages of processing to compare the relative P-sorbing capacity of the WTR samples. (2) Evaluate the effect of BR on the WEP concentration in coprocessed biosolids. (3) Evaluate the effect of BR on the P concentration in the reject water from dewatering operations. (4) Examine the impact of the WEP results on P-Index scores and the resulting land application rates. 5.2 Methods and Materials Materials The wastewater solids were collected at the City of York Wastewater Treatment Plant (WWTP) located in York, PA. This facility was selected because it employs biological P removal (BPR) for nutrient control, and therefore, generates a biosolids with relatively high WEP concentrations. Samples of anaerobically digested sludge ( 1 to 1.5% solids) were collected from the sludge holding tanks prior to polymer addition and dewatering. Because the sludge at the point of collection meets the U.S. EPA pathogen and pollutant requirements for land application, hereafter the material is referred to as biosolids. The polyelectrolyte used by the York WWTP, Pollu-treat CL-455 from (Pollu-tech Inc.), was also used in the laboratory experiments. Pollu-treat CL-455 is a branched chain, cationic product that is commonly used for hard to dewater sludges. For laboratory experiments, a polymer solution was prepared by diluting concentrated polymer with distilled water to a concentration of 0.2% (1 ml polymer / 499 ml water). The diluted solution was

88 76 slowly mixed for 45 min until no visible clumps of concentrated polymer remained. Polymer solution was used within 48 h of preparation. The City of York WWTP flow schematic, sludge processing procedures, and sludge chemical characteristics are described in Chapter 3. Alum-based WTR (Al-WTR) samples were collected at the York Water Company s (YWC) Grantley Road facility, located in York County, Pennsylvania. The Al-WTR is a composite of solids from the sedimentation basins and filter backwashing operations. Samples used in coprocessing experiments were collected after the thickening tank (~ 2 to 4% solids) prior to chemical addition and dewatering. WTR samples used for oxalate-extractable Al (Al ox ) determination were collected from three locations representing WTR at various stages of processing. Samples were collected from the bottom of the sedimentation basins, storage lagoons and stock piles. Sampling procedures are described in Chapter 3. Samples were prepared and analyzed for Al ox using the method of McKeague and Day (1993), as modified for WTR by Dayton and Basta (2005a). The YWC Grantley Road WTP flow schematic, WTR processing procedures, and WTR chemical characteristics are described in Chapter Coprocessing Procedure Total solids, specific gravity and ph were measured on the WTR and biosolids samples in accordance with standard procedures (APHA et al., 1998). Volumes of Al-WTR were mixed with 200 ml of biosolids to give various blending ratios (BR), defined as the mass ratio of Al- WTR to biosolids on a dry weight basis. Dried WTR samples were crushed with a mortar and pestle and sieved to < 5 mm. The blended samples were mixed for 1 h on a reciprocating shaker at 200 evolutions per min (EPM). Mixed samples were removed from the mixer and placed in 500 ml jar-testing beakers. For coprocessing experiments in which polymer was added, a 0.2%

89 77 polymer solution was injected into the WTR/ biosolids blends. Samples were placed on the jar tester at 200 rpm and polymer solution was injected with a syringe into the vortex of the mixing zone approximately 2 cm below the liquid surface. Mixing continued at 200 rpm for 20 s and then was reduced 60 rpm for an additional 90 s. Mixing speeds and duration were selected so that the product of the velocity gradient (G) and detention time (t) was approximately 10,000 (dimensionless). For the dewatering experiments, samples were centrifuged for 20 min at 3,500 rpm to thicken the solids fraction to approximately 5 to 10%. This initial dewatering step was necessary prior to belt-filter pressing because the bench-top belt-press simulator used in experiments cannot accommodate liquid slurries. The centrate was immediately coarse-filtered using Whatman #1 paper to remove floating debris, adjusted to ph 2 with HCl for preservation, and retained for subsequent P analysis. This liquid was considered compositionally representative of the liquid discharged during dewatering, hereafter called reject water. Thickened solids were dewatered using a bench-top Crown Press Belt Filter Press (BFP) Simulator. Samples were pressed for 4 min at 200 psi to achieve solids content between 18 and 25%, typical of full-scale BFP dewatering operations. The BFP cake samples were then stored in airtight containers to maintain a consistent moisture content until analysis Phosphorus Analysis Randomly selected filtrate samples were passed through a 0.45 µm filter (commonly used to distinguish suspended and dissolved solids) in order to quantify the particulate and soluble fractions of TP in the filtrate. Results are presented in Figure 6.4.

90 78 Figure 5.1. Comparison of dissolved and total P in BFP filtrate samples. Analysis using a one-way ANOVA (α = 0.05) yielded a p value < 0.001, indicating no statistically significance difference between filtrate TP and DP when dewatering the liquid biosolids alone or the WTR / biosolids blend. Based on this result, it is concluded that the amount of particulate P in the samples after 11 μm filtration is negligible and thus samples did not need to be digested prior to analysis of TP. The total P (TP) content of the reject water was analyzed using inductively coupled plasma (ICP) atomic emission spectroscopy. BFP cake samples were analyzed for WEP using the 100:1 WEP ratio procedure developed by Kleinman et al. (2007). Samples were measured for total solids in accordance with standard procedures (APHA et al., 1998). A dry weight equivalent of 2.0 g of samples was extracted and mixed with sufficient deionized water to a achieve a final