UNESCO-IHE INSTITUTE FOR WATER EDUCATION

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1 UNESCO-IHE INSTITUTE FOR WATER EDUCATION Coupling enhanced biological phosphorus removal (EBPR) and magnesium ammonium phosphate (MAP) through ozonated sludge for nutrient recovery Jorge Guillermo Merizalde-Dobles MSc Thesis (MWI ) April 2011

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3 Coupling Enhanced Biological Phosphorus Removal (EBPR) and magnesium ammonium phosphate (MAP) through ozonated sludge for nutrient recovery Master of Science Thesis by Jorge Guillermo Merizalde-Dobles Supervisor Prof. Dr. D. Brdjanovic (UNESCO-IHE) Mentors Dr. Carlos M. López-Vázquez (UNESCO-IHE) Dr. Mariska Ronteltap (UNESCO-IHE) Dr. Devendra Saroj (University of Surrey, UK) Examination committee Prof. Dr. Damir Brdjanovic (UNESCO-IHE), Chairman Dr. Carlos M. López-Vázquez (UNESCO-IHE) Ir. Rogier van Kempen (Delfluent Services BV) This research is done for the partial fulfillment of requirements for the Master of Science degree at the UNESCO-IHE Institute for Water Education, Delft, the Netherlands Delft April 2011

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5 The findings, interpretations and conclusions expressed in this study do neither necessarily reflect the views of the UNESCO-IHE Institute for Water Education, nor of the individual members of the MSc committee, nor of their respective employers.

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7 To Sara and Ana Lucía, for they are my strength.

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9 Abstract Sludge ozonation was investigated in the laboratory as a means of nutrient recovery in the context of an activated sludge (AS) wastewater treatment plant (WWTP) performing enhanced biological phosphorus removal (EBPR). Experiments for different ozone doses were carried out and the resulting supernatants were fed as carbon source to induce anaerobic phosphorus release (APR) in batch tests. Volatile fatty acid (VFA) exhaustion -rather than production- and nitrate accumulation were observed with increasing dose when ozonating continuously, and deemed as the cause for the low APR kinetic rate observed when using ozonated supernatant as substrate: 78±6 and 100±8 mgp/gcod PAO /d for ozonation doses of 86±7 and 118±9 mgo 3 /gvss respectively. In contrast, no significant amounts of nitrate accumulated when ozonating intermittently in 60-minute steps, up to 6 hours. Calculations based on analysis of experimental samples showed struvite (magnesium ammonium phosphate or MAP) as the solid phase closest to saturation, although precipitation would require mixing the anaerobic reactor's effluent supernatant with an ammonium-rich stream and/or chemical addition of magnesium at ph around 9.3. Keywords: Ozonation, activated sludge, EBPR, nutrient recovery, struvite, magnesium ammonium phosphate, MAP, nitrate, VFA. J. G. Merizalde-Dobles i

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11 Acknowledgements I would like to express my gratitude to my mentors, Dr. Carlos M. López-Vázquez and Dr. Mariska Ronteltap for their ideas, criticism, guidance and support throughout my journey; to my supervisor Prof. Damir Brdjanovic and to Dr. Devendra Saroj for their input and encouragement, as well as to all those lecturers in IHE that have made our learning possible by challenging our thinking and sharing their knowledge and experience. I am extremely grateful to Ferdi Battes, Peter Heerings and all the laboratory staff at IHE that have had the kindness to allow us in their sacred space and the patience to bear our learning mistakes; as well as to Laurens Welles, Becky de Vera, Henock Belete Asfaw, Ervin Buçpapaj and all those fellow laboratory MSc researchers who were always willing to lend a hand when I needed it. My gratitude goes to the Instituto Costarricense de Acueductos y Alcantarillados in Costa Rica and to the Joint Japan/World Bank Graduate Scholarship Program for believing in me and making this journey materially possible. Finally, I would like to thank my parents, wife and daughter for all their love and unconditional support. J. G. Merizalde-Dobles iii

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13 Table of contents Abstract... i Acknowledgements... iii Table of contents...v List of tables... vii List of figures... viii List of abbreviations... ix 1 INTRODUCTION Literature Review Sludge reduction and ozonation Enhanced Biological Phosphorus Removal (EBPR) Nutrient recovery Struvite precipitation A problem, an opportunity Ionic fractions Saturation and crystallization The effect of ionic strength Ion activity product Problem statement Research objectives METHODOLOGY Collection and storage of sludge samples Measurements Ozonation Experiments Laboratory setup Procedure Continuous and intermittent ozonation Anaerobic Phosphorus Release (APR) Experiment Setup Procedure Pretreatment (RBCOD and NO X depletion) Batch test Chemical Calculations RESULTS AND DISCUSSION Ozonation Experiments Ozone dose COD solubilization Volatile fatty acids COD mass balance TKN solubilization Nitrogen mass balance Nitrate accumulation Ammonium oxidation TP solubilization ph change COD ratios Solubilization of cations APR batch tests Nitrate and ph rise J. G. Merizalde-Dobles v

14 3.2.2 Phosphate release Nutrient Recovery From ozonated sludge filtrate From APR batch test filtrate Optimum ph Required cation concentration Ammonium as limiting ion Overall assessment CONCLUSIONS AND RECOMMENDATIONS REFERENCES vi MSc thesis

15 List of tables TABLE 1. CLASSIFICATION OF THE DIFFERENT ANALYSES PERFORMED IN THIS RESEARCH TABLE 2. SOLID AND GASEOUS PHASES CONSIDERED FOR CALCULATION TABLE 3. CHANGE IN PH OF FRESH SLUDGE DURING APR BATCH TESTS TABLE 4. ANAEROBIC PHOSPHATE RELEASE FOR THE BATCH TESTS AND FROM LITERATURE TABLE 5. SOLUTION USED FOR OPTIMUM PH CALCULATION AT 20 C J. G. Merizalde-Dobles vii

16 List of figures FIGURE 1. PHOSPHATE, AMMONIUM AND MAGNESIUM IONIZATION FRACTIONS FIGURE 2. STRUVITE BOUNDARY CONDITIONAL SOLUBILITY PRODUCT PS 0 (SNOEYINK AND JENKINS, 1980) FIGURE 3. WWTP LAYOUT WITH UCT CONFIGURATION MODIFIED TO ACCOMMODATE SLUDGE OZONATION FOR NUTRIENT RECOVERY FIGURE 4. SETUP FOR SLUDGE OZONATION EXPERIMENTS FIGURE 5. PROCEDURE SCHEME FOR THE ANAEROBIC PHOSPHORUS RELEASE BATCH TEST FIGURE 6. OZONE DOSE VS. OZONATION TIME FIGURE 7. SOLUBLE CHEMICAL OXYGEN DEMAND OBSERVED AT DIFFERENT OZONE DOSES FIGURE 8. VOLATILE FATTY ACIDS CONCENTRATION VS. OZONE DOSE FIGURE 9. COD MASS BALANCE FOR DIFFERENT INTERMITTENT OZONATION DOSES (RUN 2) FIGURE 10. NITROGEN MASS BALANCE FOR DIFFERENT INTERMITTENT OZONATION DOSES FIGURE 11. NITRATE CONCENTRATION VS. OZONE DOSE FIGURE 12. AMMONIUM CONCENTRATION VS. OZONE DOSE FIGURE 13. PHOSPHORUS MASS BALANCE FOR DIFFERENT OZONATION DOSES FIGURE 14. ORTHOPHOSPHATE CONCENTRATION VS. OZONE DOSE FIGURE 15. CHANGE IN PH FOR DIFFERENT OZONE DOSES FIGURE 16. SOLUBLE COD TO SOLUBLE TKN RATIO FOR DIFFERENT OZONE DOSES FIGURE 17. SOLUBLE COD TO NITRATE RATIO FOR DIFFERENT OZONE DOSES FIGURE 18. SOLUBLE COD TO ORTHOPHOSPHATE RATIO FOR DIFFERENT OZONE DOSES FIGURE 19. CATION CONCENTRATIONS VS. OZONE DOSE FIGURE 20. ANAEROBIC PHOSPHORUS RELEASE BATCH TEST # FIGURE 21. ANAEROBIC PHOSPHORUS RELEASE BATCH TEST # FIGURE 22. SATURATION INDEXES OF THE OZONATED SLUDGE FILTRATE AT DIFFERENT OZONATION DOSES FIGURE 23. SATURATED PHASES OF THE APR BATCH TEST FILTRATES FIGURE 24. SATURATION INDEXES OF SOLID PHASES FOR DIFFERENT PH VALUES AT 20 C FIGURE 25. SATURATION INDEXES OF SOLID PHASES AT PH=9.3 FOR DIFFERENT APR KINETIC RATES FIGURE 26. ACTUAL AND MINIMUM REQUIRED IAP FOR DIFFERENT APR KINETIC RATES FIGURE 27. SATURATION INDEXES FOR DIFFERENT AMMONIUM CONCENTRATIONS viii MSc thesis

17 List of abbreviations AAS Atomic Absorption Spectrometry AES Atomic Emission Spectrometry a.k.a. Also known as APR Anaerobic phosphorus release AS Activated sludge ATU Allylthiourea (C 4 H 8 N 2 S) BNR Biological nutrient removal BCOD Biodegradable COD = RBCOD + SBCOD C Carbon COD Chemical oxygen demand COD PAO COD of PAO biomass DO Dissolved oxygen DS Dry solids EBPR Enhanced biological phosphorus removal FCOD Fermentable COD f CV COD to VSS ratio f PAO/VSS PAO biomass fraction of VSS GC Gas chromatography HAc Acetic acid: CH 3 COOH HP Harnaschpolder (the WWTP) HvH Hoek van Holland (the WWTP) IAP Ion activity product IC Ion chromatography ISS Inorganic suspended solids K Potassium Ksp Solubilization product "constant" MAP Magnesium ammonium phosphate a.k.a. struvite Mg Magnesium N Nitrogen N 2 (Di)nitrogen gas N/DN Nitrification / denitrification + NH 4 Ammonium - NO 2 - NO 3 Nitrite Nitrate NO x Nitrite and/or nitrate O 3 Ozone OHO Ordinary heterotrophic (micro)organisms OrgN Organically-bound nitrogen = SON + XON op 3- Orthophosphate PO 4 OS Ozonated sludge P Phosphorus PAO Phosphate-accumulating (micro)organisms PHA Polyhydroxyalkanoates 3- PO 4 Orthophosphate polyp Polyphosphate RBCOD Readily biodegradable COD = VFA + FCOD SBCOD Slowly biodegradable COD SI Saturation index = Log(IAP) - Log(Ksp) SON Soluble OrgN SS Suspended solids TKN Total Kjeldahl nitrogen + = OrgN + NH 4 TP Total phosphorus TSS Total suspended solids USGS United States Geological Survey VFA Volatile fatty acids VSS Volatile suspended solids WAS Waste activated sludge WWTP Wastewater treatment plant X COD Particulate COD Particulate OrgN X ON J. G. Merizalde-Dobles ix

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19 1 Introduction Biological oxidation is the most economical method of treatment for sewage; the most commonly used biological oxidation technology is the so-called activated sludge process (Weijers, 2000). However, round-the-clock aeration and high production of excess sludge make running costs for activated sludge (AS) systems high. Treatment, handling and disposal of waste activated sludge (WAS) comprise up to 60% of the total operational costs (Zhang et al., 2009). This significant amount can be cost-effectively reduced by implementation of sludge reduction technologies. Besides extended aeration, sludge reduction methods rely on lysis (cell destruction) of a considerable fraction of the biomass by physical, chemical or biological means, which bring intracellular contents out to the bulk liquid (solubilization). Advanced oxidation methods include ozonation, which can cause direct and indirect damage to the cell wall by forming hydroxyl radicals, thus solubilizing the organics and minerals intracellularly contained in the sludge biomass. On the other hand, in most industrialized countries nutrient removal is mandatory for facilities treating municipal wastewater. If readily biodegradable chemical oxygen demand (RBCOD) is limiting, biological nutrient removal (BNR) systems would necessitate an external source of carbon to achieve the required effluent concentrations for nitrogen and phosphorus. The RBCOD-rich supernatant stream produced by ozonation can aid the BNR processes in the plant by avoiding the need for an external carbon source. About 60% of the COD released during ozonation is biodegradable (Chu et al., 2009). However, solubilized intracellular nutrients need to be removed to avoid their recirculation in the system. Since this supernatant stream is small and relatively concentrated, struvite precipitation could cost-effectively achieve nutrient recovery, although magnesium is usually limiting and ph needs to be raised above 8 or 9 for optimal precipitation (Saidou et al., 2009). Finally, ozonation has the potential to cost-effectively reduce excess sludge volumes and enhance BNR in a full-scale AS system treating municipal wastewater. The implications of integrating WAS reduction by ozonation -for production of RBCOD and nutrient recovery through struvite precipitation- into an enhanced biological phosphorus removal (EBPR) system were investigated during the course of this research. J. G. Merizalde-Dobles 3

20 1.1 Literature Review In the late 19 th century, water-borne sanitation came as a response to overcrowding cities in the industrializing Western World and the subsequent public health impacts. By means of sewers, the pollution from human excreta was moved out of the cities and into the surface waters. The increasing pollution loads pushed natural systems beyond their selfpurification capacities, killing higher organisms due to low dissolved oxygen (DO) concentrations, which forced the developed counties to create disposal regulations that grew stricter with time. At that time, the easiest, most effective way of oxidizing the organic matter in sewage was by means of cultivation of ordinary aerobic microorganisms -already present in domestic wastewater- by providing aeration and enough residence time to allow transformation of most incoming biodegradable matter into biomass and carbon dioxide (the principle of the activated sludge process). A very important downside to AS technology is the high growth yield of ordinary heterotrophic organism (OHO) biomass, which needs to be treated and disposed of. In developed countries, environmental and regulatory restrains limit options for sludge disposal to landfilling or incineration with very high inherent costs. Treatment before final disposal must remove a considerable fraction of the sludge water content by thickening (gravitational separation), digestion (biological treatment), dewatering (chemical/mechanical treatment) and drying (thermal treatment). Obviously, this constitutes a significant operational (i.e. financial) burden (Zhang et al., 2009), hence the need for effective ways of minimization. 4 MSc thesis

21 1.1.1 Sludge reduction and ozonation Pérez-Elvira et al. (2006) categorize sludge reduction measures in three groups. The first: introduction of lower-yield stages in the wastewater process (lysis-cryptic growth, uncoupling and maintenance metabolism, predation, anaerobic treatment). The second: enhancing the sludge digestion process by either pretreatment (physical, chemical or biological), or modification of the digestion configuration. The third: removal (incineration, gasification, advanced oxidation). Ozonation is a sort of advanced oxidation widely applied for disinfection. Ozone (O 3 ) is an unstable -thus highly oxidative- gas that can be produced by applying an electrical discharge to pure oxygen or even air. It has the potential to destroy bacterial cell walls (Yan et al., 2009) causing solubilization and mineralization of its contents. Cell damage is the most important mechanism of sludge disintegration. Sludge disintegration to ozone dose is a direct non-linear relationship. There is a dosing threshold between 25 and 50 mg O 3 /g DS beyond which ozone penetrates bacterial cell walls causing their breakage (Zhang et al., 2009). When ozone is applied for WAS reduction it makes sense to maximize solubilization and minimize mineralization, thus creating a source of easily biodegradable carbon that could be used to enhance BNR performance of the AS system. Furthermore, rupture of the cell wall should release nutrients (N, P) and other minerals (Mg, K, etc.) to the bulk liquid, making them available for recovery Enhanced Biological Phosphorus Removal (EBPR) It was only until the 1960's that the role of nutrients (N, P) in eutrophication was understood, thus chemical and biological removal technologies for wastewater streams were developed (Henze et al., 2008). Bacterial removal of phosphorus beyond metabolic needs -i.e. other than biomass formation- was first observed in the 1950's, which led to the development and spread application of enhanced biological phosphorus removal (EBPR) processes (Henze et al., 2008). Phosphate accumulating organisms (PAO) are present in the activated sludge of EBPR systems. In anaerobic conditions, PAO uptake VFA to replenish their intracellular stores of glycogen and PHA, while excreting phosphate from their poly-p reserves. Under aerobic conditions, PAO take up and accumulate phosphate in their intracellular polyphosphate (poly- P) reserves. Since WAS is usually discarded from the aerobic part of an AS plant, phosphorus leaves the system inside the replenished poly-p reserves of PAO in the WAS; meanwhile, a negligible portion exits via the effluent total suspended solids (TSS). J. G. Merizalde-Dobles 5

22 1.1.3 Nutrient recovery Verstraete et al. (2009) estimate that wastewater as a resource is worth 0.35/m 3, out of which nutrients have a value of 0.02/m 3. Since 2003, prices of phosphate fertilizers are linearly rising; the price of nitrogen fertilizer produced by fixation of atmospheric nitrogen has also increased. Therefore, nutrient recovery and minimization of diffuse emissions (e.g. N 2 from N/DN, CO 2 from oxidation) are necessary for the sustainability of wastewater treatment (Verstraete et al., 2009). Recovery of phosphorus can be performed at different locations within an EBPR system. In a PhoStrip configuration, VFA are added to a fraction of the return sludge to encourage phosphorus release then the P-rich supernatant is subjected to chemical precipitation (Henze et al., 2008). In a BCFS process, phosphorus is chemically precipitated from the supernatant of mixed liquor taken from the anaerobic reactor; in this case, phosphorus can be precipitated with iron, aluminum, lime or magnesium. Phosphate-rich sludge (obtained from P precipitation) can be alkalized and the released phosphorus can be transformed into calcium phosphate, which is a raw material for industrial phosphorus processes (Verstraete et al., 2009). Although maximal recovery might seem sensible, increased operation of the P-stripper in a BCFS system could make phosphorus unavailable to PAO or even limit autotrophic growth, which would destabilize the EBPR process or deteriorate nitrification, respectively (Barat and van Loosdrecht, 2006). Therefore, recovery of nutrients must abide to the restrictions of the main biological processes in the WWTP. 6 MSc thesis

23 1.1.4 Struvite precipitation A problem, an opportunity In wastewater treatment practice, struvite deposits are well known as an important operational problem, especially when treating sludge from BNR systems, mostly after anaerobic digestion. In relatively short time, pipes can be clogged by struvite deposits. Factors influencing the occurrence of this problem are low pressure -causing CO 2 to off-gas-, pipe roughness, low suspended solids, and high surface-to-volume ratios (Doyle et al., 2000). It must be noted that struvite formation and struvite deposition are not necessarily the same. Formation starts if the right ph and ion activities are given, and it is even considered beneficial since it might aid EBPR (by increasing the VFA-to-P ratio). Struvite crystals form in two steps: nucleation and growth, the former being rate limiting. Suspended solids provide a seeding core for struvite crystals to grow without the need of forming nuclei. In a stream with high suspended solids, crystals are washed away with the rest of the solids. Therefore, struvite deposits occur more frequently on pipe and equipment surfaces where SS are low. In this case, roughness (even at micro-scale) provides the seeding effect. This means that pipes conveying filtrate (or centrate) are more likely to suffer from struvite deposits than sludge lines (WEF, 2006). Struvite deposits are especially common at the suction side of pumping stations, where the low pressure strips CO 2 out of solution, yielding an increase in ph. Deposits are avoided either by ph decrease or reduction of at least one constituent. Operational means usually involve both measures and are limited to chemical addition (WEF, 2006). Struvite recovery could represent a more attractive solution to the problem, since it has the potential to generate an income at the same time that it avoids clogging. According to Shu et al. (2006) and Roeleveld et al. (2004), approximately 1 kg of struvite can be crystallized from 100 m 3 of wastewater, which in Japan is commercialized at 250/ton dry solids as fertilizer (as cited by Verstraete et al., 2009), although struvite production from wastewater might cost between 220 and 2750/ton (Doyle and Parsons, 2002). Moreover, pretreatment by struvite precipitation has proven to enhance BNR in an AS system treating swine wastewater (high ammonia concentration) by increasing the carbon-tonitrogen and carbon-to-phosphorus ratios (Ryu and Lee, 2010). Available methods of struvite precipitation are: mechanical stirring, magnetic stirring, aeration, electromechanical precipitation, ion exchange and dissolved CO 2 degasification (Saidou et al., 2009). J. G. Merizalde-Dobles 7

24 Ionic fraction (a) Ionic fractions In order to estimate the potential for struvite precipitation in any given solution, the ionic speciation of all constituents need to be known. Equilibrium constants for all involved ions are shown in the following equations: (Snoeyink and Jenkins, 1980) With this information, ionic fractions can be calculated, as plotted in Figure 1. It can be noticed that the Mg 2+ species is dominant for the expected WAS ph range (6-8). However, phosphate and ammonium peaks are apart from each other, leaving only a ph window between 10 and 11.5 with considerably low ionic fractions in which struvite could be "optimally" precipitated. 100% 90% 80% 70% 60% 50% 40% 30% 20% 10% 0% H3PO4 H2PO4- HPO42- PO43- NH4+ NH3 Mg2+ MgOH- ph Figure 1. Phosphate, Ammonium and magnesium ionization fractions. 8 MSc thesis

25 Saturation and crystallization Struvite precipitation could occur only if the concentrations of the constituents are such that they exceed the minimum -boundary- conditional solubility product, i.e. Ps > Ps 0. The calculation follows: Ps = C T,Mg C T,NH3 C T,PO4 > Ps 0 Where C T,i is the (total) analytical concentrations of constituent i (including all its related species). Nevertheless, if the concentration of a solution is gradually increased above the conditional solubility product (Ps 0 ) of struvite, nucleation will not occur for some time until a supersaturation threshold has been crossed, since nuclei require a minimum activation energy to start (Stumm and Morgan 1996). On the other hand, the Ps "constant" or boundary conditional solubility varies with ph (see Figure 2), since the precipitation potential is dependent upon the actual concentrations of the ion species that constitute the struvite crystal (i.e. Mg 2+, NH +4, PO 4 3- ), which varies with ph, as seen in Figure 1. Figure 2. Struvite boundary conditional solubility product Ps 0 (Snoeyink and Jenkins, 1980). J. G. Merizalde-Dobles 9

26 The effect of ionic strength Ionic strength ( or IS) is a measure of the electrical forces among charged particles (ions). Struvite precipitation will depend on the actual activity of its constituent ions, rather than their analytical concentration. When the influence of other ions is negligible, the effect of ionic strength can be dismissed and activity will be equal to concentration. However, in very concentrated solutions, the apparent activity of dissolved substances in reactions is lower than expected for a given measured (analytical) concentration. The more ions in a solution, the stronger their electrostatic interactions, which in turn affect the chemical equilibrium (Snoeyink and Jenkins 1980). In thermodynamic terms, this is called non-ideal behavior of ions and molecules in solution. In other words, the ions involved in struvite crystallization are hampered by all other ions present in the solution. Ionic strength ( ) can be calculated by: Where C i and Z i are the concentration and valence of ion i, respectively. 10 MSc thesis

27 Ion activity product For high concentration solutions, concentrations need to be corrected by an activity factor before calculating chemical equilibria. The activity factor for a particular ion ( i ) can be estimated using the DeBye-Hükel limiting law as follows: log 2 i 0.5 Z i 1 2 This way, activities of the struvite constituents (between braces) can be determined in function of the corresponding ion concentrations (between brackets) by the expressions: Hence, by combining the ionic fraction and ion activity effects, the ion activity product (IAP) can be estimated, as follows: Where a i and i are the ionic fraction and activity coefficient of species i, respectively. Therefore, when precipitation potential is to be calculated for a high concentration solution, the IAP must be computed and compared to the solubility product (Ksp) of the equilibrium phase of interest. The degree of saturation of a solution can be expressed by the saturation index (SI), as follows: SI = Log(IAP) - Log(Ksp) Finally, for precipitation to occur, the saturation index (SI) needs to be greater than zero. J. G. Merizalde-Dobles 11

28 1.2 Problem statement Expenses for treatment and disposal of waste activated sludge comprise a substantial fraction of total operating costs in a WWTP. Ozonation has the potential to reduce WAS volumes while solubilizing some of the intracellularly stored organic and mineral compounds. If the COD solubilized by sludge ozonation is suitable for BNR, further savings could be achieved by avoiding the need for an external carbon source and enhancing the anaerobic phosphorus release. Moreover, these additional ions released to the liquid phase during ozonation (PO 4 3-, Mg 2+, Ca 2+, K + ) could be later precipitated for recovery. The context of the proposal described in this research report is that of an AS WWTP performing biological nitrogen and phosphorus removal, as depicted in Figure 3. Figure 3. WWTP layout with UCT configuration modified to accommodate sludge ozonation for nutrient recovery. The costs and benefits inherent to implementing sludge ozonation for nutrient recovery are summarized as follows: Costs Benefits Implementation (capital) investment 42% less WAS 1 Energy to produce ozone 31% less WWTP operational costs 1 More complex WWTP layout Nutrient recovery Chemical addition (Mg) Lower C-source costs 1 Inchauste-Daza (2010) 12 MSc thesis

29 1.3 Research objectives 1. To define the optimal ozone dose for maximized anaerobic phosphorus release by PAO as an indicator of the availability of RBCOD suitable for anaerobic uptake by PAO. 2. To assess the conditions and potential for precipitation of MAP from ozonated sludge. 3. To assess the conditions and potential for precipitation of MAP from the anaerobic reactor's supernatant, after being fed with ozonated sludge (to increase the anaerobic phosphorus release). 4. To quantify the potential of using other intracellular elements excreted during the anaerobic phosphorus release of PAO (such as Mg 2+ and K 2+ ) for struvite precipitation, as an additional resource recovery measure derived from WAS ozonation. J. G. Merizalde-Dobles 13

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31 2 Methodology This research was mainly carried out in the laboratory. WAS collected from the Hoek van Holland WWTP was subjected to ozonation in order to cause lysis and solubilization of intracellularly stored compounds, particularly of RBCOD to promote biological P-removal. Next, the filtrate from the ozonated sludge was anaerobically fed to a batch of activated sludge, as carbon source to encourage phosphorous release by PAO. The aim of these experiments was to find an optimal ozone dose for maximizing the induction of anaerobic phosphorus release when the ozonated filtrate was fed to the APR batch test. The next step was then to assess the optimum conditions to favor the precipitation of struvite and other minerals that would allow for nutrient recovery from the supernatant resulting from the APR batch tests. Finally, ozonation experiments were performed in both continuous and intermittent fashion. A first set of continuous experiments provided the supernatant for use in the APR batch tests. The second set, comprised of intermittent ozonation experiments was used for characterization of the ozonated sludge and mass balance calculations. J. G. Merizalde-Dobles 15

32 Parameter Gravimetry Spectrophotometry Digestion Distillation Atomic Spectroscopy Gas Chromatography Fresh Ozonated APR batch 2.1 Collection and storage of sludge samples The activated sludge used in the experiments was collected weekly from the end of the aeration tank at Hoek van Holland (HvH) wastewater treatment plant. After settling for 30 minutes, a supernatant volume of two thirds of the total volume was discarded; thereby producing a solids concentration similar to that of return activated sludge (RAS). The fresh sludge was then stored for up to one week at 4ºC. Ozonated filtrate was used in the anaerobic phosphate release (APR) batch tests within 48 hours to avoid bacterial growth and change in its composition. 2.2 Measurements A complete list of analyses performed to the samples drawn from the different experiments in this research is given in Table 1. All analytical procedures were carried out in accordance to Standard Methods (American Public Health Association, 2005) as described in the UNESCO-IHE Laboratory Manual entitled "Environmental Chemistry, Selected Methods for Water Quality Analysis" (LN0168/09/1). Table 1. Classification of the different analyses performed in this research. Type Sludge Samples # 01 VSS X T T 02 COD X X S S&T S 03 PO 4 X S S S 04 TP X X T T 05 NH 4 X S S S 06 TKN X X X T&S T&S - 07 NO 3 X S S S 08 Al X S S 09 Ca X S S 10 Fe X S S 11 K X S S 12 Mg X S S 13 Na X S S 14 VFA X S S S T = total sludge, S = soluble (filtrate) 16 MSc thesis

33 ozone generator 2.3 Ozonation Experiments For this experiment, activated sludge is ozonated in a continuously mixed glass vessel where ozone-rich air is diffused in, and ozonation time is the only control variable. All excess ozone is either thermally converted back to oxygen before escaping to the atmosphere, or bubbled through a trap bottle containing a potassium iodide solution. Cumulative gas volumes are recorded by gauges connected to the exit of the trap bottles. Ultimately, an ozone dose can be calculated and expressed in milligrams of ozone per gram of suspended solids (mgo 3 /gss). The resulting sludge is characterized by analysis of the drawn samples, as described in section 2.2 Measurements Laboratory setup The same laboratory setup was used for all tests, and all batches contained 2000±20 ml of the 3-time-concentrated activated sludge (see section 2.1). Figure 4 shows a schematic diagram of the laboratory setup used. air off-gas off-gas volume gauge V air ozone destructor sludge vessel off-gas trap bottle w/ KI solution air 0 3 -rich air by-pass volume gauge V air by-pass by-pass trap bottle w/ KI solution Figure 4. Setup for sludge ozonation experiments. In detail, iodide (I - ) ions in the trap bottles are oxidized to elemental iodine (I 2 ) during the experiment, causing the development of color in the initially clear KI solution. The amount of ozone bubbled through a trap bottle (containing 200 ml of 40 g/l KI solution) is J. G. Merizalde-Dobles 17

34 later determined by titration with 0.1 M sodium thiosulphate (Na 2 S 2 O 3 ), which -under acidic conditions- will reduce iodine back to iodide, rendering a clear solution Procedure At the beginning and at the end of the experiment, the by-pass line is kept open for 10 minutes in order to later estimate the average ozone concentration supplied to the sludge vessel as follows: Ozone mass[mg] Ozone concentration = Air volume [L] The amount of off-gassed ozone is also determined by titration and subtracted from the input ozone to calculate the mass that effectively reacted with the sample in the vessel. After a TSS/VSS measurement of the fresh sludge, the ozonation dose can be estimated in mgo 3 /gss. Finally, the ozonated sludge is filtered to produce about half a liter of "substrate" to be fed to the APR batch test. For high ozonation doses, a 10 minute centrifugation at 4500 RPM is performed to facilitate filtration. Nitrate is measured by means of Merck strips and samples are taken for determination of VFA, soluble COD, anions (Mg, K, Ca, Fe, Na), orthophosphate and suspended solids. These analyses are carried out in accordance to the Standard Methods (American Public Health Association, 2005), by the means described in section Continuous and intermittent ozonation Two ozonation runs were carried out for this research. Run 1 consisted of three continuous ozonation tests, while run 2 comprised six intermittent ozonation experiments. In detail, the first run was composed of three experiments in which a two-liter sludge sample was continuously exposed to ozone for 90, 180 and 270 minutes. Fresh sludge was used for every experiment. Both, the fresh and ozonated sludge were sampled for testing, and the applied ozone dose was determined as explained before. On the other hand, for the second run a single two-liter sludge sample was used throughout the entire run, which consisted of 6 experiments. In every experiment, the sludge was exposed to ozone for 60 minutes. Then the reactor was opened for sampling, the trap bottle solution was retrieved and replaced, and the reactor was closed again for the next 60- minute ozonation cycle. This process was repeated 6 times to achieve a total of 360 minutes of ozone exposure, while sampling and calculating ozone doses at every intermediate step. Ozone dose, however was calculated in terms of the initial VSS concentration of the fresh sludge sample. 18 MSc thesis

35 2.4 Anaerobic Phosphorus Release (APR) Experiment In this experiment, filtrate from an ozonation experiment is fed under anaerobic conditions to a fresh sample of mixed liquor activated sludge, with the aim of providing VFA (and FCOD) to induce phosphorus release from PAO Setup A laboratory setup was put together at the Sanitary Engineering Laboratory of UNESCO-IHE. The equipment consists of a glass fermenter with a built-in water bath compartment connected to a BioConsole and a BioController from the manufacturer Applikon. The 2-liter fermenter has controlled inputs for: air, nitrogen, acid, base and substrate (peristaltic pumps); and parameter monitoring for: temperature, ph and DO Procedure During the experiments, temperature is kept at 20 C by use of the water bath, and nitrogen and air sparging are set at ~70 L/h to provide anaerobic and aerobic conditions, respectively. About 1 liter of fresh activated sludge is placed in the reactor, ph control and mixing at 250 RPM are started and 20 mg ATU are added to avoid ammonium oxidation during aeration. A conditioning pretreatment is provided before the actual test is performed Pretreatment (RBCOD and NO X depletion) First, the batch is tested for nitrate 2. If nitrate is detected, the batch is subjected to a 30- minute period of sparging with nitrogen gas (N 2 ) to induce biological denitrification. The nitrate test is repeated afterwards to confirm its absence. Then, a 60-minute aeration period follows in order to oxidize any remaining RBCOD in the sludge. Biological ammonium oxidation was inhibited by the previous addition of ATU. Therefore, it can be assured that the sludge is free of electron acceptors and that the only substrate present are the organics in the ozonated filtrate dosed during the phosphorus release test. 2 Nitrate concentrations are measured range-wise by use of Merck strips (type ). In order to measure, the strip is dipped in the solution for about 1 second, the excess water is shaken off and the result must be read after 60 seconds. The measuring strip has two pads, one for nitrite and the other one for nitrate, which change color according to the nitrite/nitrate concentration. The nitrite scale has three standard values to compare with: -, +, ++ (none, some, much). The nitrate pad is to be compared against a white-to-purple color scale with seven standard values: 0, 10, 25, 50, 100, 250, 500 mg N/L. J. G. Merizalde-Dobles 19

36 Finally, a filtered sample is taken for determination of the remaining soluble COD, which is assumed to correspond to the inert COD concentration Batch test After pretreatment, the conditioned batch is considered to be free of nitrate and RBCOD. Nitrogen sparging is then started to create anaerobic conditions in the batch reactor. A measured amount of ozonated filtrate (about 500 ml) is fed to the batch reactor; this is the beginning of the batch test. This test lasts 120 minutes, during which samples are taken every 15 minutes for VFA and orthophosphate measurements (see Figure 5). A time profile for both concentrations can be drawn to estimate the PAO activity induced during the test. ~1 L fresh activated sludge Yes NO 3 - in sludge? No ~0.2 L ozonated sludge filtrate Anaerobic Pretreatment 30' N 2 Aerobic Pretreatment 60' Aeration APR Batch Test 120' N 2 Inert SCOD [PO 4 3- ] & [VFA] Figure 5. Procedure scheme for the anaerobic phosphorus release batch test. 20 MSc thesis

37 2.5 Chemical Calculations For this purpose, the analytical concentrations were processed with PHREEQC 3 to determine chemical equilibria for the interest phases in the solutions (fresh sludge and sludge from ozonation and APR experiments). The considered saturation phases were: Table 2. Solid and gaseous phases considered for calculation Log # Phase Name State K sp h [kj/mol] 1 Ca 5 (PO 4 ) 3 OH Hydroxyapatite Solid Fe 3 (PO 4 ) 2 8H 2 O Vivianite Solid H 2 Hydrogen Gaseous KMgPO 4 Potassium struvite Solid NH 3 Ammonia Gaseous MgNH 4 PO 4 Struvite Solid O 2 Oxygen Gaseous PHREEQC (Version 2)--A Computer Program for Speciation, Batch-Reaction, One-Dimensional Transport, and Inverse Geochemical Calculations. United States Geological Survey. J. G. Merizalde-Dobles 21

38 22 MSc thesis

39 3 Results and Discussion 3.1 Ozonation Experiments In this section, the results from the sludge ozonation tests are presented. As explained in section , two ozonation test sets (runs 1 & 2) were performed. Run 1 was done continuously for each batch; Run 2 was done on one single two-liter batch in 1-hour steps i.e. intermittently, for up to six accumulated hours. Indeed, the aim of the first run was to produce substrate to feed the APR batch tests, and in that way, find the optimum ozonation dose in order to produce a substrate that would maximize anaerobic PAO activity. Nevertheless, unexpected results during the first ozonation run and APR experiments forced a change in strategy. Consequently, the aim of the second run was to confirm the findings of run 1, especially regarding nitrate accumulation, and at the same time, characterize the ozonated sludge and calculate mass balances for nitrogen, phosphorus and COD. Finally, the results are presented in charts, where horizontal and vertical error bars represent the standard error of the measurements for the parameters graphed in the X and Y axes respectively. J. G. Merizalde-Dobles 23

40 Ozone dose [mg / g VSS] Ozone dose The results from all ozonation experiments are presented in Figure 6, it can be noted that ozonation time and ozone dose have a clear direct relationship. O 3 dose vs. Time Ozonation time [minutes] R1: continuous R2: intermittent Quad. fit Figure 6. Ozone dose vs. ozonation time Even though ozonation time is much easier to measure, ozone dose has the advantage of evening out most experimental particularities, thus making the results more easily comparable and reproducible. Therefore, in the rest of this discussion ozonation results will be presented in relation to the ozone dose rather than the ozonation time. Finally, all ozone doses are stated in relation to the fresh sludge VSS concentration. This is particularly important to avoid any misunderstanding especially when expressing doses for intermittent ozonation experiments, as is the case for run MSc thesis

41 mgcod/l COD solubilization Solubilization of organic matter is one of the main goals of waste sludge ozonation (Zhang et al., 2009), since it reduces the amount of suspended solids while supposedly procuring a carbon source for BNR. The results from these experiments confirm that ozonation indeed releases organic matter from the sludge solids into the liquid phase. As shown in Figure 7, soluble COD was found to increase along with ozone dose. SCOD vs. O 3 dose R1: continuous R2: intermittent R2*: intermittent Ozone dose [mg/gvss] Linear Figure 7. Soluble chemical oxygen demand observed at different ozone doses Moreover, although not measured in this research, the soluble BCOD of ozonated sludge comprises around 80-98% of the soluble COD depending on the ozone dose, with an optimum around 130 mgo 3 /gtss (Inchauste-Daza, 2010). This would make the produced supernatant a suitable carbon source for BNR. Another important aspect to be noted is the fact that after the fourth ozonation of run 2, the resulting sludge was kept overnight at room temperature (~23 C) in the same vessel for about 15 hours; a sample was taken next morning just before performing the last two ozonation experiments. An important increase in SCOD was observed (Figure 7: green hollow triangle at 102 mgo 3 /gvss, ~3000 mgcod/l). This last fact suggests that sludge ozonation followed by a lag period might be more effective for solubilization of particulate organics than ozonation alone. One underlying mechanism could be that a higher efficacy is achieved for solubilizing organics when ozone is applied at lower doses for longer contact times, i.e. the small amount of remaining dissolved ozone had 15 hours of contact time with the sludge. Another possible explanation could be that ozonation facilitates the biological hydrolysis of particulates during the following lag period. Further research is needed to conclude on this fact, and to optimize the active/lag ozonation timing as well as the ozone concentration for maximizing the production of soluble BCOD. J. G. Merizalde-Dobles 25

42 mg VFA / L Volatile fatty acids VFA are required in order to induce anaerobic P-release. Therefore, ozonation is expected to produce either VFA or fermentable organics from which VFA could be generated. Despite the fact that increasing ozone dose produces more soluble COD, it seems that any VFA released during ozonation are immediately oxidized, as shown in Figure 8. For run 1, the first experiment was performed on fresh sludge with a very low starting VFA concentration (Figure 8: blue hollow diamond at 0 mgo 3 /gvss, VFA ~1 mg/l). This concentration remained very low even after ozonating for 90 minutes, indicating that no VFA accumulation was produced (Figure 8: blue hollow diamond at 34 mgo 3 /gvss, VFA ~1 mg/l). For the remaining two experiments in run 1, new activated sludge was collected, which had a considerably higher starting VFA concentration (Figure 8: blue diamond at 0 mgo 3 /gvss, VFA ~42 mg/l). Since the sludge is collected from the end of the aeration tank, it is not expected to contain any VFA. However, this fresh sludge was ozonated 4 days after collection from the WWTP, time during which VFA might have been produced by fermentation even though the sample was kept in the refrigerator the entire time. Nevertheless, after a 180-minute ozonation (86 mgo 3 /gvss), VFA went down by about 25%, and after a 270-minute ozonation (118 mgo 3 /gvss) VFA decreased by 90%. This reinforces the notion that any VFA that might be produced during the ozonation of activated sludge are immediately oxidized and thus do not accumulate in the liquid phase. VFA vs. O 3 dose R1: Continuous R1*: continuous R2: intermittent Ozone dose [mg / g VSS] Figure 8. Volatile fatty acids concentration vs. ozone dose Furthermore, the intermittent ozonation experiments of run 2 showed a 90% decrease in VFA concentrations after only 60 minutes of ozone exposure (14 mgo 3 /gvss), these concentrations remained at these very low levels even after higher ozone doses. 26 MSc thesis

43 Nevertheless, it must be noted that, as stated before, after the fourth intermittent ozonation of run 2, the resulting sludge was kept overnight at room temperature (~23 C) in the same vessel for about 15 hours. Therefore, the observed increase in VFA (Figure 8: green triangle at 102 mgo 3 /gvss, VFA 50 mg/l) is consistent with the observed increase in soluble COD reported before. However, a steep decrease in VFA concentration was observed in the next two ozonation steps, confirming the findings of the first ozonation run. This further underpins the notion that ozonation followed by a lag phase -rather than ozonation alone- is more effective in producing soluble COD, including VFA which are required for EBPR. On the other hand, further research is needed to characterize the produced COD by respirometry, as well as to optimize this particular destruction approach and determine the mechanism by which the provided lag phase facilitates BCOD solubilization and fermentation. As stated before, it could be hypothesized that during the lag phase, the remaining lower ozone concentrations combined with longer contact times are responsible for the observed effect. Another possible explanation is that the initial ozonation sets the stage for the enduring live biomass in the ozonated sludge to further hydrolyze and ferment the remaining particulate organics during the lag phase. In conclusion, no VFA accumulation was achieved by ozonation alone, whether continuous or intermittent (in 60-minute steps); in fact any VFA present in the fresh sludge were almost completely mineralized beyond a dose of 100 mgo 3 /gvss. J. G. Merizalde-Dobles 27

44 mgcod/l COD mass balance From the samples taken during run 2 (intermittent ozonation), COD was measured for total sludge and filtrate. The latter was considered to be equal to soluble chemical oxygen demand (SCOD), while the particulate chemical oxygen demand (X COD ) was determined by difference as: X COD Total COD SCOD In order to close the balance, the difference from the total fresh sludge COD (i.e. the balance) was regarded as COD lost as CO, CO 2, VFA and other volatile carbon species that off-gassed from the solution (see Figure 9). In summary, the balances account for 82-93% of the starting COD COD balance Off-gas SCOD XCOD Ozone dose [mg/gvss] Figure 9. COD mass balance for different intermittent ozonation doses (run 2) Finally, the COD solubilization trend observed in Figure 7 was confirmed by the particulate COD measurements expressed in this balance chart. Moreover, this is consistent with results from previous research reports (Inchauste-Daza, 2010). 28 MSc thesis

45 mg N / L TKN solubilization The main nitrogenous component in fresh activated sludge was found to be particulate organically bound nitrogen (X ON ). As a result, an OrgN solubilization trend was observed as the ozone dose increased, i.e. increasing S ON, decreasing X ON (see Figure 10). Moreover, total sludge TKN tended to decrease when increasing the ozone dose Nitrogen mass balance A nitrogen mass balance was prepared with the measurements from run 2. Similarly, in order to close the balance, it was assumed that the missing nitrogen was off-gassed from the solution perhaps as N 2 O, NO or N Nitrogen balance Ozone dose [mg / g VSS] Off-gassed NO3- NH4+ SON XON Figure 10. Nitrogen mass balance for different intermittent ozonation doses J. G. Merizalde-Dobles 29

46 mgno 3 - -N/L Nitrate accumulation During the first ozonation run, nitrate and nitrite were measured using Merck dip strips, which provide an estimation of the concentrations present in the water phase. Nitrite concentrations were found to be below detection limits for all samples. However, since nitrate accumulation was observed, the last two samples -which had the highest uncertainty-, were re-tested using the standardized method (APHA, AWWA and WEF, 2005). As depicted in Figure 11, nitrate accumulation was found to be directly related to ozone dose when ozonating continuously (run 1). In contrast, when ozonation was intermittent (run 2), no substantial nitrate accumulation occurred regardless of the dose. Nitrate vs. O 3 dose R1: continuous R2: intermittent Linear Ozone dose [mg/gvss] Figure 11. Nitrate concentration vs. ozone dose In fact, nitrate accumulation might be the most important impediment for the use of ozonated sludge filtrate as carbon source for APR, since PAO would lose their competitive advantage to uptake VFA in an anoxic environment. Moreover, depending on the amount of ozonated sludge recycled in a full-scale WWTP, nitrate could impair or even stop EBPR altogether. Henze et al. (2008) explain that for a recycled activated sludge (RAS) ratio of 1:1, phosphorus removal decreases linearly for increasing nitrate concentrations in the RAS flow from zero until about 11 mgn/l; at this point, PAO activity is completely impeded and P removal is only due to "wastage of sludge with normal metabolic P content (0.03 gp/gvss)" (Henze et al., 2008, p.202). In conclusion, the results of this research suggest that in a full-scale application RAS could only be ozonated at very low doses to avoid disrupting the EBPR process. Indeed, nitrate accumulation is another reason to explore an alternative low-dose long-contact-time approach to sludge ozonation, as suggested in section MSc thesis