Novel technologies for WWTP optimization in footprint, nutrients valorization, and energy consumption

UNIVERSIDADE DE SANTIAGO DE COMPOSTELA Departamento de Enxeñaría Química Novel technologies for WWTP optimization in footprint, nutrients valorization, and energy consumption Memoria presentada por: Nicolás Morales Pereira Para optar al grado de Doctor por la Universidade de Santiago de Compostela Santiago de Compostela, Enero de 2014 UNIVERSIDADE DE SANTIAGO DE COMPOSTELA Departamento de Enxeñaría Química Ramón Méndez Pampín, Catedrático de Ingeniería Química, José Luis Campos Gómez, Profesor Contratado Doctor de Ingeniería Química y Anuska Mosquera Corral, Profesora Titular de Ingeniería Química de la Universidad de Santiago de Compostela, Informan: Que la memoria titulada “Novel technologies for WWTP optimization in footprint, nutrients valorization, and energy consumption”, que para optar al grado de Doctor de Ingeniería Química, Programa de Doctorado en Ingeniería Química y Ambiental, presenta Don Nicolás Morales Pereira, ha sido realizada bajo nuestra inmediata dirección en el Departamento de Ingeniería Química de la Universidad de Santiago de Compostela. Y para que así conste, firman el presente informe en Santiago de Compostela, el 10 de enero de 2014. Ramón Méndez Pampín José Luis Campos Gómez Anuska Mosquera Corral i TABLE OF CONTENTS Resumen 1 Resumo 11 Summary 19 CHAPTER 1 26 INTRODUCTION 1.1 Water and wastewater in the 21st century 29 1.2 Water demand and sanitation as population grows 33 1.3 Water and energy 36 1.4 Nutrients in wastewater 44 1.4.1 Nutrients recovery 50 1.4.1.1 Urine separation 54 1.4.1.2 Fertilizer production by ammonia stripping from sludge liquid fraction 57 1.4.2 Nutrients removal by Anammox based processes 61 1.4.2.1 Conventional Nitrogen removal processes 61 1.4.2.2 The Anammox processes 62 1.4.2.3 CANON process 64 1.4.2.4 Application of Anammox processes to the main stream of WWTP 67 1.5 Biological compact treatment systems 68 1.5.1 Membrane systems 69 1.5.2 Biofilm systems 70 1.5.3 Aerobic granular systems 73 1.5.3.1 Advantages and drawbacks of aerobic granulation 74 1.5.3.2 Aerobic granules formation 76 1.5.3.3 Aerobic granular reactor configuration. Continuous or discontinuous? 77 1.5.3.4 Scale-up applications 81 1.6 Scope of this thesis 82 1.7 References 85 CHAPTER 2 100 MATERIALS AND METHODS 2.1 Analysis of the liquid phase 103 2.1.1 Chemical Oxygen Demand (COD) 103 2.1.2 Total Organic Carbon (TOC) 105 2.1.3 Nitrogen 105 2.1.4 Inorganic ions 109 2.1.5 Other control parameters 110 2.2 Biomass characterization 111 2.2.1 Solids concentrations (TSS and VSS) 111 Table of contents ii 2.2.2 Sludge Volume Index 111 2.2.3 Density of the granules 111 2.2.4 Average diameter of the granules 112 2.2.5 Elemental analysis 113 2.2.6 Poly-Hydroxy-Alkanoates (PHA) 114 2.2.7 Specific Anammox Activity assays 115 2.2.8 Respirometric assays 116 2.3 Microbiological determinations 117 2.3.1 Identification of bacteria populations by FISH 117 2.3.2 Reagents preparation 118 2.4 Calculations 120 2.4.1 SBR operational cycle analysis 120 2.4.2 Estimation of the nitrogen assimilated and denitrified. 121 2.4.3 Biomass production 122 2.4.4 Nitrogen removal rates in the Anammox Processes 123 2.5 References 124 CHAPTER 3 126 OPERATION OF AN AEROBIC GRANULAR PILOT SCALE SBR PLANT TO TREAT SWINE SLURRY 3.1 Introduction 129 3.2 Materials and methods 130 3.2.1 Reactor set-up and operational conditions 130 3.2.2 Analytical methods 132 3.2.3 Calculations 133 3.3 Results and discussion 134 3.3.1 Granule formation and physical properties 134 3.3.2 Organic matter and Nitrogen removal 138 3.3.3 Operational cycles 141 3.3.4 Microbial populations 144 3.4 Conclusions 148 3.5 References 149 CHAPTER 4 152 AEROBIC GRANULAR-TYPE BIOMASS DEVELOPMENT IN A CONTINUOUS STIRRED TANK REACTOR 4.1 Introduction 155 4.2 Materials and methods 156 4.2.1 Reactor description 156 4.2.2 Feeding composition 158 4.2.3 Inoculum 158 4.2.4 Operational conditions 158 4.2.5 Analytical methods 159 4.3 Results and discussion 160 4.3.1 Reactor performance 160 Table of contents iii 4.3.2 Biomass characteristics 163 4.3.3 Microbial Populations 169 4.4 Conclusions 172 4.5 References 174 CHAPTER 5 178 RECOVERY OF N AND P FROM URINE BY STRUVITE PRECIPITATION FOLLOWED BY COMBINED STRIPPING WITH DIGESTER SLUDGE LIQUID AT FULL SCALE 5.1 Introduction 181 5.2 Materials and methods 183 5.2.1 Urine separation, storage and treatment 183 5.2.2 Ammonia air stripping process including a CO2 pre-stripper 187 5.2.3 Chemical and crystals analysis 190 5.2.4 Chemical cost, electricity and fertilizer price 190 5.3 Results and discussion 191 5.3.1 Urine pre-treatment 191 5.3.2 Ammonia stripping with sludge liquid and treated urine 195 5.3.2.1 Experiment I 195 5.3.2.2 Experiment II 197 5.3.2.3 Removal efficiency and fertilizer recovery 197 5.3.2.4 Chemical and energy cost 198 5.4 Conclusions 199 5.5 References 201 CHAPTER 6 204 DYNAMICS OF THE CANON PROCESS TREATING LOW NITROGEN LOAD AT LOW TEMPERATURE 6.1 Introduction 207 6.2 Materials and methods 213 6.2.1 Reactors description 213 6.2.2 Cycle distribution 215 6.2.3 Influent 216 6.2.4 Inoculums 216 6.2.5 Analytical methods 217 6.2.6 Calculations 217 6.2.6.1 Removal rates 217 6.2.6.2 Biomass production 219 6.2.6.3 Oxygen depth penetration 219 6.2.6.4 Ammonia oxidation rate ratio 221 6.3 Results 221 6.3.1 Operation at 20 ºC and moderate nitrogen loading rates 222 6.3.2 Operation at 15 ºC and moderate nitrogen loads 227 6.3.2.1 Stage A-II 227 6.3.2.2 Stage A-IV 231 Table of contents iv 6.3.3 Operation at 15 ºC and 70 mg NH4+-N/L 232 6.3.4 Operation at 15 ºC and 50 mg NH4+-N/L with mechanical stirring 236 6.3.5 SBR 2 operation 238 6.4 Discussion 242 6.4.1 Biomass retention 245 6.4.2 Balance between AOB and Anammox activities 246 6.4.2.1 Importance of oxygen depth penetration in the process stability 246 6.4.2.2 Effect of granules size on the aerobic fraction of biomass 247 6.4.2.3 Biomass density effect on oxygen depth penetration 250 6.4.2.4 Ammonia oxidizing bacteria activity prediction 252 6.4.3 Nitrite Oxidizing Bacteria 254 6.4.3.1 NOB development 254 6.4.3.2 NOB repression 256 6.5 Conclusions 258 6.6 References 260 Conclusiones generales 265 Conclusións xerais 271 General conclusions 277 List of acronyms and symbols 281 List of publications 287 RESUMEN Y OBJETIVOS El continuo desarrollo de la sociedad está causando una gran presión sobre los sistemas naturales: incremento de la contaminación, deforestación, agotamiento de recursos, etc. El agotamiento de los recursos hídricos, así como la contaminación de los mismos, están entre los mayores retos a afrontar por la humanidad en el siglo XXI. Las tecnologías actuales de tratamiento de aguas residuales se desarrollaron a finales del siglo XIX y a lo largo del XX. Dichas tecnologías evolucionaron de simples alcantarillados, que vertían directamente en ríos y mares sin un mayor tratamiento del agua residual, hasta llegar a complejos sistemas de tratamiento que incluyen sistemas físicos, químicos, y biológicos; sistemas que consiguen eliminar sólidos, materia orgánica, así como los principales nutrientes. Actualmente, nuevas limitaciones ambientales, económicas y sociales obligan a desarrollar nuevos conceptos a la hora de afrontar el tratamiento de las aguas residuales. Dichos nuevos conceptos deben afrontar, además, aspectos como el crecimiento de la población mundial, el cambio climático o el agotamiento de los recursos hídricos, entre otros. Hasta ahora, las estaciones depuradoras de aguas residuales (EDAR) pueden ser definidas como sistemas donde la materia orgánica, el nitrógeno y el fósforo son eliminados del agua residual utilizando energía para ello. Sin embargo, este concepto deberá modificarse en los próximos años para lograr una mejora de los sistemas de tratamiento en términos de viabilidad económica y de reducción del impacto ambiental. De este modo, el agua residual ha de verse como una fuente renovable y recuperable tanto de agua, como de recursos y energía. Esto será factible gracias a la aplicación de nuevos sistemas y tecnologías de tratamiento. Además, los nuevos tratamientos de aguas residuales tendrán que requerir menor superficie para su implantación, y producir menor cantidad de sub-productos, como gases de efecto invernadero o lodo en exceso, comparados con los sistemas de tratamiento actualmente en uso. De acuerdo a esta nueva visión, la presente tesis se ha focalizado en el estudio de algunas alternativas de mejora de las EDAR, mediante la aplicación de nuevas tecnologías de Resumen y objetivos 8 El uso de microorganismos autótrofos de baja velocidad de crecimiento, como es el caso de las bacterias Anammox, implica la necesidad de usar un sistema caracterizado por una alta capacidad de retención de biomasa. En este caso, la retención de biomasa se acentuó mediante la utilización de biomasa en forma de gránulos cultivados en un sistema SBR. Las bacterias oxidantes de amonio se desarrollan en las capas más externas del gránulo, donde el oxígeno y amonio están presentes. Por otro lado, las bacterias Anammox crecen en capas más profundas, donde el oxígeno ya está agotado y el amonio y nitrito están disponibles para su consumo. El sistema CANON se operó inicialmente a 20 ºC alimentado con una concentración moderada de amonio (150-250 mg NH4+-N/L). Más tarde, la temperatura de operación se redujo a 15 ºC. Y finalmente, se alimentó el sistema con un influente con una baja concentración de amonio (50-80 mg NH4+-N/L), la cual simula las condiciones de operación que se pueden hallar en la línea de agua de una EDAR municipal donde se le haya eliminado previamente la materia orgánica. La retención de biomasa en el sistema SBR se consiguió de manera satisfactoria cuando el reactor se operó a 20 ºC. Se alcanzaron concentraciones de más de 12 g SSV/L en el reactor. Las tasas de eliminación de nitrógeno alcanzaron valores de 0,45 g N/L, alcanzando una eficiencia de eliminación de nitrógeno del 70%. Sin embargo, cuando el proceso se operó a 15 ºC, la baja tasa de crecimiento de la biomasa y las dificultades para controlar las bajas concentraciones de oxígeno disuelto en el medio líquido, hicieron necesario la utilización de diferentes configuraciones de reactor. Además, la baja concentración de oxígeno, temperatura y carga nitrogenada, facilitaron el desarrollo de un tercer grupo de microorganismos, las bacterias oxidantes de nitrito (BON). Las BON compiten con las Anammox por el nitrito, y con las BOA por el oxígeno. La eliminación de estos microorganismos no deseados supone el reto de la última parte de esta investigación. Para resolver estos problemas se probó la inclusión de un sistema de agitación mecánica, el uso de un reactor con una configuración de altura diámetro distinta y variaciones en el tiempo de sedimentación. Durante la experimentación se emplearon dos reactores SBR a escala laboratorio, con un volumen operativo de 1,5 y 4,0 L, respectivamente. A la hora de controlar el desempeño de las BOA y de las Anammox en un proceso CANON que opere a baja temperatura tratando corrientes con bajas concentraciones de nitrógeno se deben tener en cuenta tres factores principales: 1) alcanzar una alta retención de biomasa; 2) obtener un equilibrio entre las actividades de las bacterias oxidantes de amonio y de las bacterias Anammox; y 3) evitar el desarrollo de bacterias oxidantes de nitrito en la biomasa. Resumen y objetivos 9 En una EDAR, el reactor convencional de lodos activos podría ser sustituido por un reactor granular aerobio operando en continuo, donde se eliminaría la materia orgánica, mientras que el nitrógeno sería eliminado usando un proceso basado en Anammox. Se espera que las diferentes alternativas de tratamiento investigadas en esta tesis contribuyan al desarrollo de sistemas de tratamiento de aguas residuales más eficientes y sostenibles. RESUMO E OBXECTIVOS O continuo desenvolvemento da sociedade está a causar unha gran presión sobre os sistemas naturais: incremento da contaminación, deforestación, esgotamento de recursos, etc. O esgotamento dos recursos hídricos, así como a contaminación dos mesmos, están entre os maiores retos a afrontar pola humanidade no século XXI. As tecnoloxías actuais de tratamento de augas residuais desenvolvéronse a finais do século XIX e ó longo do XX. Estas tecnoloxías evolucionaron dende simples redes de sumidoiros que vertían directamente ós ríos e mares, sen un maior tratamento da auga residual, ata chegar a complexos sistemas de tratamento que inclúen sistemas físicos, químicos, e biolóxicos; sistemas que conseguen eliminar sólidos, materia orgánica, así como os principais nutrientes. Actualmente, novas limitacións ambientais, económicas e sociais obrigan a desenvolver novos conceptos á hora de afrontar o tratamento das augas residuais. Estes novos conceptos deben afrontar, ademais, aspectos como o crecemento da poboación mundial, o cambio climático ou o esgotamento dos recursos hídricos, entre outros. Ata o de agora, as estacións depuradoras de augas residuais (EDAR) poden ser definidas como sistemas onde a materia orgánica, o nitróxeno e o fósforo son eliminados da auga residual utilizando enerxía para iso. Con todo, este concepto deberase modificar nos próximos anos para lograr unha mellora dos sistemas de tratamento en termos de viabilidade económica e de redución do impacto ambiental. Deste xeito, a auga residual terá que verse como unha fonte renovable e recuperable, tanto de auga, coma de recursos e enerxía. Isto será factible grazas á aplicación de novos sistemas e tecnoloxías de tratamento. Ademais, os novos tratamentos de augas residuais terán que requirir menor superficie para a súa implantación, e producir menor cantidade de sub-produtos, coma gases de efecto invernadoiro ou lodo en exceso, comparados cos sistemas de tratamento actualmente en uso. De acordo con esta nova visión, a presente tese focalizouse no estudo dalgunhas alternativas de mellora das EDAR, mediante a aplicación de novas tecnoloxías de tratamento. Deste xeito, as principais melloras agardadas a partir do desenvolvemento desta tese están relacionadas con: Resumo e obxectivos 12 • A redución da produción de lodo en exceso nos procesos biolóxicos empregados para eliminar materia orgánica e nitróxeno das augas residuais. • A mellora da sedimentabilidade da biomasa nos reactores biolóxicos. • O tratamento das augas residuais para facilitar a súa aplicación como fertilizantes e a recuperación de nutrientes da auga residual. • A redución dos requirimentos enerxéticos para o tratamento das augas residuais, co obxectivo de alcanzar a auto-suficiencia enerxética. Deste xeito, no Capítulo 3, aplicouse a tecnoloxía aerobia granular para o tratamento de xurros porcinos. Esta tecnoloxía permite reducir a produción de lodo en exceso nas EDAR e reducir os requirimentos de terreo necesario para a súa implantación, grazas á formación de biomasa granular con boas propiedades de sedimentabilidade. A formación de gránulos aerobios acádase principalmente en reactores secuenciais SBR, como é o caso do reactor a escala piloto utilizado no Capítulo 3. Con todo, no Capítulo 4, investigouse a posibilidade de obter este tipo de biomasa con boas propiedades de sedimentabilidade nun reactor operado en continuo. No Capítulo 5, recuperáronse estruvita e sulfato de amonio, minerais ricos en nutrientes, a partir de correntes de urina recollida separadamente e do sobrenadante dun dixestor anaerobio. Esta combinación permite recuperar nitróxeno e fósforo das augas residuais. Finalmente, no Capítulo 6, estudouse a aplicación do proceso Anammox (ANaerobic AMMonium Oxidation) para a eliminación de nitróxeno na liña de auga das EDAR, co obxectivo de incrementar a eficiencia enerxética do tratamento de augas residuais. Os contidos principais e obxectivos específicos correspondentes a cada capítulo da presente tese descríbense máis en detalle a continuación. No Capítulo 1 preséntase un resumo actualizado dos retos futuros relacionados co tratamento das augas residuais, e detállanse algunhas das novas tecnoloxías e solucións que xurdiron nos últimos anos relacionadas con este ámbito. Préstase especial atención ás novas tecnoloxías relacionadas coa eliminación e recuperación de nutrientes e á biomasa granular aerobia, temas que se estudan na presente tese. Preséntase unha revisión actualizada dos estudos realizados ata a data nos devanditos temas. No Capítulo 2, proporciónase unha descrición das metodoloxías analíticas empregadas para determinar os parámetros convencionais usados ó longo da tese para caracterizar as augas residuais e a biomasa. Os parámetros máis convencionais como o pH, a concentración de osíxeno disolto (OD), demanda química de osíxeno (DQO) ou a concentración de sólidos, medíronse seguindo as instrucións do “Standard Methods”. Descríbese brevemente a técnica de Hibridación in situ con Fluorescencia (FISH), que se aplica para a identificación das distintas poboacións microbianas implicadas nos procesos biolóxicos de tratamento de augas residuais. Resumo e obxectivos 13 Tras a súa adaptación á investigación desta tese, foron realizadas outras análises, coma como a caracterización da biomasa granular aerobia mediante o índice volumétrico de lamas (IVL), a determinación da densidade de biomasa nos gránulos aerobios, a aplicación de análises dixitais de imaxes para determinar a morfoloxía e diámetro medio dos gránulos, ou a medida da concentración de poli-hidroxi-alcanoatos (PHA) no interior dos gránulos. Tamén se presentan neste capítulo os cálculos empregados para determinar os parámetros que permiten analizar os resultados obtidos ó longo da tese. Os métodos analíticos máis específicos e os cálculos empregados soamente nalgún capítulo, así coma as descricións das montaxes experimentais detállanse no capítulo correspondente. Con respecto á granulación aerobia, a maioría de investigacións centradas nas propiedades físicas e no desempeño da biomasa granular aerobia leváronse a cabo en reactores SBR a escala laboratorio. Estes reactores operan en ciclos secuenciais que normalmente inclúen as seguintes fases: enchido, reacción, sedimentación e baleirado. Os ciclos caracterízanse pola curta duración das fases de enchido e sedimentación, que teñen como obxectivo o desenvolvemento de biomasa en forma de gránulos en condicións aerobias. Con todo, ata o de agora, non existe un número relevante de estudos realizados a escala piloto ou industrial con este tipo de biomasa. Por todo iso, no Capítulo 3, estudouse o arranque e operación dun reactor SBR aerobio a escala de planta piloto, cun volume de operación de 100 L, encamiñado a tratar o efluente dunha granxa porcina. Unha das características asociadas á biomasa aerobia granular é a súa capacidade de soportar variacións da carga aplicada, xa que logo, o obxectivo deste traballo foi investigar o efecto de grandes variacións da composición de entrada dun reactor granular SBR a escala piloto. Esta información é indispensable á hora de proceder ó escalado e aplicación destes sistemas a nivel industrial, xa que este é o obxectivo final do deseño de novas tecnoloxías. Os xurros de porco usados nesta investigación caracterizábanse por unha alta variabilidade na súa composición, en termos de contido de materia orgánica e nitróxeno, e tamén na relación C/N. Esta alta variabilidade foi adicionalmente incrementada mediante a dilución con auga potable. Nunha primeira etapa, estudouse o desempeño da planta piloto durante o arranque do proceso e o desenvolvemento da biomasa granular. Nunha etapa posterior, investigouse a capacidade tanto do reactor coma da biomasa granular para soportar condicións variables de alimentación. Logrouse obter un proceso de granulación no reactor, e as propiedades físicas da biomasa permaneceron estables durante o período de operación, de máis de 300 días. A biomasa granular formada no reactor mostrou excelentes propiedades de sedimentabilidade, con índices volumétricos de lamas que variaron entre 27 e 60 mL/g SST. Estas excelentes propiedades de sedimentabilidade facilitaron a separación da biomasa granular do efluente, e a súa retención no interior do reactor. Os primeiros gránulos aerobios observáronse logo de 9 Resumo e obxectivos 14 días de operación, e o seu tamaño medio durante a operación variou entre 2,0 e 2,8 mm. O reactor mostrou unha boa capacidade de retención da biomasa granular. Así, a concentración de sólidos alcanzada variou entre os 5 e 12 g SSV/L. Con todo, a eficiencia de retención dos sólidos contidos nos xurros alimentados foi máis baixa, e como consecuencia, a concentración de sólidos no efluente do reactor foi similar á medida no influente. A eficiencia de eliminación de materia orgánica non se viu afectada polas flutuacións da velocidade de carga orgánica (VCO) aplicada, senón pola fracción de materia orgánica non biodegradable que contiñan os xurros porcinos. Por esta razón, a eficiencia de eliminación de materia orgánica nunca superou valores superiores ó 80%. A carga amoniacal alimentada foi oxidada fundamentalmente a nitrito, e deste xeito a eficiencia de eliminación de amonio achegouse ó 76%, malia a alta variabilidade da auga residual que constituía a alimentación do reactor. Con todo, practicamente non se observou desnitrificación durante o período experimental. Co obxectivo de avaliar a eficiencia do reactor en maior profundidade realizáronse medidas ó longo de ciclos de operación en diferentes días de experimentación. Observouse que a materia orgánica biodegradable era facilmente eliminada durante os primeiros minutos da fase de aireación en todos os ciclos, no denominado período de saciedade. Durante este tempo, mediuse unha redución na concentración de OD debido á rápida oxidación da materia orgánica. Unha vez que a materia orgánica biodegradable era consumida, a concentración de OD no líquido comezaba a incrementarse (durante o período de fame), namentres que a fracción de DQO non biodegradable permanecía inalterada. Cando a materia orgánica se esgotaba, as concentracións de nitrito e nitrato no medio de reacción aumentaban, pero a concentración de nitróxeno total permanecía practicamente constante. Esta observación suxire que o proceso de nitrificación/desnitrificación simultáneo non ocorría durante a fase de aireación. A concentración de PHA medida na biomasa confirma que as bacterias acumulan materia orgánica en forma dos devanditos compostos durante a fase de saciedade e consómena durante os primeiros minutos da fase de fame. A técnica FISH aplicouse para a caracterización da biomasa obtida co obxectivo de seguir a evolución das poboacións microbianas durante o proceso de granulación e durante a operación do reactor, unha vez que os gránulos estaban formados. As poboacións bacterianas detectadas indicaron unha evolución dende aquela presente no lodo inoculado á que compón os gránulos aerobios. Os organismos de tipo filamentoso estaban presentes fundamentalmente no inóculo, mentres que eses organismos laváronse do sistema unha vez que se desenvolveron os gránulos. As poboacións microbianas nitrificantes estaban compostas principalmente por membros de Nitrosomonas spp. coma bacterias oxidantes de amonio, e dunha menor cantidade de bacterias oxidantes de nitrito, pertencentes ó filo Nitrospirae, o que corresponde coa acumulación de nitrito observada durante a operación do reactor. Resumo e obxectivos 15 A aplicación dos conceptos da tecnoloxía de granulación en reactores SBR para a ampliación ou renovación de EDAR existentes pódese ver limitada polas diferentes condicións de operación e as diferenzas na xeometría entre os reactores en forma de columna SBR e os reactores convencionais de lamas activas. Transformar un sistema continuo nun reactor SBR adecuado para a obtención de gránulos aerobios é en ocasións dificultoso. Por todo iso, o obxectivo do Capítulo 4 foi definir as condicións operacionais axeitadas para o desenvolvemento de biomasa granular nun reactor de tanque axitado operado en continuo (RCTA) con xeometría similar á dos reactores convencionais de lamas activas normalmente empregadas nas EDAR, que teñen relacións de altura/diámetro ó redor de 1. A produción de gránulos aerobios nun reactor en continuo abriría unha nova perspectiva á aplicación desta tecnoloxía para a mellora das EDAR actualmente en operación. Para acadar o desenvolvemento de biomasa aerobia granular colocouse no reactor un sistema de selección de biomasa, baseado na hidrodinámica do reactor e no progresivo acurtamento do tempo de retención hidráulico (TRH). Este sistema consistiu nun tubo semi mergullado no medio de reacción a través do cal se baleiraba o efluente do reactor. Fixouse unha velocidade ascensional de ó redor de 10 m/h neste tubo. Consecuentemente, as partículas que tiñan unha velocidade de sedimentación menor que o valor prefixado eran lavadas do reactor, mentres que os sólidos con boas propiedades de sedimentabilidade eran retidos. Probáronse tempos de retención hidráulicos de 6, 3 e 1 hora en dous reactores de 3 e 6 L de volume útil. Formouse biomasa en forma de gránulos aerobios no reactor cando se aplicou un TRH de 1 hora e se seleccionou unha velocidade de sedimentación de 10 m/h no tubo de descarga de efluente. Doutra banda, acumulouse biomasa de tipo floculenta no reactor cando o TRH foi de 3 e 6 horas. Os gránulos formados tiñan un diámetro medio que chegaba ata os 7 mm, así como altas velocidades de sedimentación, de arredor de 36-48 m/h. Estes valores eran notablemente máis altos que os típicos valores de sedimentabilidade que presenta o lodo activo, e eran comparables a aqueles atopados en gránulos aerobios formados en reactores SBR. Pola contra, os valores de IVL e de densidade eran peores comparados cos valores correspondentes dos gránulos aerobios desenvolvidos en reactores SBR. Durante o proceso de formación dos agregados, o valor de IVL10 reduciuse gradualmente, dende os 700 mL/g SST do inóculo ata os 127 mL/g SST dos gránulos, mentres que o valor de IVL dos gránulos formados en reactores SBR adoitan chegar a ser duns 30-40 mL/g SST. A densidade de biomasa dos agregados formados no reactor operado en continuo variou entre 7 e 11 g SSV/Lgránulo. Estes valores son relativamente baixos, comparados co valor correspondente dos gránulos formados en reactores SBR, que poden chegar a 43,5 g SSV/Lgránulo. De todos os xeitos, a biomasa producida neste sistema operado en continuo é máis facilmente separable da fase líquida, reducíndose deste xeito a necesidade dun gran sedimentador, e facilitando a xestión e eliminación do lodo producido durante o tratamento da auga residual. Resumo e obxectivos 16 En canto á eliminación de materia orgánica e de nitróxeno, tratáronse cargas orgánicas de 4,8-6,0 g DQO/L·d con eficiencias de eliminación do 60%, debido á presenza dunha fracción de materia orgánica non biodegradable. A eliminación de nitróxeno variou entre un 10 e un 15%, e pode ser atribuída totalmente ó crecemento da biomasa, xa que non ocorreron procesos de nitrificación e/ou desnitrificación no reactor durante a experimentación. Esta biomasa de forma granular sería a primeira formada en reactores operando en continuo con xeometrías similares á dos reactores de lamas activas, cuxas relacións altura/diámetro son de arredor de 1. Os obxectivos do Capítulo 5 foron a avaliación do pre-tratamento dunha corrente de urina recollida de forma separada e o seu posterior co-tratamento nun sistema de stripping por aire. No reactor de stripping a urina mesturouse co sobrenadante dun dixestor anaerobio. Os experimentos realizados neste traballo proporcionaron resultados preliminares, polo que será necesario levar a cabo máis experimentos para comprobar a viabilidade do sistema; pero os primeiros resultados foron prometedores. Neste capítulo utilizouse un novedoso sistema de stripping de amonio para tratar o sobrenadante dun dixestor anaerobio. Esta corrente líquida das EDAR caracterízase por ter un caudal relativamente baixo pero por posuír unha alta concentración de nitróxeno. O sistema operouse a escala industrial nunha EDAR municipal, combinándoo cun pre-stripper de CO2 previo e unha unidade de absorción a continuación. Con esta combinación o amonio recuperábase como unha solución de sulfato de amonio que se vendía a agricultores da rexión, os cales usaban este produto como fertilizante. Deste xeito, un produto de desfeito convertíase nun produto con valor. Con respecto ó sistema de operación, a presenza do pre-stripper de CO2 reducía o consumo de produtos químicos do sistema, incrementando o pH do líquido e facilitando a desprotonación do ión amonio. Consecuentemente, o pre-stripper de CO2 fixo aumentar a recuperación de nitróxeno na unidade de absorción á vez que reduciu o consumo de NaOH nun 50%. A eficiencia do sistema incrementouse adicionalmente cando a urina pre-tratada se mesturou co sobrenadante. A urina recollíase separadamente mediante o uso de WC NonMix e de ouriñais sen auga. Despois era pre-tratada nun reactor co obxectivo de eliminar e recuperar o fósforo, que se atopa en altas concentracións na urina. No reactor formábanse cristais de estruvita, un mineral rico en fósforo con aplicacións como fertilizante, unha vez que se engadía óxido de magnesio á urina. Os cristais de estruvita tiñan un tamaño medio de entre 42 e 80 µm. Os primeiros experimentos demostraron a viabilidade do sistema de tratamento combinado do sobrenadante e de urina nun reactor de stripping de amonio que inclúe un prestripper de CO2. Deste xeito, a adición dun 10% en volume de urina tratada ó sobrenadante alimentado ó sistema de tratamento por stripping, produce un incremento do 40% na Resumo e obxectivos 17 concentración de amonio. Malia que as eficiencias obtidas nesta unidade foron menores que as acadadas durante as simulacións, alcanzouse un incremento na produción de sulfato amónico durante os dous experimentos levados a cabo a escala real nunha EDAR municipal. Ademais, os operadores da EDAR non informaron de incidencias relacionadas coa adición de urina tratada ó sistema de stripping. Con todo, a duración dos experimentos foi demasiado curta como para descartar esta posibilidade completamente. No Capítulo 6, levouse a cabo a eliminación autótrofa de nitróxeno a baixas temperaturas (15-20 ºC) usando un sistema baseado no proceso CANON (Completely Autotrophic Nitrogen Removal Over Nitrite) con biomasa granular. Con este sistema avaliouse a posibilidade de aplicar o devandito proceso para a eliminación de nitróxeno na liña de auga das EDAR. Os sistemas CANON baséanse na combinación da nitrificación autótrofa parcial, oxidación da metade do amonio presente ata nitrito por parte das bacterias oxidantes de amonio (BOA); e do proceso Anammox, onde o amonio e o nitrito se combinan para producir nitróxeno gas en condicións limitantes de osíxeno. Os procesos Anammox foron aplicados satisfactoriamente para o tratamento de correntes ricas en amonio a temperaturas superiores a 25 ºC, como é o caso do sobrenadante de dixestores anaerobios de lamas. Con todo, existe pouca información sobre a posible aplicación dos procesos Anammox a correntes con baixas concentracións de nitróxeno e a baixas temperaturas, como é o caso da liña de augas dunha EDAR. Os sistemas CANON presentan como vantaxes menores requirimentos de enerxía para a aireación e que non necesitan dunha fonte de carbono orgánico para a conversión das especies nitroxenadas a nitróxeno gas, comparando cos sistemas convencionais de nitrificación/desnitrificación. Ademais, o feito de non requirir materia orgánica para levar a cabo a desnitrificación heterótrofa fai que queden dispoñibles maiores cantidades de materia orgánica para a súa conversión a metano na dixestión anaerobia, incrementando así a recuperación enerxética das EDAR. O uso de microorganismos autótrofos de baixa velocidade de crecemento, como é o caso das bacterias Anammox, implica a necesidade de usar un sistema caracterizado por unha alta capacidade de retención de biomasa. Neste caso, a retención de biomasa acentuouse mediante a utilización de biomasa en forma de gránulos cultivados nun sistema SBR. As bacterias BOA desenvólvense nas capas máis externas do gránulo, onde o osíxeno e o amonio están presentes. Doutra banda, as bacterias Anammox crecen en capas máis profundas, onde o osíxeno xa está esgotado e o amonio e nitrito están dispoñibles para o seu consumo. O sistema CANON operouse inicialmente a 20 ºC alimentado cunha concentración moderada de amonio (150-250 mg NH4+-N/L). Máis tarde, a temperatura de operación reduciuse a 15 ºC. E finalmente, alimentouse o sistema cun influente cunha baixa concentración de amonio (50-80 mg NH4+-N/L), a cal simula as condicións de operación que Summary and objectives 24 sorber unit. With this combination, the ammonia was recovered as an ammonia sulphate solution at the end of the treatment. This nitrogen rich solution was sold to local farmers who used it as fertilizer. In this way, a waste product was turned into a valuable product. With respect to the system operation, the presence of the CO2 pre-stripper, which reduced the consumption of chemicals in the overall system, increased the pH of the liquid and facilitated the deprotonation of the ammonia. Consequently, the CO2 pre-stripper increased the nitrogen recovery in the sorber unit, while allowed reducing 50% of the NaOH consumption. The efficiency of this system was further increased when pre-treated urine was added to the supernatant liquid. Urine was collected separately by means of No-mix toilets and dry urinals. Then, it was pre-treated in a reactor in order to remove and recover the phosphorus which was present in high concentrations in the urine. Struvite, a phosphorus rich mineral used as fertilizer, precipitated in the reactor after magnesium oxide was added to the collected urine. Struvite crystals with an average size of 42 to 80 µm were formed in the reactor. The first experiments showed the feasibility of the combined treatment system of supernatant liquid and urine in the ammonia stripping reactor including a CO2 pre-stripper. In this way, an increase of 10% in the liquid flux by the addition of the urine represented a 40% increase of the ammonia concentration in the inlet of the stripping unit. Even if the efficiency of the nitrogen removal was lower than the values reached during the optimization and simulations performed during the operation without urine addition, the achievement of these percentages generated a proportional increase in the fertilizer production. The fertilizer production rise partially covered the chemical and operational costs of the ammonia stripping system. In addition, operational problems due to the treated urine addition were not reported by the WWTP operators; however, the length of the experiments was too short to discard this possibility. In Chapter 6, the autotrophic nitrogen removal at low temperatures (15-20 ºC) was performed using a Completely Autotrophic Nitrogen Removal Over Nitrite (CANON) system with granular biomass. This system was operated to research its application to remove nitrogen from the water line of the WWTPs. The CANON system is the combination of the partial autotrophic nitritation, oxidation of half of the present ammonium to nitrite by the Ammonia Oxidizing Bacteria (AOB), and the Anammox process, where ammonia and nitrite are combined to produce nitrogen gas under oxygen limiting conditions. The Anammox processes have successfully been applied to rich ammonia streams at temperatures above 25 ºC, as it is the case of the supernatant from anaerobic sludge digesters. However, little information about the application of Anammox process at low temperature and nitrogen concentration, like it is the case of the main stream of the WWTP, is available. As advantages, the CANON system presents the fact that it requires less aeration energy and no organic carbon source for the conversion to nitrogen gas, compared to the conventional nitrification/denitrification processes. Furthermore, larger Summary and objectives 25 amounts of organic matter are available for methane production, which increase the energy recovery in the WWTPs, as the organic matter is not needed for heterotrophic denitrification. The use of slow growing autotrophic microorganisms, as it is the case of the Anammox bacteria, implies the use of reactor systems characterized by high biomass retention capacities. In this case, the biomass retention was enhanced through the use of biomass in the form of granules grown in an SBR system. AOB bacteria developed in the outer layers of the granules where oxygen and ammonia were present. On the other hand, Anammox bacteria grown in deeper layers, where oxygen was depleted and ammonia and nitrite were available. The CANON system was initially operated at 20 ºC and fed with moderate concentrations of ammonia (150-250 mg NH4+-N/L). Later the temperature of operation was decreased to 15 ºC. Finally, an influent containing low ammonia concentrations (50-80 mg NH4+-N/L) was applied in order to simulate the conditions of the water line of a municipal WWTP where the organic matter was previously removed. At 20 ºC, the biomass retention in the SBR was satisfactorily achieved, and solids concentrations of more than 12 g VSS/L were obtained. Nitrogen removal rate values of 0.45 g N/L·d were achieved, reaching a 70% of nitrogen removal efficiency. However, when the process was operated at 15 ºC the low biomass growth and the difficulties to control the low dissolved oxygen concentrations in the bulk liquid implied the need of testing different reactor configurations. In addition, the low oxygen concentration, temperature and nitrogen load facilitated the development of a third microbial group, the nitrite oxidizing bacteria (NOB). The NOB compete with Anammox for the nitrite and with the AOB for the oxygen. The overcoming of these undesired microorganisms was a challenge in the last part of this research. In order to solve these drawbacks, the inclusion of mechanical stirring, the use of a reactor with a different height to diameter ratio, and variations in the settling time were tested. Two laboratory scale SBRs were used during the experimentation with a working volume of 1.5 L and 4.0 L, respectively. Three main factors should be pointed out as crucial to control the performance of the AOB and Anammox bacteria in a CANON system operated at low temperature and nitrogen load: 1) to achieve a high biomass retention; 2) to achieve an equilibrium between the AOB and Anammox activities and 3) to avoid the NOB development in the biomass. In a WWTP, the conventional activated sludge reactor could be substituted by a continuous aerobic granular reactor, where the organic matter would be removed, while the nitrogen would be removed in the Anammox based process. The different alternatives of treatment research in this thesis are expected to lead to more efficient and sustainable wastewater treatment systems. Chapter 1 INTRODUCTION Summary In this chapter, some of the challenges that the water and wastewater treatment processes have to face up nowadays are briefly described. New processes, new technologies and new paradigm shifts in view of water are under development. The solutions that were feasible and adequate for the 20th century are not valid anymore for the 21st century. Global population growth, climate change, energetic crisis, new pollutants, more and more stringent effluent quality requirements… seriously challenge the efficiency of the classical wastewater treatment plants. The present thesis was focused on some of the new technologies under development in recent years which cover some of these challenges. Aerobic granulation was used in some of the research experiments. It is a new biological process which produces less sludge, allows for high sludge retention times and has a small plant footprint, among other advantages, compared to the conventional activated sludge process. In addition, aerobic granules can combine different biological processes in the same reactor. Wastewater carries a high amount of valuable nutrients; their recovery is one of the challenges of water research. The combination of urine separate treatment, production of struvite, and ammonia stripping can increase the nutrients recovery from wastewater through the production of fertilizers. Anammox based processes have proved their efficiency in the nitrogen removal of the reject water. Nowadays, the scientific challenge is to apply these processes to the main stream of wastewater treatment plants, at low temperatures and nitrogen concentrations, with the multiples advantages that can be achieved. Chapter 1 28 1 OUTLINE 1.1 Water and wastewater in the 21st century 29 1.2 Water demand and sanitation as population grows 33 1.3 Water and energy 36 1.4 Nutrients in wastewater 44 1.4.1 Nutrients recovery 50 1.4.1.1 Urine separation 54 1.4.1.2 Fertilizer production by ammonia stripping from sludge liquid fraction 57 1.4.2 Nutrients removal by Anammox based processes 61 1.4.2.1 Conventional Nitrogen removal processes 61 1.4.2.2 The Anammox processes 62 1.4.2.3 CANON process 64 1.4.2.4 Application of Anammox processes to the main stream of WWTP 67 1.5 Biological compact treatment systems 68 1.5.1 Membrane systems 69 1.5.2 Biofilm systems 70 1.5.3 Aerobic granular systems 73 1.5.3.1 Advantages and drawbacks of aerobic granulation 74 1.5.3.2 Aerobic granules formation 76 1.5.3.3 Aerobic granular reactor configuration. Continuous or discontinuous? 77 1.5.3.4 Scale-up applications 81 1.6 Scope of this thesis 82 1.7 References 85 Introduction 29 1 1.1 WATER AND WASTEWATER IN THE 21ST CENTURY Modern wastewater treatment technology was developed during the 19th and 20th centuries. It evolved from simple aqueduct sewerages that dumped in waterways without treatment to a complete treatment. Modern wastewater treatments technologies include physical, chemical and biological treatments, aimed to remove solids, organic matter and nutrients (Figure 1.1 and Figure 1.2). Development of sanitation was considered the greatest medical advance since 1840, according to a reader survey run by the British Medical Journal (Ferriman, 2007). However, to ensure an adequate supply and quality of water was one of the most serious challenges faced by human groups from the beginnings of civilization (Cassardo and Jones, 2011). Channels for conveying wastewater to fields for fertilization purposes have been used in Syria and Palestine since the earliest Neolithic societies (De Feo et al., 2010). Sewers were already used in some households in the Mesopotamian Empire, 3500 to 2500 BC, as well as latrines leading to cesspits (Lofrano and Brown, 2010). The world's first urban sanitation systems were constructed in Harappa and Mohenjo-Daro in the Indus Valley, in the present Pakistan (Webster, 1962). Initially, wastewater was passed through tapered terracotta pipes into a small sump. Then solids settled and accumulated in the sump, while the liquids overflowed into drainage channels in the street. In Europe, the Minoan (3200-1100 BC) and Etruscan (800-100 BC) civilizations developed basic water and wastewater technologies: water harvesting and distribution systems, cisterns, groundwater and wells, drainage and sewerage systems (Koutsoyiannis et al., 2008, Angelakis et al., 2013). All the Minoan palaces applied strategies to dispose of water and wastewater with open terracotta or stone masonry conduits (Angelakis et al., 2013). These systems were found in the sewage system of Knossos palace (Angelakis et al., 2005). Furthermore, the Romans developed and improved these water and wastewater infrastructures. In this way, the largest known ancient sewer, the Cloaca Maxima in Rome, was built in the 6th century BC (Lofrano and Brown, 2010). Nevertheless, after the Roman Empire collapse, the infrastructures, culture and knowledge of water were forgotten and abandoned. In the western countries the sanitary “dark ages” began (Lofrano and Brown, 2010). In an unprecedented historical regression, water came to be drawn from rivers and wells and to be discharged without treatment. Unless some exception, sewage treatment did not reemerge until second half of 19th century, with the development of some new technologies (Figure 1.1). Primary treatment was developed: the Mouras automatic scavenger (1860), the septic tanks patented by Donald Cameron in 1895, the Imhoff Tank designed by Karl Imhoff in 1906, among others. Secondary treatment consisted of: trickling filter (1893), first patent related to attached growth processes for a moving cylinder with wooden slats (1900) or the activated sludge (Ardern and Lockett, 1914). Advanced treatment came later by means of: biological denitrification in an activated Chapter 1 30 1 sludge process (Ludzack and Ettinger, 1962), biological removal of nitrogen and phosphorous in a single sludge system (Barnard, 1975), membrane systems (1980s) and so on. Nowadays, new concepts are being taken into account in order to face up to new environmental, economic and social limitations, coping with the population growth, the global climate change, water scarcity, etc. (Cassardo and Jones, 2011). Completely new approaches, or the recovery of some old forgotten solutions, are necessary in order to design the modern water and wastewater treatment systems. The new designs should be done with sustainability in mind, as the energy-intensive and chemical-dependent systems in current use are completely unsustainable. The present water cycle in most urban areas includes: • a water collecting and transport system, usually several kilometers away from urban population centers; • a treatment of all the volume to make it drinkable; • a simple water usage, as only a small fraction is used for drinking or cooking; • and then its treatment in huge energy and chemical consuming wastewater treatment plants, which treat a mix of black, grey, yellow, industrial and rain waters. Consequently, the present water cycle is a totally unsustainable cycle. This situation will be further aggravated by global climate change, which is altering water supply and storage patterns in ways that make existing water management infrastructures less effective. Introduction 31 1 Figure 1.1. Evolution of wastewater treatment: AS - Activated sludge; CW - constructed wetlands; RBC - Rotating biological reactors; UASB - Upflow anaerobic sludge blanket; MBRs - Membrane biological reactors, SBR - Sequencing Batch Reactors; MBBR - Moving Bed Biofilm Reactors. Adapted from: Lofrano and Brown (2010). Chapter 1 32 1 A) B) Figure 1.2. A) Simplified configuration of a conventional wastewater treatment plant. B) Image of a wastewater treatment plant (Annabel, 2009). In a step forward (Figure 1.3), wastewater has to be seen as a renewable and recoverable source of energy, resources and water by means of the use of advanced treatment systems, which allow the obtaining of the following improvements: A) B) Figure 1.3. From the present-day conventional wastewater treatment plant (A) to the wastewater treatment plant of the future (B). • to achieve a better water quality, for a full water reuse (Molinos-Senante et al., 2011); • to be more energy efficient, minimizing energy requirements and even producing energy (Siegrist et al., 2008, Garrido et al., 2013); • to produce fewer by-products, as sludge or gaseous emissions, and to optimize their management (Préndez and Lara-González, 2008, Fine and Hadas, 2012, Yan et al., 2013); • to have a small plant footprint in urban areas. Decentralized systems can be seen as an option (Libralato et al., 2012, Suriyachan et al., 2012); • to recover nutrients (Etter et al., 2011, Hülsen et al., 2014) and other compounds, as the biopolymers (Moralejo-Gárate et al., 2011), cellulose as a biomass resource (Honda et al., 2002) or different kind of metals (Petrov and Nenov, 2004, Manipura and Burgess, 2008, Fabbricino et al., 2013); Introduction 33 1 • to minimize impacts on environment (emissions, greenhouse gases) and to treat emerging contaminants, as pharmaceutical and personal-care substances (Fernandez-Fontaina et al., 2012, Rodríguez-García et al., 2012). According to this new approach the present thesis has been focused on the study of some alternatives to improve the WWTPs by the application of new technologies. This is the case of the systems based on aerobic granular biomass, systems which use separated collecting units for nutrients recovery or those based on new processes, as the Anammox processes. All these different alternatives are expected to lead to more efficient and sustainable wastewater treatment systems. These alternatives act over one or more aspects that are considered as susceptible of improvement. In this way, the main improvements expected from this thesis are related to: • the reduction in the sludge production in the biological processes used to remove organic matter and nitrogen from the wastewater; • the improvement of the settleability of the biomass in the biological reactors; • the treatment of wastewater in order to facilitate its application as fertilizer and the recovery of nutrients from the wastewater; • the decrease of energy requirements for the wastewater treatment to achieve the energy self-sufficiency. However, this is a small part of the water cycle where humans can act upon. Furthermore, in order to understand the global situation of the water in the world several aspects have to be considered. 1.2 WATER DEMAND AND SANITATION AS POPULATION GROWS The world population reached 7 thousand million in 2011 and is projected to reach 9 thousand million in 2050 (U.N., 2009). However, its growth has slowed down since the 1970s and it is expected to continue its downward trend. On the contrary, water demand is continually increasing at higher rates than population growth. This water demand growth comes mainly from countries with high rates of economic growth and large current populations. In these countries, people, agriculture and industry demand more water and with higher quality as stated by the United Nations World Water Assessment Programme (2009). Although water covers 70% of the Earth surface, less than 3% of this water is fresh (Figure 1.4). As less than 1% of the freshwater is readily accessible and is unevenly distributed throughout the planet, water scarcity (Rijsberman, 2006) (Figure 1.5) can be the origin of local and international conflicts (Figure 1.6). Some authors talk about “water war” or conflicts due to water disputes (Barnaby, 2009, Rahaman, 2012). By 2025, nearly 2 thousand million people will be living in countries or regions with absolute water scarcity, and two-thirds of the world population could be living under stressed water conditions (World Economic Chapter 1 40 1 The most typical wastewater to energy process is based on the use of sludge produced during both primary settling and biological wastewater treatment, thought anaerobic digestion. In the anaerobic digestion, the readily biodegradable portion of the volatile solids in sludge is converted to biogas (mainly methane, 60-65% and carbon dioxide, 35-40%). Biogas is collected and converted into electricity, and heat can be recovered from the power generation units to heat the digesters, or to generate steam power. Research is focused on the use of codigestion, in order to increase the energy recovery with the addition of some high-strength organic wastes (food-processing operation wastes, fats, oil, grease, animal manure...) (Li et al., 2011, Iacovidou et al., 2012, Marañón et al., 2012, Regueiro et al., 2012). Moreover, different pretreatments, as thermal hydrolysis, mechanical disintegration, electrical pulse treatment, etc. are applied in order to improve the methane production and its generation rate (Val del Río et al., 2011, Carlsson et al., 2012, Appels et al., 2013). The other widely used alternative is the thermal conversion through incineration, gasification, pyrolysis, supercritical water oxidation or steam reformation. In thermal conversion the entire volatile fraction of the biomass is either completely or partially oxidized. Energy can be recovered from the heat liberated during the oxidation, or in some technologies, from gaseous or carbon-based solid residues. Other emerging alternatives convert solids to gases under aerobic conditions to produce synthetic gas or biofuel: • Bio-hydrogen can be produced from wastewater using microbial electrolysis cells (Escapa et al., 2012). MFCs generate electricity from the organics present in wastewater and are a promising innovative approach of renewable energy production from wastewater (Oh et al., 2010). • In the algae bioreactors wastewater and nutrients are used to stimulate algae growth (Pruvost et al., 2011). Then, the residual algae are digested with wastewater solids to produce biogas, which is purified and used as fuel for vehicles (Figure 1.10). To reach the enhanced algal yield, additional CO2 can be obtained by the thermal transformation of external biomass (i.e. sludge from a wastewater treatment plant), together with internal biomass as the digestate from residual algae and wastewater solids. • Cellulosic biofuel can be generated in constructed wetlands where energy output is used for biofuel production (Liu et al., 2012a). Energy can also be recovered from heat from warm water, as hot water is discharged into the sewer system. The average temperature of wastewater leaving the house is 27 ºC (SenterNovem, 2006, Lazarova et al., 2012), while that temperature was 10 ºC when it entered the house (Hofman et al., 2011). The total heat loss through sewage water in the Netherlands is about 110 W/cap (Watt per capita) and accounts for approximately 40% of the total heat loss of a modern house (Hofman et al., 2011). This means that the heat loss via Introduction 41 1 sewage is nearly ten times higher than the energy demand for drinking water supply and wastewater treatment together (Hofman et al., 2011). Energy consumption of drinking water supply is 2.6 W/pe, (Watt population equivalent) and wastewater treatment around 7 W/pe (Hofman et al., 2011). Temperature of grey water is higher, about 38-40 ºC because of warm showers, baths and hot laundry (Roest et al., 2010). Figure 1.10. Micro algae water treatment processes. Several options to recover this heat embedded in water are under research. The heat content of water from households can be recovered within buildings combined with Aquifer Thermal Energy Storage (ATES) systems (Lee, 2013) (Figure 1.11), within houses, i.e. shower heat exchangers (Figure 1.12) (Wong et al., 2010), from the sewer, or at wastewater treatment plants (Frijns et al., 2013). Similar to a geo-exchange system, electric heat pumps transfer thermal energy from warm sewage (12–25 ºC) to a higher temperature useful for residential space heating and domestic hot water (Figure 1.13). Compared to geoexchange, sewer heat recovery is more efficient due to higher heat source temperature and lower installation costs (Roger Bayley Inc., 2009). WWTPs can also recover hydraulic energy by installing large turbines to capture this energy and produce electricity, or installing micro-hydro water turbines or hydrokinetic devices in channels and conduits prior to discharge (NACWA, 2009). Chapter 1 42 1 However, these emerging and promising alternatives are still under development. Consequently, until now, most of the WWTPs are energy consumers. As an example, the cost of lipid production for diesel production from algae, to achieve a 10% return was determined to be around US$ 32/L for open ponds and US$ 68/L for photo bioreactors (Davis et al., 2011). And by the year 2009, there were only three sewer heat recovery systems worldwide that recovered heat from untreated sewage, two in Oslo, Norway and one in Tokyo, Japan (Roger Bayley Inc., 2009). A) B) Figure 1.11. Aquifer Thermal Energy Storage: A) summer operation - cooling; B) winter operation - heating. Reprinted from: Ghaebi et al. (2014). A) B) Figure 1.12. A) Scheme of a shower installation with a single-pass counter-flow heat exchanger. Reprinted from: Wong et al. (2010). B) Diagram of a drain water heat recovery system. Reprinted from: U.S. Department of Energy (2013). Introduction 43 1 Figure 1.13. Heat recovery from the wastewater: the Neighborhood Energy Utility (NEU) provides heat and hot water in the City of Vancouver. Adapted from: Roger Bayley Inc. (2009). The main objective of the wastewater industry has been to meet water quality requirements without major energy considerations. In consequence, WWTPs are hardly ever designed with energy efficiency in mind. Hernández-Sancho et al. (2011) found that only 10% of the WWTPs analyzed in their research were energetically efficient. Until these promising new technologies for energy recovery are developed and widely applied, the main objective of WWTPs has to be the optimization of the energy. A global analysis of the wastewater treatment train in terms of the current water quality requirements and the optimal integration of the energy issues has been performed: to improve energy efficiency, to maximize the use of sludge for energy production and to recover energy from internal or external sources. The National Renewable Energy Laboratory of the U.S. Department of Energy (Daw et al., 2012) and the United States Environmental Protection Agency (E.P.A., 2013) suggested some strategies to improve the energy efficiency in the municipal wastewater treatment facilities: • Operational control. System controls that use supervisory control and data acquisition (SCADA) feedback and variable frequency drives (VFD) can be used to optimize effluent quality and energy consumption. • Installation of variable-frequency drives which adjust the speed of an electric motor by modulating the power being delivered. • Upgrade to energy-efficient motors and motor systems, doing that energy can be saved, maintenance costs reduced and the environment protected. • Heating, cooling and ventilation system upgrades and use of bright lights. • Managing electrical load using energy-efficient strategies such as reducing peak demand, shifting to off-peak hours, and improving the power factors of motors. Chapter 1 44 1 • Repair and replacement. A best practice for facilities is to regularly evaluate the condition, performance and remaining useful life of process equipment. • Biomass. There are significant environmental and financial tradeoffs between providing higher quality biomass and the subsequent energy demands. Sustainable biomass treatment, transport, and end-use can reduce economic and environmental costs. • Infiltration, inflow and leaks. A considerable amount of energy is wasted treating groundwater that infiltrates systems through pipes that are broken or out of alignment. Similarly, breaks and leaks in the collection system increase the energy required to pump sewage to the treatment plant. • On-site renewable energy. A best practice for all communities is to look for opportunities to incorporate on-site renewable energy at their facilities: solar, wind, biodiesel… • Use of co-generation or combined heat and power energy and enhance the production of biogas. Wastewater treatment facilities that have anaerobic digesters create methane gas as a by-product of digestion of biomass. • Conservation. Effective outreach within the community can yield substantial reductions in water use and wastewater generation. Organic matter content of wastewater has a chemical energy by oxidation of 18 W/pe. In contrast, the total energy input of the water cycle is approximately 10 W/pe. Consequently, new technologies in wastewater treatment plants that improve the energy recovery from this organic matter can create an energy self-sufficient water cycle. The best options for energy production from wastewater are production of biogas and using dried sludge for power generation (Hofman et al., 2011). In addition, energy content of nutrients present in water is about 45 MJ/kg N for N-fertilizer and 29 MJ/kg P for P-fertilizer production. The energy recovery potential of these nutrients is 6.1 W/cap for nitrogen and 0.7 W/cap for phosphorus (Hofman et al., 2011). 1.4 NUTRIENTS IN WASTEWATER The production of artificial fertilizers has been increasing continually after the Second World War (Figure 1.14), with a rate of 600% between 1950 and 2000 (IFA, 2006), reaching a rate of 100 million tonnes of nitrogen used per year (Glass, 2003). However, this high rate of fertilizer production for modern agriculture is mainly dependent on phosphorus derived from phosphate rock, which is a non-renewable resource. Introduction 45 1 Figure 1.14. Commercial fertilizer use in the U.S.A. in the period 1960-2005 based on sales data. Reprinted from: U.S. Environmental Protection Agency, (2008). On the other hand, nitrogen and phosphorus are the primary causes of environmental eutrophication in surface waters (E.P.A., 2007). Presence of excess of these nutrients (Figure 1.15 and Figure 1.16) produces different alterations in the natural cycles of rivers, lakes and seas: it promotes excessive plant growth and decay (Figure 1.17), severely reduces the general water quality, provokes undesirable impacts on biodiversity, water quality, fish stocks and reduces the recreational use value of the environment. The annual cost of eutrophication in the U.S.A. is estimated in US $ 2200 million (Dodds et al., 2009). Figure 1.15. Exceedance of critical loads for eutrophication due to the deposition of nitrogen in 2000 and 2010. Reprinted from: European Environment Agency (2012a). Chapter 1 46 1 Figure 1.16. Phosphorus concentration in European lakes and reservoirs Reprinted from: European Environment Agency (2012b). Major sources of nutrients discharges include wastewater, agriculture and atmospheric deposition of nitrogen from burning fossil fuels. Agricultural sources, as fertilizer leaching, runoff from cultivated fields and manure from concentrated livestock operations, and aquaculture are the principal sources of nutrient impairment in waterways in the United States and the European Union (Figure 1.18). On the other hand, urban wastewater is the primary source in Asia and Africa (Selman and Greenhalgh, 2009). In Table 1.3 the main phosphorus sources in each EU member state are indicated. The data are based partly on phosphate input information (detergents) and partly on calculations. The calculation for human phosphate production assumes that each person in the EU generates 0.7 kg P each year and that 50% of this is available. Release of nutrients to the environment is expected to increase in the future (Selman and Greenhalgh, 2009) as a result of the enlarging global trends in population growth, energy use and agricultural production. However, eutrophication can be reversed by controlling the nutrient inputs to waterways (Smith and Schindler, 2009). Introduction 47 1 Table 1.3. Phosphate sources in Europe in 1992 (percentage from each source and total) (Morse et al., 1993) Member State Human Detergents Livestock Fertilizers Industry Background Total (1000 t P/year) Austria 20 10 36 16 6 12 13 Belgium 26 11 43 7 8 5 13 Denmark 12 11 55 11 5 6 15 Finland 18 9 17 15 3 38 9 France 18 15 31 19 6 11 106 Germany 28 3 44 12 6 7 97 Greece 21 7 18 34 5 15 17 Ireland 9 7 49 24 2 9 15 Italy 35 2 26 18 8 11 56 Netherlands 23 3 57 9 5 3 24 Portugal 24 14 27 16 7 12 15 Spain 19 16 18 26 7 14 72 Sweden 21 10 15 14 7 33 14 UK 24 19 29 14 8 6 82 Key assumptions and calculations: Human sources: 1.6 g P/cap·d (human waste) + 0.3 g P/cap·d (household waste) = 0.7 kg P/cap·year * population * 50% (availability). Detergent use: kg P/cap·year * population * 50% (availability). Animal sources: 9.5 kg P/ce·year * ce (cattle equivalent) * 10% (availability). Fertilizer/Manure: 0.6 kg P/ha·year loss from 30 kg P/ha·year application * agric area. Industry inputs: 20 % of domestic sources. Background: 0.2 kg P/ha·year (0.1 in Finland, Norway and Sweden) * whole area. A) B) Figure 1.17. A) Eutrophication at a waste water outlet in the Potomac River, Washington, D.C. Retrieved from: Trubetskoy (2012). B) Algal bloom warning in Loch Melvin, Ireland. Retrieved from: Webb (2012). Chapter 1 48 1 Figure 1.18. Phosphate sources in Europe in 1992. Data from Morse et al. (1993). In order to reverse eutrophication trends and to mitigate nutrient losses, different legislation and normative were approved throughout the world to minimize the amounts of nutrients discharged (Table 1.4), as European Union directives: • The urban wastewater treatment directive aims to protect the environment from the adverse effects of urban wastewater discharges and discharges from certain industrial sectors (EEC, 1991b). • The nitrates directive aims to protect water quality across Europe by preventing nitrates from agricultural sources polluting ground and surface waters and by promoting the use of good farming practices (EEC, 1991a). • The Water Framework Directive (WFD) lays down a strategy to fight against the pollution of water, including adopting specific measures against pollution by individual pollutants or groups of pollutants presenting a significant risk to the aquatic environment (EEC, 2000). Both problems, phosphorus run out and eutrophication can be solved at the same time with technologies for nutrients recovery from wastewater streams (Figure 1.19). The fertilizer value of the nutrients discharged to the sewer systems in Norway per year is US$ 30 million. In addition, 15-20% of the current mineral fertilizer used could be substituted by fertilizer derived from wastewater. In the case of developing countries these percentages can achieve values up to 40-50% (Vråle and Jenssen, 2005). The fertilizer value of human excreta was already recognized from the beginning of civilization. In this way, the Ancient Greeks used public latrines and later conducted the wastewater outside the city to agricultural fields, where wastewater was used for irrigation and to fertilize crops and orchards (Henze et al., 2008). At the beginning of 20th century, a plan for Background; 10.0% Human; 23.0% Detergents; 11.0% Livestock; 32.0% Fertilizers; 17.0% Industry; 7.0% Introduction 49 1 separate collection of toilet water through a vacuum sewer and grey and storm water was developed by Mr. Liermur. The collected sewage was not treated, but directly spread out over land as fertilizer. However, water logging became a problem, and the continuous expansion of the cities made it more difficult to find sufficient land nearby (Henze et al., 2008). Table 1.4. Requirements for discharges from urban wastewater treatment plants to sensitive areas which are subject to eutrophication. One or both parameters may be applied depending on the local situation. The values for concentration or for the percentage of reduction shall apply (EEC, 1991b). Parameters Concentration Minimum percentage of reduction (1) Total phosphorus 2 mg/L P (10000-100000 pe) 1 mg/L P (more than 100000 pe) 80 Total nitrogen (2) 15 mg/L N (10000-100000 pe) 10 mg/L N (more than 100000 pe) (3) 70-80 (1) Reduction in relation to the load of the influent. (2) Total nitrogen means: the sum of total Kjeldahl-Nitrogen (organic N+NH3), nitrate (NO3–)-nitrogen and nitrite (NO2–)-nitrogen. (3) Alternatively, the daily average must not exceed 20 mg N/L N. This requirement refers to a water temperature of 12 °C or more during the operation of the biological reactor of the wastewater treatment plant. As a substitute for the condition concerning the temperature, it is possible to apply a limited time of operation, which takes into account the regional climatic conditions. Figure 1.19. Sustainable scenario for meeting long-term future phosphorus demand through phosphorus use efficiency and recovery. Reprinted from: Cordell et al. (2011). Chapter 1 56 1 Urine from healthy persons is quite stable and hardly contains any microorganisms (Udert et al., 2006). The average composition of fresh and stored urine is summarized in Table 1.5. Table 1.5. Reference values for fresh urine and stored urine per person and day (Udert et al., 2006). Fresh urine Stored urine Average coefficient of variation % Data range Total nitrogen g N/m3 9200 20 - 9200 Total ammonia g N/m3 480 29 - 8100 Ammonia NH3 g N/m3 0.3 - - 2700 Urea g N/m3 7700 20 - 0 Total Phosphate g P/m3 740 14 - 540 Calcium g/m3 190 22 - 0 Magnesium g/m3 100 21 - 0 Potassium g/m3 2200 - 1300-3100 2200 Total carbonate g C/m3 0 - - 3200 Sulphate g SO4/m3 1500 29 - 1500 Chloride g/m3 3800 - 2300-7700 3800 Sodium g/m3 2600 - 1800-5800 2600 pH 6.2 8 - 9.1 Alkalinity mM 22 - - 490 COD g O2/m3 10000 4000 - 10000 Volume L 1.25 0.61 - 1.25 In fresh urine, about 85% of nitrogen is fixed as urea and about 5% as total ammonia. Ubiquitous urea-hydrolyzing bacteria catalyze the hydrolysis of urea to ammonia and bicarbonate (Udert et al., 2006), causing a strong pH increase, as it is shown in equation (1.1). Only few days are necessary for complete urea depletion in the collection tank (Udert et al., 2003). 2 2 2 34 3 NH (CO)NH 2 H O NH NH HCO     (1.1) After urine is collected separately, it is available for its processing. The main options are: • Urine used directly as a liquid fertilizer (Kirchmann and Pettersson, 1995). However, direct reuse of urine is not widely applied in industrialized countries. This is mainly due to issues related to possible effect of micropollutants contained in urine, a perceived hygiene risk, and an expected soil and plant sensitivity to the high salt concentrations (Maurer et al., 2002). • The existing sewer network could be used for transporting urine to a treatment facility at night when the sewers are empty (Larsen and Gujer, 1996), or it can be transported Introduction 57 1 by trucks. Another option is to improve the existing collection system for source separated resource streams (Figure 1.28 A and B). • A wide variety of processes can be performed to treat the collected urine. Maurer et al. (2006) defined seven main purposes of a treatment unit: volume reduction, Precovery, N-recovery, stabilization, hygienization, removal of micropollutants and biological nutrient removal. A) B) Figure 1.28. A) Source separated wastewater streams with the modification of existing collection system (Tchobanoglous, 2012). B) Schematic drawing of an integrated urbanwater and resource-management system (Daigger, 2008). 1.4.1.2 Fertilizer production by ammonia stripping from sludge liquid fraction Ammonia stripping by means of air supply is a unit process which brings air and water into intimate contact to transfer volatile substances from the water into the air. In a waste stream, ammonia nitrogen exists in aqueous solution as either ammonium ion (NH4+) or ammonia (NH3) (Idelovitch and Michail, 1981). The degree to which ammonia forms the ammonium ion depends on the pH of the solution, following equation (1.2). This dependence is shown in Figure 1.29 at 0, 20, and 40 °C (Huang and Shang, 2004). This equilibrium is the basis of the ammonia-stripping process, a simple desorption process used to lower the ammonia content of a wastewater stream. In this process, the pH of the water is raised in such a way that the equilibrium of equation (1.2) displaces to NH3 (aq) formation. Then, this gas is removed from the solution by water/gas exchange.     43 4 32 NH NH aq H NH OH NH aq H O      (1.2) 1. Below pH 7 and at 20 ºC, virtually all the ammonia will be soluble in the form of ammonium (Idelovitch and Michail, 1981). Chapter 1 58 1 2. Above pH 11.5, virtually all the ammonia will be present as a dissolved gas. At pH 10.5, most of the ammonia (about 95%) is found in the gaseous form (Idelovitch and Michail, 1981). 3. In the range between 7 and 11, both ammonium ions and dissolved gas coexist. 4. Percentage of dissolved gas increases with temperature and pH and as a consequence, temperature and pH favor the removal of ammonia from solution. Figure 1.29. Effects of pH and temperature on the distribution of ammonia and ammonium ion in water. Reprinted from: Huang and Shang (2004). After the ammonia desorption from the wastewater, this compound can be recovered as ammonium sulfate or other salt, using a closed-loop reactor. In the closed-loop system, ammonia is absorbed in sulphuric acid, to produce ammonium sulphate, which may have potential for being used as an agricultural fertilizer. The clean gas coming from this unit is then recycled back to the stripper. One of the most ammonium concentrated streams in a WWTP is the sludge liquid fraction. It results out of the anaerobic sludge digestion and subsequent sludge treatment, like dewatering and drying. Sludge liquid contains between 600 and 1200 mg NH4+-N/L, as well as presents a higher temperature, typically 30 ºC, compared to influent wastewater (Wilsenach, 2005). In anaerobic conditions, nitrogen remains in the form of ammonium and its concentrations are not reduced during this process (Rousseau et al., 2008). Consequently, the anaerobic effluent needs to be additionally treated to remove the nitrogen load. Ammonia stripping of the anaerobically digested effluent in a biogas plant is feasible due to the available heat from biogas utilization, low investment and basic pH of an anaerobically digested effluent (Bonmatí and Flotats, 2003). These authors showed that with pH above 11.5, temperature of Introduction 59 1 80 ºC and 3 hours of retention time in the batch stripping column, up to 99% of ammonium nitrogen can be stripped from the wastewater. Thus, an important pH increase is necessary to obtain a good ammonia removal from wastewater. That increase can be produced by the addition of chemicals as NaOH or Ca(OH)2 for example. However, this can significantly increase the cost of the process. An alternative to reduce the costs is the use of a CO2-pre-stripper column, previously to the stripping reactor. The pH of the anaerobic digestion effluent can be increased from about 7 to about 9 using this alternative (Lei et al., 2007). After heating the ammonia rich liquid, ammonium and bicarbonate are partly transformed to the gaseous components free ammonia and carbon dioxide. The stripping rate for CO2 is higher, by two orders of magnitude, than that for ammonia. This happens since the dimensionless Henry’s law constant is 0.011 and 0.95 for ammonia and CO2, respectively (Fattah et al., 2008). Carbon dioxide could be stripped in a first stripper column without losing too much of free ammonia in the off-gas and with significantly lower air flow than in the ammonia stripper. With the inclusion of the additional CO2 stripping column (Figure 1.30), most of the nitrogen is in the form of NH3 prior to enter the ammonia stripper because of the pH increase, reducing the amount of base solution that is needed for the removal of ammonia in the overall system (Figure 1.31). As a by-product of the stripping process, the wastewater treatment plant will produce ammonium sulphate, a marketable fertilizer. This product can be applied to soils by means of the Controlled Uptake Long Term Ammonium Nutrition fertilization (CULTAN), a type of injection of fertilizer (Figure 1.32). Using this technique, the entire amount of nitrogen needed for a plant to grow is injected at one time, with a more precise application and a more uniform distribution of the fertilizer. Injecting nitrogen into the soil leads to a higher dry matter content of the plant. In addition, it allows the spread of nitrogen regardless the conditions of the field, reduces soil compaction caused by tractors moving across the field and produces lower nitrate and N2O losses than conventional fertilization (Kozlovský et al., 2009). Chapter 1 60 1 Figure 1.30. CO2 stripping pre-treatment scheme. Figure 1.31. Functional principle of free ammonia stripping and sorption with pretreatment of the supernatant by CO2 stripping. A) B) Figure 1.32. A) Fluid fertilizers injection. Retrieved from: (Werktuigendagen, 2009) B) Block diagram of a spur wheel. Retrieved from: (Sommer et al., 2010). Pre-treatment with CO 2 stripping – reduction in base demand for pH shift Sludge liquid preheated to 65°C pH = 7-8 Air Off gas to biofilter rich in CO 2 Shift in pH with partial deprotonation of NH 4 : NH 4 + + HCO 3 - => NH 3 + CO 2 Sludge liquid pH = 8-8.5 NH 3,L CO 2,L Q L Q air CO2 Stripper NH 4,F W NH 3,L NH 3,air NH 3 -free sludge liquid Sulfuric acid (H 2 SO 4 ) (NH 4 ) 2 SO 4 (pH ∼5) NH 3,air Q air Off gas to remove CO 2 QL Q air Stripper Sorber Sludge liquid Fresh air Sludge liquid Gas circulation system Sodium hydroxide (NaOH) NH 3,L CO 2,L Q L Q air CO2 Stripper NH 4,F W NH 3,L NH 3,air NH 3 -free sludge liquid Sulfuric acid (H 2 SO 4 ) (NH 4 ) 2 SO 4 (pH ∼5) NH 3,air Q air Off gas to remove CO 2 QL Q air Stripper Sorber Sludge liquid Fresh air Sludge liquid Gas circulation system Sodium hydroxide (NaOH) Introduction 61 1 1.4.2 Nutrients removal by Anammox based processes 1.4.2.1 Conventional Nitrogen removal processes Nitrogen is usually present in the wastewaters as ammonium or organic nitrogen. The conventional technologies for nitrogen removal in current wastewater treatment plants are based on the Nitrification/Denitrification processes. Initially, ammonia is oxidized to nitrite in the nitritation process, equation (1.3) by Ammonia Oxidizing Bacteria (AOB). Then, the nitrite is oxidized to nitrate by Nitrite Oxidizing Bacteria (NOB), following equation (1.4), in the nitratation process. These processes imply an oxygen consumption of 4.57 kg O2/kg Noxidized. In the denitrification process, the nitrate and/or nitrite presents in the wastewaters is reduced to molecular nitrogen in anoxic conditions by the action of heterotrophic bacteria, following equation (1.5). Organic carbon is needed as the electron donor, while nitrate acts as the last electron acceptor in the respiratory chain substituting the O2 molecule. The most common and widely distributed denitrifying bacteria are Pseudomonas species, which can use hydrogen, methanol, carbohydrates, organic acids, alcohols, benzoates, and other aromatic compounds for denitrification (Metcalf & Eddy et al., 2002). For wastewaters with lower chemical oxygen demand to nitrogen ratios (COD/N), external addition of a carbon source such as methanol is needed to avoid the limitation of the denitrification process. 4 2 3 22 2 NH 1.5 O 2 HCO NO 3 H O 2 CO       (1.3) 2 23 NO 0.5 O NO   (1.4) 33 3 22 2 NO 0.625 CH COOH HCO 0.25 CO 0.5 N 0.75 HO      (1.5) The AOB and NOB are two phylogenetically unrelated groups with different growth rates. The way their growth rates are affected by parameters like temperature, pH, free ammonia, dissolved oxygen (DO) concentration, etc. is different. These differences can be used to outcompete NOB and to uncouple both reaction rates (Sinha and Annachhatre, 2007). NOB inhibition conduces to an accumulation of nitrite, which can be exploited by the partial nitrification/denitrification process (Figure 1.34). With the application of partial nitrification/denitrification process, a saving of 25% of oxygen and 40% of organic carbon source can be obtained. An additional advantage of this technology is that the sludge production is 40% of that corresponding to nitrification/denitrification processes (van Kempen et al., 2001). The partial nitrification/denitrification strategy has been implemented in different reactor configurations (Sinha and Annachhatre, 2007), as the SHARON reactor (Single reactor for High-activity Ammonia Removal Over Nitrite) developed by Hellinga et al. (1998). Chapter 1 62 1 Figure 1.33. Comparison of nitrification/denitrification and partial nitrification/denitrification processes. 1.4.2.2 The Anammox processes The ANaerobic AMMonium OXidation (Anammox) process, involving a new route in the nitrogen cycle (Figure 1.34), was predicted thermodynamically by Broda, (1977) and experimentally observed in the 1990’s decade (Mulder et al., 1995). Soon after its discovery, the potentiality of this unknown process was applied to the environmental field (Strous et al., 1997). Furthermore, Anammox bacteria have been found to be essential components of the global nitrogen cycle and account for 50% of the world nitrogen turnover (Pennisi, 2012). Anammox are chemolitoautotrophic bacteria capable of converting ammonium together with nitrite directly to dinitrogen gas, in the absence of any organic carbon source, following equation (1.6) (Strous et al., 1998). Nitrite acts as electron acceptor in the reaction. The highly toxic “rocket fuel” hydrazine (N2H4) and nitric oxide (NO) are the two intermediates of this process (Kartal et al., 2011). A small amount of nitrate is also produced in the anabolism of Anammox bacteria. + 42 3 2 3 2 0.5 0.15 2 NH 1.32 NO 0.066 HCO 0.13 H 1.02 N 0.26 NO 0.066 CH O N 2 H O         (1.6) Ammonium and nitrite are consumed on an almost equimolar basis, thus, half of ammonia in the wastewater treatment plant must be oxidized to nitrite by means of a partial nitritation process. Introduction 63 1 Figure 1.34. The updated nitrogen cycle with autotrophic nitrogen removal (Anammox) Adapted from: Campos et al. (2010). Anammox bacteria share features with all three domains of life, Bacteria, Archaea and Eukarya, making them extremely interesting from an evolutionary perspective. Anammox bacteria belong to the genus Plantomycetes (Strous et al., 1999a), a phylum of the domain Bacteri. Different species of Anammox bacteria have been identified in the last years: Candidatus Brocadia anammoxidans, Candidatus Kuenenia stuttgartiensis, Candidatus Scalindua brodae, Candidatus Anammoxoglobus propionicus, Candidatus Brocadia fulgida… (Schmid et al., 2000, Schmid et al., 2003, Kartal et al., 2007, Kartal et al., 2008). They are structurally distinct from other bacteria in that they have intracytoplasmic membranes that compartmentalize the cell in organelles. These membranes are called anammoxosomes (Jetten et al., 2001). These bacteria also contain unique membrane lipids named ladderanes (Damste et al., 2005, van Niftrik and Jetten, 2012). The optimal temperature and pH of operation of Anammox bacteria have been found to be 35 ºC and 8, respectively. These bacteria are characterized by a low productivity; 0.038 g VSS/g Nremoved, and a slow growth rate with large doubling times as long as 11 d (Strous et al., 2002). Although Anammox bacteria were recently identified, these microorganisms are truly ubiquitous. They appear to be present in virtually any N-containing ecosystems with a pronounced suboxic zone or chemocline (a layer caused by a strong, vertical chemistry gradient within a body of water), marine sediments, wastewater treatment plants, etc. (Thamdrup and Dalsgaard, 2002, Dalsgaard et al., 2003, Kuypers et al., 2003, Francis et al., 2007). The application of Anammox processes to the removal of nitrogen in the wastewater treatment plants can produce an important reduction of energy consumption and resources needs (Figure 1.35). Chapter 1 64 1 Figure 1.35. Comparison between the combination of nitrification/denitrification processes and the partial nitrification/Anammox processes for the treatment of wastewaters with low COD/N ratio. The cost of the nitrogen removal using Anammox process has been estimated as 1 €/kg Nremoved, while other conventional nitrogen removing techniques cost around 2–4 €/kg Nremoved (van Dongen et al., 2001). The combination of the partial nitritation and Anammox processes requires less than half of the aeration energy, and no carbon source presence, compared to conventional nitrification and heterotrophic denitrification processes (Siegrist et al., 2008). Furthermore, less sludge is produced as a by-product in comparison to the application of the nitrification/denitrification or the partial nitrification/denitrification strategies. 1.4.2.3 CANON process Two different configurations are used to combine ammonia oxidation and Anammox processes: a two reactors configuration (Hellinga et al., 1998) with each process carried out in different units, and a single stage system under oxygen limiting conditions. Different acronyms have been used to define the single stage alternative: OLAND (Oxygen-Limited Aerobic Nitrification and Denitrification) (Windey et al., 2005) and aerobic Deammonification DEMON (Wett, 2006). These names were based on the idea that the ammonia oxidizing bacteria carried out the denitrification process. However, nowadays, it is known that Anammox bacteria are the responsible bacteria for the denitrification process, becoming the CANON acronym (Completely Autotrophic Nitrogen removal Over Nitrite) (Sliekers et al., 2002) the most suitable one to define the process. Introduction 65 1 The low biomass yield of Anammox and ammonia oxidizing bacteria (Strous et al., 1999b) is at the same time an advantage, due to its low sludge production, and a disadvantage, due to the need of good biomass retention capacity of the reactor. In consequence, systems with a high sludge retention time are needed, such as systems based in a biofilm-based setup: rotating contactors (Pynaert et al., 2003), moving bed reactors (Cema et al., 2006), fixed bed reactors (Furukawa et al., 2006) and granular biomass reactors (Fernandez et al., 2008); or based in membrane bioreactors (Trigo et al., 2006, van der Star et al., 2008). Reactors based on granular biomass are suitable to develop the CANON process according to the “magic bead concept” (Santos et al., 1996). The system is based on the establishment of aerobic and anoxic zones within the granule and on the physical separation of the nitrifying and Anammox populations. In this way, AOB can grow in the outer part of the aggregates, where they produce nitrite and consume oxygen to provide anoxic conditions in the inner part of the granule. In this anoxic zone, ammonium, left from the incomplete AOB activity, and nitrite, produced during partial nitrification, have to be present in order to allow the growth of Anammox bacteria (Vázquez-Padín et al., 2009a). A schematic overview of these reactions in a biofilm system is shown in Figure 1.36. This stratification of microorganisms was demonstrated through fluorescence in situ hybridization (FISH) and microsensors measurements (Nielsen et al., 2005, Vázquez-Padín et al., 2010b). A) B) Figure 1.36. Schematic overview of the aerobic and anoxic reactions in a CANON biofilm system A) without and B) with the inclusion of NOB. In order to maintain the stability of the process some operational considerations must be taken into account: a) The nitrifying activity should be high enough to deplete oxygen in the outer part of the biofilm and to avoid its presence in the inner layers. Thus, to obtain anoxic conditions in the inner part of the biofilm, the thickness of the biofilm must be larger than the oxygen penetration depth. Chapter 1 72 1 Figure 1.40. Scheme of a biocarrier in a MBBR (HeadworksBIO). Biological aerated filters (BAFs) are submerged three-phase fixed-media reactors for wastewater treatment, which combine filtration with biological carbon reduction, nitrification or denitrification. A major characteristic of BAF reactors is the use of filter media. The media is either in suspension or supported by a gravel layer at the base of the filter. The dual purpose of this media is to support highly active biomass that is attached to it and to filter suspended solids in one unit. BAF is operated either in upflow or downflow configuration, depending on design specified by manufacturer. Trapped solids and growing biomass gradually block the filter pathways. These obstructions are cleared from time to time by air-scouring or backwashing with treated effluent. The BAFs have been applied to nitrogen removal using external carbon sources, as well to simultaneous biological phosphorus and nitrogen removal using a wastewater carbon source (Sammut et al., 1994). Figure 1.41. Rotating biological contactor. Retrieved from: Mbeychok (2007) Introduction 73 1 A) B) Figure 1.42. A) Scheme of a typical trickling filter (Mbeychok, 2011). B) Sewage treatment trickling filter bed using plastic media in a small rural treatment plant at Beddgelert sewage treatment, Gwynedd, Wales, U.K. (Velela, 2005). 1.5.3 Aerobic granular systems Granular growth is a particular case of biofilm growth, in which no carrier material is required. Bacteria are self-immobilized forming microbial aggregates. Initial applications of granular biomass were associated to anaerobic process: Upflow Anaerobic Sludge Blanket (UASB), Expanded Granular Sludge Blanket (EGSB) or Internal Circulating (IC) Reactors. The first studies about granular biomass developed in aerobic conditions were published in the early 1990’s. Mishima and Nakamura (1991) developed aerobic granules in an Aerobic Upflow Sludge Blanket (AUSB) reactor. However, until the late 1990’s the aerobic granular technology did not emerge (Campos et al., 2009) with the use of Sequencing Batch Reactors (SBRs). The definition of “aerobic granule” emerged from the discussions which took place at the 1st IWA-Workshop Aerobic Granular Sludge, in Munich (2004) and literally stated that: “Granules making up aerobic granular activated sludge are to be understood as aggregates of microbial origin, which do not coagulate under reduced hydrodynamic shear and which settle significantly faster than activated sludge flocs” (de Kreuk et al., 2005a). According to this definition, aerobic granules are biomass structures, which fulfill the following requirements: • The values of the sludge volume index after 10 and 30 minutes of settling (SVI10 and SVI30, respectively) do not differ more than a 10% (Schwarzenbeck et al., 2004) meaning that no thickening of the biomass occurs after settling. • Granules do not coagulate and settle as separate units. Chapter 1 74 1 • The position of microorganisms is fixed and it does not change quickly as in an activated sludge floc due to the existence of a matrix of biomass and EPS. • No carrier material is intentionally involved or added. • The minimum size of the granules is considered to be around 0.2 mm (de Kreuk et al., 2005a), in order to be able to separate them from a sludge sample by sieving. • Bacterial aggregates with an SVI10 value equal or lower than 60-70 mL/g TSS are conventionally considered as aerobic granules. 1.5.3.1 Advantages and drawbacks of aerobic granulation Aerobic granular technology has several advantages compared to activated sludge processes, such as good biomass retention capacity, ability to withstand shock and toxic loadings (Tay et al., 2005b), and presence of aerobic and anoxic zones inside the granules (Tay et al., 2002) to perform simultaneously different biological processes. The reduction in the sludge production compared with activated sludge is another great advantage. Initially, the aerobic granules were developed in SBR systems to oxidize the organic matter from the wastewaters (Beun et al., 1999). Later, the operational conditions allowed the development of aerobic granules, which are capable to remove at the same time nutrients: nitrogen and phosphorus (de Kreuk et al., 2005b). When organic matter oxidation and nitrogen removal occur inside a granule, the microorganisms which require aerobic conditions are situated on the granule surface, being in the outermost layer the aerobic heterotrophs which oxidize the organic matter. While the nitrifying organisms, which oxidize the ammonia to nitrite/nitrate, are located a little in depth (Figure 1.45). Inside the granules, where the dissolved oxygen is absent, the denitrifiers are located to perform the reduction of nitrite/nitrate to nitrogen gas. If the granule is big enough, an anaerobic zone in the core of the granule can exist. In those cases where phosphorus is simultaneously removed, the distribution of the different processes is as indicated in Figure 1.46. This stratification also happens during the formation of CANON granules, where ammonia oxidizing bacteria occupy an external nitrification zone while Anammox locate in a more internal zone (Vázquez-Padín et al., 2010b) where oxygen cannot penetrate. The biomass grows as compact and dense microbial granules, enabling better biomass retention in the reactor, which is important for substrate conversion capabilities and for implantation. The high biomass concentration that can be achieved with aerobic granular reactors allows to treat high organic loading rates (OLRs), up to 15 g COD/L·d (Moy et al., 2002). A comparison between the activated sludge systems and the aerobic granular technology indicates that the latter presents several improvements regarding costs and quality of the produced effluent. Some of these advantages are indicated in Table 1.6. Introduction 75 1 Figure 1.43. Distribution of the organic matter and nitrogen removal processes inside a granule. Adapted from: Mosquera-Corral et al. (2012). Figure 1.44. Distribution of the organic matter, nitrogen and phosphorous removal processes inside a granule. Adapted from: Mosquera-Corral et al. (2012). Table 1.6. Comparison between activated sludge systems and aerobic granular systems (de Bruin et al., 2004). Parameter Activated sludge to remove nitrogen Aerobic granular system Effluent quality Good Similar or better Process stability Good Similar or better Implantation surface 100% 25% Energy consumption 100% < 65-75% Sludge production 100% Similar or better Building costs 100% Significantly lower Operational costs 100% Significantly lower AEROBIC CO 2 Organic source O 2 AEROBIC RESPIRATION NITRIFICATION O 2 NH 4+ NO 2NO 3N 2 DENITRIFICATION ANOXIC ANAEROBIC PHB AEROBIC CO 2 Organic source O 2 AEROBIC RESPIRATION NITRIFICATION O 2 NH 4+ NO 2NO 3N 2 DENITRIFICATION ANOXIC ANAEROBIC PHB AEROBIC PHB Acetate NITRIFICATION O 2 NH 4+ NO 2NO 3N 2 Poly-P P-REMOVAL (ANAEROBIC STAGE) PO 4-3 P-REMOVAL ANOXIC AEROBIC PO 4-3 PHB O 2 AEROBIC PHB Acetate NITRIFICATION O 2 NH 4+ NO 2NO 3N 2 Poly-P P-REMOVAL (ANAEROBIC STAGE) PO 4-3 P-REMOVAL ANOXIC AEROBIC PO 4-3 PHB O 2 Chapter 1 76 1 However, there are still some drawbacks, which need further research like the fact that the produced effluents contain solids concentrations over 75 mg TSS/L. Consequently, these effluents need additional treatment to accomplish the discharge requirements. This solids content can be reduced by means of filtration systems, as membrane systems, settlers, sand filters, etc. Furthermore, the aeration cost is relatively high due to the need of large air flows to keep the required DO concentration and the appropriated mixture. In contrast, when slowgrowing microorganisms, such as the nitrifiers or phosphorous removal bacteria, are used, the DO requirements decreased (de Kreuk and van Loosdrecht, 2004). Additionally, during the start-up, large quantities of the added inoculum are washed out from the reactor, decreasing temporarily the quality of the produced effluent. This can be avoided by inoculating the reactor with previously developed granular biomass (Liu et al., 2005). 1.5.3.2 Aerobic granules formation Aerobic granulation is a gradual process from seed sludge to compact aggregates, further to granular sludge and finally to mature granules (Tay et al., 2001a). Verawaty et al. (2013) proposed a conceptual model describing how granules grow up to a certain critical size. Then, granules that have managed to grow larger than the critical size tend to break/attrite and in this way, they reduce in size down to the critical one, achieving a steadystate distribution of granule sizes (Figure 1.45). Critical size and size distribution depend on the operating conditions of the reactor, wastewater characteristics, aeration, reactor geometry, mixing, and solids concentration. Figure 1.45. Conceptual model for granule formation and breakage/attrition. Reprinted from: Verawaty et al. (2013). The key operational factors that promote aerobic granulation in SBRs have already been established: Introduction 77 1 • Feast-famine regime. The heterotrophic biomass must be cyclically subjected to periods of availability (feast) and lack (famine) of organic substrate in the liquid phase. During the feast period, the organic matter is oxidized and stored inside the bacteria cells as glycogen, lipids or Poly-β-Hidroxyalkanoates (PHA), like the Polyhydroxybutyrate (PHB). On the contrary, during the famine period, the bacteria grow on the stored compounds (Beun et al., 2002). • Short settling time. Biomass with good settling properties is retained inside the reactor, while the flocculent biomass is washed out. In aerobic granulation research, a short settling time has been commonly used to enhance aerobic granulation in SBRs (McSwain et al., 2004, Qin et al., 2004). Settling time is considered as a key factor for granulation. In fact, when long settling times are applied, poorly settling sludge flocs cannot be effectively withdrawn; and they may overtake granule-forming microorganisms (Campos et al., 2009). • Hydrodynamic shear force. Formation of aerobic granules and their physical granule integrity is stimulated by high shear forces (Chen et al., 2007). The shear force in aerobic granular systems is achieved by means of mechanical stirring and/or aeration. The collisions between the granules provoke the detachment of weakly attached materials on the surface of the aggregates, helping to maintain their high densities and smooth surfaces. In fact, some authors reported that aerobic granules could be formed only above a threshold shear force value, in terms of superficial upflow air velocity, of 1.2 cm/s in a column SBR (Tay et al., 2001b). More regular, rounder and more compact aerobic granules were developed at high hydrodynamic shear forces (Liu and Tay, 2002). • Other factors. Several parameters have been reported as important for aerobic granulation, such as: substrate composition, fed organic load, exopolymeric substances formation, presence of divalent cations, and DO concentration. However, once these important parameters are controlled, aerobic granule formation shows an important flexibility regarding operation conditions. This technology has been applied to treat different kinds of wastewaters (Arrojo et al., 2004, Inizan et al., 2005, Figueroa et al., 2008) and under different operational conditions of oxygen limitation or low temperature (de Kreuk et al., 2005c). 1.5.3.3 Aerobic granular reactor configuration. Continuous or discontinuous? Nowadays, the development of aerobic granules is mostly focused on the application of column-type Sequencing Batch Reactors (SBRs) with a large height to diameter ratio (H/D), since these systems fulfill most of the needed requirements for this aim: short settling periods, alternative feast-famine periods and high shear forces. These SBRs operate in sequential cycles distributed in different operational phases: filling, reaction, settling, effluent withdrawal and idle time (Figure 1.46). Chapter 1 78 1 Figure 1.46. Operational phases of a cycle from a SBR . Retrieved from: MosqueraCorral et al. (2012). Nevertheless, the application of the concepts of the granular SBR technology for the upgrading of existing WWTPs could be limited by the different operational conditions and geometry of both column-type SBR and conventional activated sludge reactors. Transforming a continuous system into a SBR suitable to obtain aerobic granules is difficult. Curiously, first applications of granular biomass grown in aerobic systems were developed on continuous systems, on the Aerobic Upflow Sludge Blanket (AUSB) reactor in a research performed by Mishima and Nakamura (1991) (Figure 1.47). This system consisted of two different units: 1) an AUSB reactor where high upflow velocities were maintained in order to achieve a high hydraulic selection pressure to retain only granular biomass and to wash out suspended biomass, 2) an external oxygenation chamber where pure oxygen was added to the recycled effluent. This kind of operational strategy promoted sludge aggregation and also allowed most of the produced biomass to be retained in the form of a well-settled sludge blanket (Sharma and Huang, 2004). However, this system required high recirculation rates in order to obtain suitable oxygen transfer rates to remove pollutants and used pure oxygen, which increased the operational costs. Thus, this work was not really appreciated at that moment (Mosquera-Corral et al., 2012). In consequence, there are very few reported works on AUSB and this kind of reactor was no further developed. Introduction 79 1 Figure 1.47. Schematic of the AUSB reactor set-up. Reprinted from: Sharma and Huang (2004). Biofilm Airlift Suspension reactors (BAS) (Figure 1.48) were also used for the development of aerobic biofilm and granular biomass (van Loosdrecht et al., 1995). In the top of the reactor, a phase separator allowed separating the biomass from the treated liquid (Figure 1.49 A). Campos et al. (2000) used a similar system for the development of nitrifying granules in the so called Nitrifying Activated Sludge Airlift reactor (NASA) (Figure 1.49 B). Figure 1.48. Schematic representation of the BAS reactor. Reprinted from: van Loosdrecht et al. (1995). Chapter 1 80 1 A) B) Figure 1.49. Design of the tree separator phase: A) BAS design. B) NASA design. Reprinted from: Campos et al. (2000). Tsuneda et al. (2003) used an Aerobic Upflow Fluidized Bed reactor (AUFB) for the development of nitrifying granular biomass. From the bottom of the reactor, wastewater was continuously fed. Aeration was carried out via a porous air diffuser ball at a rate of 1.0 L/min. A solid–liquid separator was placed on the top of the reactor to prevent the outflow of suspended sludge from the reactor (Figure 1.50). Figure 1.50. Schematic illustration of the AUFB reactor. Reprinted from: Tsuneda et al. (2003). More recently, Liu et al. (2012b) developed a new reactor for the growth of aerobic granules: the Continuous-flow bioreactor with aerobic Granular Sludge and Self-Forming Dynamic MemBRane (CGSFDMBR). This system includes four principal reaction tanks: the Introduction 81 1 sequencing batch airlift reactor tank, the settling tank, the dynamic membrane bioreactor tank and the sludge selection tank (Figure 1.51). However, the formation of aerobic granules in that reactor was not completely demonstrated, as the reactor was completely inoculated with already formed granules, the characteristics of the granules worsened after a few weeks of operation, and no biomass concentration data was published. Figure 1.51. Diagram of the CGSFDMBR wastewater treatment process. Reprinted from: Liu et al. (2012b). 1.5.3.4 Scale-up applications The research on aerobic granulation was initially focused on the use of lab-scale reactors fed with synthetic media. Nowadays, the technology has been developed and demonstration plants have been implemented. Tay et al. (2005a) operated a pilot plant for the development of aerobic granular biomass treating a synthetic effluent. The pilot plant reactor had a height of 1.6 m and a diameter of 0.19 m, being the working volume 34 L. Inizan et al., (2005) performed two experiments at pilot scale using a synthetic medium and the effluent of a pharmaceutical company. The reactors had 1.8 m of height and 0.2 m of diameter, with a working volume of 40 L. More recently, a 1 m3 pilot plant was operated by Ni et al. (2009) to treat urban wastewater reaching COD and nitrogen removal percentages of 90-95% (Figure 1.52 A). Scale-up of granular systems leads to modification of the hydrodynamic conditions, which are very important for the formation and maintenance of stability of aerobic granules. Jungles et al. (2011) studied the start-up and performance at different OLRs of a pilot-scale aerobic granular SBR reactor (Figure 1.52 B) fed with synthetic effluent. During the last phase of the experiment, the stability of the system was tested by rapid changes in the applied OLR. The obtained results showed that this technology is suitable to obtain high efficiencies in terms of COD and nutrient removal. These authors also stated that the selection of an adequate Chapter 1 88 1 Etter, B., Tilley, E., Khadka, R. and Udert, K. M. (2011). Low-cost struvite production using sourceseparated urine in Nepal. Water Research 45(2): 852-862. European Environment Agency. (2012a). "Exceedance of critical loads for eutrophication due to the deposition of nutrient nitrogen in 2010 " Retrieved February, 2013, from http://www.eea.europa.eu/data-and-maps/figures/exceedance-of-critical-loads-foreutrophication-due-to-the-deposition-of-nutrient-nitrogen-in-2010. European Environment Agency. (2012b). "Phosphorus concentration in European lakes and reservoirs." Retrieved February, 2013, from http://www.eea.europa.eu/data-andmaps/figures/phosphorus-concentration-in-european-lakes-and-reservoirs. Fabbricino, M., Naviglio, B., Tortora, G. and d'Antonio, L. (2013). An environmental friendly cycle for Cr(III) removal and recovery from tannery wastewater. Journal of Environmental Management 117(0): 1-6. Fattah, K. P., Sabrina, N., Mavinic, D. S. and Koch, F. A. (2008). Reducing operating costs for struvite formation with a carbon dioxide stripper. Water Science and Technology 58: 957-962. Fenu, A., Roels, J., Wambecq, T., De Gussem, K., Thoeye, C., De Gueldre, G. and Van De Steene, B. (2010). Energy audit of a full scale MBR system. Desalination 262(1–3): 121-128. Fernandez-Fontaina, E., Omil, F., Lema, J. M. and Carballa, M. (2012). Influence of nitrifying conditions on the biodegradation and sorption of emerging micropollutants. Water Research 46(16): 5434-5444. Fernandez, I., Vazquez-Padin, J. R., Mosquera-Corral, A., Campos, J. L. and Mendez, R. (2008). Biofilm and granular systems to improve Anammox biomass retention. Biochemical Engineering Journal 42(3): 308-313. Ferriman, A. (2007). BMJ readers choose the “sanitary revolution” as greatest medical advance since 1840. BMJ 334(7585): 111-111. Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Mendez, R. (2008). Treatment of saline wastewater in SBR aerobic granular reactors. Water Science and Technology 58(2): 479-485. Figueroa, M., Vázquez-Padín, J. R., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2012). Is the CANON reactor an alternative for nitrogen removal from pre-treated swine slurry? Biochemical Engineering Journal 65(0): 23-29. Fine, P. and Hadas, E. (2012). Options to reduce greenhouse gas emissions during wastewater treatment for agricultural use. Science of The Total Environment 416(0): 289-299. Foladori, P., Andreottola, G. and Ziglio, G. (2010). Sludge Reduction Technologies in Wastewater Treatment Plants, IWA Publishing. Francis, C. A., Beman, J. M. and Kuypers, M. M. M. (2007). New processes and players in the nitrogen cycle: the microbial ecology of anaerobic and archaeal ammonia oxidation. ISME J 1(1): 19-27. Frijns, J., Hofman, J. and Nederlof, M. (2013). The potential of (waste)water as energy carrier. Energy Conversion and Management 65(0): 357-363. Furukawa, K., Lieu, P. K., Tokitoh, H. and Fujii, T. (2006). Development of single-stage nitrogen removal using anammox and partial nitritation (SNAP) and its treatment performances. Water Science and Technology 53(6): 83-90. Gander, M., Jefferson, B. and Judd, S. (2000). Aerobic MBRs for domestic wastewater treatment: a review with cost considerations. Separation and Purification Technology 18(2): 119-130. Garrido, J. M., Fdz-Polanco, M. and Fdz-Polanco, F. (2013). Working with energy and mass balances: a conceptual framework to understand the limits of municipal wastewater treatment. Water science and technology 67(10): 2294-2301. Geoscience Australia (2013). Long-term phosphate prices since 1960 for phosphate rock (Morocco), 70% BPL, contract, f.a.s. Casablanca. http://www.australianminesatlas.gov.au/aimr/images/phosphate-fig2.jpg Creative Commons Attribution 3.0 Australia Introduction 89 1 Ghaebi, H., Bahadori, M. N. and Saidi, M. H. (2014). Performance analysis and parametric study of thermal energy storage in an aquifer coupled with a heat pump and solar collectors, for a residential complex in Tehran, Iran. Applied Thermal Engineering 62(1): 156-170. Reprinted with permission from Elsevier. Gil, J. A., Túa, L., Rueda, A., Montaño, B., Rodríguez, M. and Prats, D. (2010). Monitoring and analysis of the energy cost of an MBR. Desalination 250(3): 997-1001. Glass, A. D. M. (2003). Nitrogen use efficiency of crop plants: Physiological constraints upon nitrogen absorption. Critical Reviews in Plant Sciences 22(5): 453-470. Gleick, P. H. (2009). "Water Conflict Chronology." Retrieved November 2012, from http://www.worldwater.org/conflict/index.html. Haegh, M., Sj, lin, C., Dahl, C. P. and Hansen, A. T. (2010). Small Footprint Wastewater Treatment Plants. The Combination of Moving Bed Biofilm Reactors with Ballasted Floc Solids Separation. Proceedings of the Water Environment Federation 2010(7): 270-283. HeadworksBIO Advantages & Applications of MBBR Technologies. Heffer, P. and Prud’homme, M. (2010). Fertilizer Outlook 2010 - 2014. 78th IFA Annual Conference, Paris, May - June 2010. Heidrich, E. S., Curtis, T. P. and Dolfing, J. (2010). Determination of the Internal Chemical Energy of Wastewater. Environmental Science and Technology 45(2): 827-832. Heijnen, S. J., Mulder, A., Weltevrede, R., Hols, P. H. and van Leeuwen, H. L. J. M. (1990). Largescale anaerobic/aerobic treatment of complex industrial wastewater using immobilized biomass in fluidized bed and air-lift suspension reactors. Chemical Engineering & Technology 13(1): 202-208. Hellinga, C., Schellen, A. A. J. C., Mulder, J. W., van Loosdrecht, M. C. M. and Heijnen, J. J. (1998). The SHARON process: An innovative method for nitrogen removal from ammoniumrich waste water. Water Science and Technology 37(9): 135-142. Hendrickx, T. L. G., Wang, Y., Kampman, C., Zeeman, G., Temmink, H. and Buisman, C. J. N. (2012). Autotrophic nitrogen removal from low strength waste water at low temperature. Water Research 46(7): 2187-2193. Henze, M., Van Loosdrecht, M. C. M. and Ekama, G. A. (2008). Biological Wastewater Treatment: Principles, Modelling and Design, IWA Pub. Hernández-Sancho, F., Molinos-Senante, M. and Sala-Garrido, R. (2011). Energy efficiency in Spanish wastewater treatment plants: A non-radial DEA approach. Science of the Total Environment 409(14): 2693-2699. Hofman, J., Hofman-Caris, R., Nederlof, M., Frijns, J. and van Loosdrecht, M. C. (2011). Water and energy as inseparable twins for sustainable solutions. Water Science and Technology 63(1): 88-92. Höglund, C. (2001). Evaluation of microbial health risks associated with the reuse of source separated human urine. PhD thesis, Royal Institute of Technology. Honda, S. i., Miyata, N. and Iwahori, K. (2002). Recovery of biomass cellulose from waste sewage sludge. Journal of Material Cycles and Waste Management 4(1): 46-50. Huang, J. C. and Shang, C. (2004). Air Stripping. In: Handbook of Environmental Engineering. Wang, L. K. et al.. (ed), The Humana Press Inc., Totowa, NJ. New Jersey Volume 4: Advanced Physicochemical Treatment Processes: Disclaimer: "Springer and the original publisher is given to the publication in which the material was originally published, by adding; with kind permission from Springer Science and Business Media". Hülsen, T., Batstone, D. J. and Keller, J. (2014). Phototrophic bacteria for nutrient recovery from domestic wastewater. Water Research 50(0): 18-26. Iacovidou, E., Ohandja, D.-G. and Voulvoulis, N. (2012). Food waste co-digestion with sewage sludge – Realising its potential in the UK. Journal of Environmental Management 112(0): 267274. Chapter 1 90 1 Idelovitch, E. and Michail, M. (1981). Nitrogen Removal by Free Ammonia Stripping from High pH Ponds. Journal (Water Pollution Control Federation) 53(9): 1391-1401. IFA. (2006). "Production and International Trade Statistics Available online: http://www.fertilizer.org/ifa/HomePage/STATISTICS/Production-and-trade." International Fertilizer Industry Association Retrieved 21 August 2013. Inizan, M., Freval, A., Cigana, J. and Meinhold, J. (2005). Aerobic granulation in a sequencing batch reactor (SBR) for industrial wastewater treatment. Water Science and Technology 52(10-11): 335-343. Jetten, M. S. M., Wagner, M., Fuerst, J., van Loosdrecht, M., Kuenen, G. and Strous, M. (2001). Microbiology and application of the anaerobic ammonium oxidation ('anammox') process. Current Opinion in Biotechnology 12(3): 283-288. Jiang, F., Beck, M. B., Cummings, R. G., Rowles, K. and Russell, D. (2004). Estimation of costs of phosphorus removal in wastewater treatment facilities. Water Policy Working Paper. Jönsson, H. (2002). Urine separation – Swedish experiences. In: Urban areas – rural areas and recycling – the organic way forward, . Magid, J. et al.. (ed), NJF (Nordic Association of Agricultural Scientists). Tjele. Jungles, M. K., Figueroa, M., Morales, N., Val del Río, A., da Costa, R. H. R., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2011). Start up of a pilot scale aerobic granular reactor for organic matter and nitrogen removal. Journal of Chemical Technology and Biotechnology 86(5): 763-768. Kartal, B., Rattray, J., van Niftrik, L. A., van de Vossenberg, J., Schmid, M. C., Webb, R. I., Schouten, S., Fuerst, J. A., Damsté, J. S., Jetten, M. S. M. and Strous, M. (2007). Candidatus “Anammoxoglobus propionicus” a new propionate oxidizing species of anaerobic ammonium oxidizing bacteria. Systematic and Applied Microbiology 30(1): 39-49. Kartal, B., van Niftrik, L., Rattray, J., de Vossenberg, J. L. C. M. V., Schmid, M. C., Damste, J. S. S., Jetten, M. S. M. and Strous, M. (2008). Candidatus 'Brocadia fulgida': an autofluorescent anaerobic ammonium oxidizing bacterium. Fems Microbiology Ecology 63(1): 46-55. Kartal, B., Maalcke, W. J., de Almeida, N. M., Cirpus, I., Gloerich, J., Geerts, W., Op den Camp, H. J. M., Harhangi, H. R., Janssen-Megens, E. M., Francoijs, K.-J., Stunnenberg, H. G., Keltjens, J. T., Jetten, M. S. M. and Strous, M. (2011). Molecular mechanism of anaerobic ammonium oxidation. Nature 479(7371): 127-130. Kirchmann, H. and Pettersson, S. (1995). Human urine - Chemical composition and fertilizer use efficiency. Fertilizer Research 40(2): 149-154. Koh, L. C., Zhang, H., Nguyen, T. T. and Ong, S. (2012). Performance of an ultra-compact biofilm reactor treating domestic and synthetic wastewaters. Water Science and Technology 65(3): 6. Koutsoyiannis, D., Zarkadoulas, N., Angelakis, A. and Tchobanoglous, G. (2008). Urban water management in ancient Greece: legacies and lessons. Journal of Water resources planning and management 134(1): 45-54. Kozlovský, O., Balík, J., Černý, J., Kulhánek, M., Kos, M. and Prášilová, M. (2009). Influence of nitrogen fertilizer injection (CULTAN) on yield, yield components formation and quality of winter wheat grain. Plant, Soil and Environment 55(12): 536-543. Krzeminski, P., Langhorst, W., Schyns, P., de Vente, D., Van den Broeck, R., Smets, I. Y., Van Impe, J. F. M., van der Graaf, J. H. J. M. and van Lier, J. B. (2012). The optimal MBR configuration: Hybrid versus stand-alone — Comparison between three full-scale MBRs treating municipal wastewater. Desalination 284(0): 341-348. Kuypers, M. M. M., Sliekers, A. O., Lavik, G., Schmid, M., Jorgensen, B. B., Kuenen, J. G., Damste, J. S. S., Strous, M. and Jetten, M. S. M. (2003). Anaerobic ammonium oxidation by anammox bacteria in the Black Sea. Nature 422(6932): 608-611. Larsen, T. A. and Gujer, W. (1996). Separate management of anthropogenic nutrient solutions (human urine). Water Science and Technology. Singapore, Singapore. 34: 87-94. Introduction 91 1 Lazarova, V., Choo, K. H. and Cornel, P. (2012). Water-Energy Interactions in Water Reuse, IWA Publishing. Lee, K. (2013). Aquifer Thermal Energy Storage. In: Underground Thermal Energy Storage. (ed), Springer London. 59-93. Lei, X., Sugiura, N., Feng, C. and Maekawa, T. (2007). Pretreatment of anaerobic digestion effluent with ammonia stripping and biogas purification. Journal of Hazardous Materials 145(3): 391397. Li, C., Champagne, P. and Anderson, B. C. (2011). Evaluating and modeling biogas production from municipal fat, oil, and grease and synthetic kitchen waste in anaerobic co-digestions. Bioresource Technology 102(20): 9471-9480. Libralato, G., Volpi Ghirardini, A. and Avezzù, F. (2012). To centralise or to decentralise: An overview of the most recent trends in wastewater treatment management. Journal of Environmental Management 94(1): 61-68. Lienert, J., Haller, M., Berner, A., Stauffacher, M. and Larsen, T. A. (2003). How farmers in Switzerland perceive fertilizers from recycled anthropogenic nutrients (urine). Water Science and Technology. 48: 47-56. Lienert, J. and Larsen, T. A. (2006). Considering user attitude in early development of environmentally friendly technology: A case study of NoMix toilets. Environmental Science and Technology 40(16): 4838-4844. Liu, D., Wu, X., Chang, J., Gu, B., Min, Y., Ge, Y., Shi, Y., Xue, H., Peng, C. and Wu, J. (2012a). Constructed wetlands as biofuel production systems. Nature Clim. Change 2(3): 190-194. Liu, H., Li, Y., Yang, C., Pu, W., He, L. and Bo, F. (2012b). Stable aerobic granules in continuousflow bioreactor with self-forming dynamic membrane. Bioresource Technology 121(0): 111118. Image reprinted with permission from Elsevier Liu, Q. S., Liu, Y., Tay, S. T. L., Show, K. Y., Ivanov, V., Benjamin, M. and Tay, J. H. (2005). Startup of pilot-scale aerobic granular sludge reactor by stored granules. Environmental Technology 26(12): 1363-1369. Liu, Y. and Tay, J. H. (2002). The essential role of hydrodynamic shear force in the formation of biofilm and granular sludge. Water Research 36(7): 1653-1665. Lofrano, G. and Brown, J. (2010). Wastewater management through the ages: A history of mankind. Science of the Total Environment 408(22): 5254-5264. Ludzack, F. J. and Ettinger, M. B. (1962). Controlling Operation to Minimize Activated Sludge Effluent Nitrogen. Journal Water Pollution Control Federation 34(9): 920-931. M.E.A. (2005). (Millennium Ecosystem Assessment) Ecosystems and human well-being: wetlands and water synthesis, World Resources Institute. Washington, DC Ma, B., Zhang, S., Zhang, L., Yi, P., Wang, J., Wang, S. and Peng, Y. (2011). The feasibility of using a two-stage autotrophic nitrogen removal process to treat sewage. Bioresource Technology 102(17): 8331-8334. Manipura, A. and Burgess, J. E. (2008). Biological nitrogenous compound removal and subsequent rhodium recovery from simulated metal refinery wastewater. Minerals Engineering 21(9): 652659. Marañón, E., Castrillón, L., Quiroga, G., Fernández-Nava, Y., Gómez, L. and García, M. M. (2012). Co-digestion of cattle manure with food waste and sludge to increase biogas production. Waste Management 32(10): 1821-1825. Matsumoto, K. and Funaba, M. (2008). Factors affecting struvite (MgNH4PO4·6H2O) crystallization in feline urine. Biochimica et Biophysica Acta (BBA) - General Subjects 1780(2): 233-239. Maurer, M., Muncke, J. and Larsen, T. (2002). Techniques for nitrogen recovery and reuse. . In: Water and Resource Recovery in Industry. Lens, P. et al.. (ed), IWA Publishing. London, UK. Maurer, M., Pronk, W. and Larsen, T. A. (2006). Treatment processes for source-separated urine. Water Research 40(17): 3151-3166. Chapter 1 92 1 Mbeychok (2007). Rotating_Biological_Contactor. http://commons.wikimedia.org/wiki/File:Rotating_Biological_Contactor.png, Public Domain Mbeychok (2011). Trickle_Filter. http://en.wikipedia.org/wiki/File:Trickle_Filter.svg, Public Domain McQuarrie, J., Rutt, K., Seda, J. and Haegh, M. (2004). Observations from the First Year of Fullscale Operation - The IFAS/BNR Process at the Broomfield Wastewater Reclamation Facility, Broomfield, CO. Proceedings of the Water Environment Federation 2004(7): 274-285. McSwain, B. S., Irvine, R. L. and Wilderer, P. A. (2004). The influence of settling time on the formation of aerobic granules. Water Science and Technology 50 (10): 195-202. Metcalf & Eddy, I., Tchobanoglous, G., Burton, F. and Stensel, H. D. (2002). Wastewater Engineering: Treatment and Reuse, McGraw-Hill Companies, Incorporated. Mishima, K. and Nakamura, K. (1991). Self-immobilization of aerobic activated sludge - a pilot study of the aerobic upflow sludge blanket process in municipal sewage treatment. Water science and technology 23(4-6): 981-990. Molinos-Senante, M., Hernández-Sancho, F. and Sala-Garrido, R. (2011). Cost–benefit analysis of water-reuse projects for environmental purposes: A case study for Spanish wastewater treatment plants. Journal of Environmental Management 92(12): 3091-3097. Montana State University Center for Biofilm Engineering. (2003). "Biofilm formation in 3 steps." Retrieved January 2013, 2013, from http://www.biofilm.montana.edu/node/2390 Reproduced with authorization from Center for Biofilm Engineering. Moralejo-Gárate, H., Mar’atusalihat, E., Kleerebezem, R. and Loosdrecht, M. M. (2011). Microbial community engineering for biopolymer production from glycerol. Applied Microbiology and Biotechnology 92(3): 631-639. Morales, N., Val del Río, A., Vázquez-Padín, J. R., Méndez, R., Mosquera-Corral, A. and Campos, J. L. (2012). Application of the Anammox process to the main stream of WWTPs. 9th Edition Int. Symposium of Sanitary and Environmental Engineering (Sidisa2012). Milano (Italy). Morse, G. K., Lester, J. N. and Perry, R. (1993). The Economic and Environment Impact of Phosphorus Removal from Wasetwater in the European Community. Selper Publications. London. Mosquera-Corral, A., Figueroa, M., Morales, N., Val del Río, Á., Campos, J. L. and Méndez, R. (2012). Aerobic granulation technology In: Innovative Technologies for Urban Wastewater Treatment Plants (2nd Edition) Omil, F. et al.. (ed), Lápices 4, . Santiago de Compostela. Moy, B. Y. P., Tay, J. H., Toh, S. K., Liu, Y. and Tay, S. T. L. (2002). High organic loading influences the physical characteristics of aerobic sludge granules. Letters in Applied Microbiology 34(6): 407-412. Mulder, A., Vandegraaf, A. A., Robertson, L. A. and Kuenen, J. G. (1995). Anaerobic Ammonium Oxidation Discovered in a Denitrifying Fluidized-Bed Reactor. Fems Microbiology Ecology 16(3): 177-183. NACWA (2009). (National Association of Clean Water Agencies), Renewable Energy Recovery Opportunities from Domestic Wastewater. Ni, B. J., Xie, W. M., Liu, S. G., Yu, H. Q., Wang, Y. Z., Wang, G. and Dai, X. L. (2009). Granulation of activated sludge in a pilot-scale sequencing batch reactor for the treatment of low-strength municipal wastewater. Water Research 43(3): 751-761. Nielsen, M., Bollmann, A., Sliekers, O., Jetten, M., Schmid, M., Strous, M., Schmidt, I., Larsen, L. H., Nielsen, L. P. and Revsbech, N. P. (2005). Kinetics, diffusional limitation and microscale distribution of chemistry and organisms in a CANON reactor. Fems Microbiology Ecology 51(2): 247-256. Odegaard, H. (2006). Innovations in wastewater treatment: the moving bed biofilm process. Water Science and Technology 53(9): 17-33. Introduction 93 1 Oh, S. T., Kim, J. R., Premier, G. C., Lee, T. H., Kim, C. and Sloan, W. T. (2010). Sustainable wastewater treatment: How might microbial fuel cells contribute. Biotechnology Advances 28(6): 871-881. Olguín, E. J. (2012). Dual purpose microalgae–bacteria-based systems that treat wastewater and produce biodiesel and chemical products within a Biorefinery. Biotechnology Advances 30(5): 1031-1046. Patwardhan, A. W. (2003). Rotating Biological Contactors: A Review. Industrial & Engineering Chemistry Research 42(10): 2035-2051. Paul, E. and Liu, Y. (2012). Biological Sludge Minimization and Biomaterials/Bioenergy Recovery Technologies, Wiley. Pennisi, E. (2012). Water Reclamation Going Green. Science 337(6095): 674-676. Petrov, S. and Nenov, V. (2004). Removal and recovery of copper from wastewater by a complexation-ultrafiltration process. Desalination 162(0): 201-209. Préndez, M. and Lara-González, S. (2008). Application of strategies for sanitation management in wastewater treatment plants in order to control/reduce greenhouse gas emissions. Journal of Environmental Management 88(4): 658-664. Pruvost, J., Van Vooren, G., Le Gouic, B., Couzinet-Mossion, A. and Legrand, J. (2011). Systematic investigation of biomass and lipid productivity by microalgae in photobioreactors for biodiesel application. Bioresource Technology 102(1): 150-158. Pynaert, K., Smets, B. F., Wyffels, S., Beheydt, D., Siciliano, S. D. and Verstraete, W. (2003). Characterization of an autotrophic nitrogen-removing biofilm from a highly loaded lab-scale rotating biological contactor. Applied and Environmental Microbiology 69(6): 3626-3635. Qiao, S., Matsumoto, N., Shinohara, T., Nishiyama, T., Fujii, T., Bhatti, Z. and Furukawa, K. (2010). High-rate partial nitrification performance of high ammonium containing wastewater under low temperatures. Bioresource Technology 101(1): 111-117. Qin, L., Liu, Y. and Tay, J. H. (2004). Effect of settling time on aerobic granulation in sequencing batch reactor. Biochemical Engineering Journal 21(1): 47-52. Rahaman, M. M. (2012). Water wars in 21st century: speculation or reality? International Journal of Sustainable Society 4(1): 3-10. Regueiro, L., Carballa, M., Álvarez, J. A. and Lema, J. M. (2012). Enhanced methane production from pig manure anaerobic digestion using fish and biodiesel wastes as co-substrates. Bioresource Technology 123(0): 507-513. Rekacewicz, P. (2006). Water Scarcity Index Rijsberman, F. R. (2006). Water scarcity: Fact or fiction? Agricultural Water Management 80(1–3): 5-22. Roberts, T. L. and Stewart, W. M. (2002). Inorganic Phosphorus and Potassium Production and Reserves. Better Crops 86(2): 6-7. Rodríguez-García, G., Hospido, A., Bagley, D. M., Moreira, M. T. and Feijoo, G. (2012). A methodology to estimate greenhouse gases emissions in Life Cycle Inventories of wastewater treatment plants. Environmental Impact Assessment Review 37(0): 37-46. Roest, K., Hofman, J. and Loosdrecht, M. C. M. (2010). The Dutch water cycle can produce energy. H2O 43(25/26): 47-51. Roger Bayley Inc. (2009). Chapter 5, Energy. In: The Challenge Series. The 2010 Winter Olympics: The Southeast False Creek Olympic Village. (ed). Vancouver, Canada: Licensed under a Creative Commons Attribution 2.5 Canada License. Rousseau, P., Steyer, J. P., Volcke, E. I. P., Bernet, N. and Béline, F. (2008). Combined anaerobic digestion and biological nitrogen removal for piggery wastewater treatment: a modelling approach. Water Science and Technology 58 (1): 133-141. Chapter 1 94 1 Sammut, F., Rogalla, F., Goncalves, R. F. and Penillard, P. (1994). Practical experiences with removing nitrogen and phosphorus on aerated biofilters. In: Nutrient Removal from Wastewaters. Horan, N. J. et al.. (ed). Basel, Switzerland: 335–345. Santos, V. A. P. M., Bruijnse, M., Tramper, J. and Wijffels, R. H. (1996). The magic-bead concept: an integrated approach to nitrogen removal with co-immobilized micro-organisms. Applied Microbiology and Biotechnology 45(4): 447-453. Schmid, M., Twachtmann, U., Klein, M., Strous, M., Juretschko, S., Jetten, M., Metzger, J. W., Schleifer, K. H. and Wagner, M. (2000). Molecular evidence for genus level diversity of bacteria capable of catalyzing anaerobic ammonium oxidation. Systematic and Applied Microbiology 23(1): 93-106. Schmid, M., Walsh, K., Webb, R., Rijpstra, W. I. C., van de Pas-Schoonen, K., Verbruggen, M. J., Hill, T., Moffett, B., Fuerst, J., Schouten, S., Damste, J. S. S., Harris, J., Shaw, P., Jetten, M. and Strous, M. (2003). Candidatus "Scalindua brodae", sp nov., Candidatus "Scalindua wagneri", sp nov., two new species of anaerobic ammonium oxidizing bacteria. Systematic and Applied Microbiology 26(4): 529-538. Schwarzenbeck, N., Erley, R. and Wilderer, P. A. (2004). Aerobic granular sludge in an SBRsystem treating wastewater rich in particulate matter. Water Science and Technology 49(1112): 41-46. Selman, M. and Greenhalgh, S. (2009). Eutrophication: Sources and Drivers of Nutrient Pollution. World Resources Institute Policy Note SenterNovem (2006). Watergebruik in Woningen en Warmteterugwinning uit Huishoudelijk Afvalwater (water use in houses and heat recovery from domestic wastewater). Utrecht, Report of TU Delft and Cauberg Huygen. Sharma, K. and Huang, J. (2004). High rate treatment by aerobic upflow sludge blanket (AUSB) with external oxygenation. Water Science & Technology 50(8): 193-200. Shiklomanov, I. (1993). World fresh water resources. In: Water in Crisis: A Guide to the World's Fresh Water Resources. Gleick, P. H. et al.. (ed), Oxford University Press. Siegrist, H., Salzgeber, D., Eugster, J. and Joss, A. (2008). Anammox brings WWTP closer to energy autarky due to increased biogas production and reduced aeration energy for Nremoval. Water Science and Technology 57(3): 383-388. Sinha, B. and Annachhatre, A. (2007). Partial nitrification: operational parameters and microorganisms involved. Reviews in Environmental Science and Biotechnology 6(4): 285313. Sliekers, A. O., Derwort, N., Campos, J. L., Strous, M., Kuenen, J. G. and Jetten, M. S. M. (2002). Completely autotrophic nitrogen removal over nitrite in one single reactor. Water Research 36(10): 2475-2482. Smith, V. H. and Schindler, D. W. (2009). Eutrophication science: where do we go from here? Trends in Ecology & Evolution 24(4): 201-207. Sommer, K., Derivative work: chris and derivative work: Fornax (2010). Funktionsskizze eines Spornrades für die CULTAN-Düngung nach Prof. Karl Sommer. http://commons.wikimedia.org/wiki/File:CULTAN-D%C3%BCngung_Spornrad.svg Creative Commons Attribution-Share Alike 3.0 Unported Steen, I. and Agro, K. (1998). Phosphorus availability in the 21st century: Management of a nonrenewable resource. Phosphorus and Potassium(217): 25–31. Stratful, I., Scrimshaw, M. D. and Lester, J. N. (2004). Removal of struvite to prevent problems associated with its accumulation in wastewater treatment works. Water Environment Research 76(5): 437-443. Strous, M., Van Gerven, E., Zheng, P., Kuenen, J. G. and Jetten, M. S. M. (1997). Ammonium removal from concentrated waste streams with the anaerobic ammonium oxidation (Anammox) process in different reactor configurations. Water Research 31(8): 1955-1962. Introduction 95 1 Strous, M., Heijnen, J. J., Kuenen, J. G. and Jetten, M. S. M. (1998). The sequencing batch reactor as a powerful tool for the study of slowly growing anaerobic ammonium-oxidizing microorganisms. Applied Microbiology and Biotechnology 50(5): 589-596. Strous, M., Fuerst, J. A., Kramer, E. H. M., Logemann, S., Muyzer, G., van de Pas-Schoonen, K. T., Webb, R., Kuenen, J. G. and Jetten, M. S. M. (1999a). Missing lithotroph identified as new planctomycete. Nature 400(6743): 446-449. Strous, M., Kuenen, J. G. and Jetten, M. S. M. (1999b). Key physiology of anaerobic ammonium oxidation. Applied and Environmental Microbiology 65(7): 3248-3250. Strous, M., Kuenen, J. G., Fuerst, J. A., Wagner, M. and Jetten, M. S. M. (2002). The anammox case - A new experimental manifesto for microbiological eco-physiology. Antonie Van Leeuwenhoek International Journal of General and Molecular Microbiology 81(1-4): 693-702. Suriyachan, C., Nitivattananon, V. and Amin, A. T. M. N. (2012). Potential of decentralized wastewater management for urban development: Case of Bangkok. Habitat International 36(1): 85-92. Sustainable Sanitation (2013). Instruction poster for urine-diverting dry toilet (UDDT) in low-income area Bulbul near Nairobi, Kenya. http://www.flickr.com/photos/gtzecosan/10543355195/sizes/o/in/set-72157637061451965/ Creative Commons Attribution 2.0 Generic Tay, J. H., Liu, Q. S. and Liu, Y. (2001a). The role of cellular polysaccharides in the formation and stability of aerobic granules. Letters in Applied Microbiology 33: 222-226. Tay, J. H., Liu, Q. S. and Liu, Y. (2001b). The effects of shear force on the formation, structure and metabolism of aerobic granules. Applied Microbiology and Biotechnology 57(1-2): 227-233. Tay, J. H., Ivanov, V., Pan, S. and Tay, S. T. L. (2002). Specific layers in aerobically grown microbial granules. Letters in Applied Microbiology 34(4): 254-257. Tay, J. H., Liu, Q. S., Liu, Y., Show, K., Ivanov, V. and Tay, S. T. L. (2005a). A comparative study of aerobic granulation in pilotand laboratory-scale SBRs. In: Aerobic Granular Sludge. (ed), IWA Publishing. Munich: 125-133. Tay, S. T. L., Moy, B. Y. P., Jiang, H. L. and Tay, J. H. (2005b). Rapid cultivation of stable aerobic phenol-degrading granules using acetate-fed granules as microbial seed. Journal of Biotechnology 115(4): 387-395. Tchobanoglous, G. and Leverenz, H. (2009). Impacts of New Concepts and Technology on the Energy Sustainability of Wastewater Management. Climate Change, Sustainable Development and Renewable Resources in Greece. Thessaloniki, Greece. Tchobanoglous, G. (2012). Wastewater treatment trends in the 21st century. The 3rd International Sede Boqer Conference on Water Technologies 2012 Advanced Technologies in Water Management, Ben-Gurion University of the Negev, Israel. Thamdrup, B. and Dalsgaard, T. (2002). Production of N-2 through anaerobic ammonium oxidation coupled to nitrate reduction in marine sediments. Applied and Environmental Microbiology 68(3): 1312-1318. The USGS Water Science School (2013). The World's Water. http://ga.water.usgs.gov/edu/earthwherewater.html, Public Domain Tijhuis, L., Hijman, B., Loosdrecht, M. C. M. and Heijnen, J. J. (1996). Influence of detachment, substrate loading and reactor scale on the formation of biofilms in airlift reactors. Applied Microbiology and Biotechnology 45(1-2): 7-17. Trigo, C., Campos, J. L., Garrido, J. M. and Méndez, R. (2006). Start-up of the Anammox process in a membrane bioreactor. Journal of Biotechnology 126(4): 475-487. Trubetskoy, S. (2012). Eutrophication at a waste water outlet in the Potomac River, Washington, D.C. http://commons.wikimedia.org/wiki/File:Potomac_green_water.JPG Creative Commons Attribution-Share Alike 3.0 Unported Chapter 1 96 1 Tsuneda, S., Nagano, T., Hoshino, T., Ejiri, Y., Noda, N. and Hirata, A. (2003). Characterization of nitrifying granules produced in an aerobic upflow fluidized bed reactor. Water Research 37(20): 4965-4973. Image reprinted with permission from Elsevier U.N. (2009). World Population Prospects: The 2008 Revision. United Nations Department of Economic and Social Affairs, Population Division. Population Newsletter 87: 1-3. U.S. Department of Energy (2013). Diagram of a drain water heat recovery system. http://energy.gov/sites/prod/files/styles/borealis_article_hero_respondlarge/public/drainwater_ heat_recover.gif, Public Domain U.S. EPA. (2008). Chapter 4. Land. In: EPA's 2008 Report on the Environment (Final Report). (ed), U.S. Environmental Protection Agency, Washington, D.C. USA. Public Domain. U.S. Geological Survey. (2012). "USGS Minerals Information: Mineral Commodity Summaries." Retrieved February, 2013, from http://www.minerals.usgs.gov/minerals/pubs/mcs/. Udert, K. M., Larsen, T. A., Biebow, M. and Gujer, W. (2003). Urea hydrolysis and precipitation dynamics in a urine-collecting system. Water Research 37(11): 2571-2582. Udert, K. M., Larsen, T. A. and Gujer, W. (2006). Fate of major compounds in source-separated urine. Water Science and Technology 54: 413-420. Ueno, Y. and Fujii, M. (2001). Three years experience of operating and selling recovered struvite from full-scale plant. Environmental Technology 22(11): 1373-1381. UNEP (2011). Year book 2011: Emerging issues in our global environment, United Nations Environment Programme. United Nations World Water Assessment Programme , UN-Water and UNESCO (2009). Water in a changing world, UNESCO Publishing. Paris Val del Río, A., Morales, N., Isanta, E., Mosquera-Corral, A., Campos, J. L., Steyer, J. P. and Carrère, H. (2011). Thermal pre-treatment of aerobic granular sludge: Impact on anaerobic biodegradability. Water Research 45(18): 6011-6020. van der Star, W., Miclea, A. I., van Dongen, U., Muyzer, G., Picioreanu, C. and van Loosdrecht, M. C. M. (2008). The membrane bioreactor: A novel tool to grow anammox bacteria as free cells. Biotechnology and Bioengineering. van der Star, W. R. L., Abma, W. R., Blommers, D., Mulder, J.-W., Tokutomi, T., Strous, M., Picioreanu, C. and van Loosdrecht, M. C. M. (2007). Startup of reactors for anoxic ammonium oxidation: Experiences from the first full-scale anammox reactor in Rotterdam. Water Research 41(18): 4149-4163. van Dongen, U., Jetten, M. S. M. and van Loosdrecht, M. C. M. (2001). The SHARON-Anammox process for treatment of ammonium rich wastewater. Water Science and Technology 44(1): 153-160. van Kempen, R., Mulder, J. W., Uijterlinde, C. A. and Loosdrecht, M. C. M. (2001). Overview: full scale experience of the SHARON process for treatment of rejection water of digested sludge dewatering. Water Science and Technology 44(1): 145-152. van Loosdrecht, M. C. M., Eikelboom, D., Gjaltema, A., Mulder, A., Tijhuis, L. and Heijnen, J. J. (1995). Biofilm structures. Water Science and Technology 32(8): 35-43. van Loosdrecht, M. C. M., Hao, X., Jetten, M. S. M. and Abma, W. (2004). Use of Anammox in urban wastewater treatment. Water Science and Technology: Water Supply 4(1): 8. van Niftrik, L. and Jetten, M. S. M. (2012). Anaerobic Ammonium-Oxidizing Bacteria: Unique Microorganisms with Exceptional Properties. Microbiology and Molecular Biology Reviews 76(3): 585-596. Vázquez-Padín, J. R., Fernández, I., Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2009a). Applications of Anammox based processes to treat anaerobic digester supernatant at room temperature. Bioresource Technology 100(12): 2988-2994. Introduction 97 1 Vázquez-Padín, J. R., Figueroa, M., Fernández, I., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2009b). Post-treatment of effluents from anaerobic digesters by the Anammox process. Water science and technology 60(5): 1135-1143. Vázquez-Padín, J. R., Pozo, M. J., Jarpa, M., Figueroa, M., Franco, A., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2009c). Treatment of anaerobic sludge digester effluents by the CANON process in an air pulsing SBR. Journal of Hazardous Materials 166(1): 336-341. Vázquez-Padín, J. R., Figueroa, M., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2010a). Nitrifying granular systems: A suitable technology to obtain stable partial nitrification at room temperature. Separation and Purification Technology 74(2): 178-186. Vázquez-Padín, J. R., Mosquera-Corral, A., Campos, J. L., Méndez, R. and Revsbech, N. P. (2010b). Microbial community distribution and activity dynamics of granular biomass in a CANON reactor. Water Research 44(15): 4359-4370. Vázquez-Padín, J. R., Fernández, I., Morales, N., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2011). Autotrophic nitrogen removal at low temperature. Water science and technology 63(6): 1282-1288. Vázquez-Padín, J. R. (2013). Personal Communication. Velela (2005). Trickling filter bed 2 w. http://commons.wikimedia.org/wiki/File:Trickling_filter_bed_2_w.JPG Creative Commons Attribution-Share Alike 3.0 Unported Verawaty, M., Tait, S., Pijuan, M., Yuan, Z. and Bond, P. L. (2013). Breakage and growth towards a stable aerobic granule size during the treatment of wastewater. Water Research 47(14): 5338-5349. Image reprinted with permission from Elsevier. Verrecht, B., Maere, T., Nopens, I., Brepols, C. and Judd, S. (2010). The cost of a large-scale hollow fibre MBR. Water Research 44(18): 5274-5283. Vråle, L. and Jenssen, P. D. (2005). Source separated wastewater a new resource for producing mineral fertilizer. 3rd. International Conferenc on Ecological Sanitation. Durban. Wang, J., Burken, J. G., Zhang, X. and Surampalli, R. (2005). Engineered struvite precipitation: Impacts of component-ion molar ratios and pH. Journal of Environmental Engineering 131(10): 1433-1440. Wang, L. P., Wong, H. Y. and Griffith, D. P. (1997). Treatment options in struvite stones. Urologic Clinics of North America 24(1): 149-162. Webb, K. M. and Ho, G. E. (1992). Struvite (MgNH4PO4·6H2O) solubility and its application to a piggery effluent problem. Water Science and Technology 26(9-11): 2229-2232. Webb, R. (2012). Algal bloom warning, Loch Melvin. http://s0.geograph.org.uk/geophotos/03/05/98/3059829_3170e9c0.jpg Creative Commons Licence. Attribution-ShareAlike 2.0 Generic Webster, C. (1962). The Sewers of Mohenjo-Daro. Journal Water Pollution Control Federation 34(2): 116-123. Wellstead, J. (2012). "Political Risks in MENA Phosphate Markets." Retrieved March, 2013, from http://potashinvestingnews.com/4799-political-risks-mena-phosphate-markets-tunisiamorocco-jordan-syria-afric.html. WERF (2011). Energy Management-Exploratory Team Report, Water Environment Research Foundation. Werktuigendagen (2009). John Deere 7430 with Joskin liquid manure trailer at Werktuigendagen 2009. http://commons.wikimedia.org/wiki/File:John_Deere_7430.jpg Creative Commons Attribution-Share Alike 2.0 Generic Wett, B. (2006). Solved upscaling problems for implementing deammonification of rejection water. Water Science and Technology 53(12): 121-128. Wett, B. (2007). Development and implementation of a robust deammonification process. Water Science and Technology 56(7): 81-88. Chapter 2 104 2 2.1.1.2 Determination procedure This procedure is applicable to samples with COD concentrations between 90-900 mg/L. Place 2.5 mL of sample in 10-mL Pyrex tubes. Add 1.5 mL of digestion solution and 3.5 mL of sulphuric acid reagent slowly on the wall of the tube slightly inclined (to avoid mixing). A blank sample using distilled water is prepared in the same way. This blank acts as “reference”, representing the COD of the distilled water. After being sealed with Teflon and tightly capped, the tubes are finally mixed completely and placed in the block digester (HACH 16500-100) preheated to 150 ºC. The duration of the digestion period is 2 h. After digestion, the tubes are cooled to room temperature. Then, the content of the tubes is transferred into a beaker and, once added 1-2 drops of ferroin indicator, the solution is titrated under rapid stirring with standard FAS. The FAS solution is standardized daily as follows: Place 5 mL of distilled water into a small beaker. Add 3.5 mL of sulphuric acid reagent. Cool to room temperature and add 5 mL of standard potassium dichromate solution (0.05 N). Add 1-2 drops of ferroin indicator and titrate with FAS titrant. The end-point is a sharp color change from blue-green to reddish brown. Molarity of FAS solution is calculated using the equation (2.1). fas fas 5 · 0.05 MV  (2.1) where: Mfas: molarity of FAS (mol/L), and Vfas: volume of FAS consumed in the titration (mL). The COD is calculated with the equation (2.2):   fas A B · M · 8000 COD V   (2.2) where: COD: Chemical Oxygen Demand (mg O2/L); A: mL of FAS solution consumed by the blank; B: mL of FAS solution consumed by the sample; 8000: milliequivalent weight of oxygen x 1000 mL/L. Materials and methods 105 2 2.1.2 Total Organic Carbon (TOC) Organic carbon in liquid samples may include a variety of organic compounds in different oxidation states. Total Organic Carbon (TOC) is a more convenient and direct expression of total organic content than Chemical Oxygen Demand (COD), but does not provide the same information. Unlike COD, TOC is independent of the oxidation state of the organic matter and does not measure other organically bound elements, such as nitrogen and hydrogen, and inorganic compounds that can contribute to the oxygen demand measured by COD (APHA-AWWA-WPCF, 2005). To determine the quantity of organically bound carbon, the organic molecules must be broken down and converted to a single carbon molecular form that can be measured quantitatively. The TOC concentration was determined by a Shimadzu analyzer (TOC-5000) as the difference between the Total Carbon (TC) and the Inorganic Carbon (IC) concentrations. The instrument was connected to an automated sampler (Shimadzu, ASI-5000-S). The TC concentrations were determined from the amount of CO2 produced during the combustion of the sample at 680 °C, using platinum immobilized over alumina spheres as catalyst. The IC concentrations were obtained from the CO2 produced during the chemical decomposition of the sample with H3PO4 (25%) at room temperature. The CO2 produced was optically measured with a non dispersive infrared analyzer (NDIR) after being cooled and dried. High-purity air was used as carrier gas at a flow of 150 mL/min. A curve comprising 4 calibration points in the range of 0 to 1 g C /L, using potassium phthalate as standard for TC and a mixture of sodium carbonate and bicarbonate (Na2CO3/NaHCO3, 3:4 w/w) for IC, was used for the quantification. 2.1.3 Nitrogen In waters and wastewaters, the forms of nitrogen of greater interest are, in order of decreasing oxidation state, nitrate, nitrite, ammonia and organic nitrogen. All these forms, as well as nitrogen gas (N2), are biochemically inter-convertible following the processes of the nitrogen cycle. Organic nitrogen is defined functionally as the organically bound nitrogen in the trinegative oxidation state, but it does not include all organic nitrogen compounds. Analytically, organic nitrogen and ammonia can be determined together and have been referred to as “Total Kjeldahl Nitrogen” (TKN), a term that reflects the technique used in their determination. Total inorganic nitrogen (IN) is the sum of the nitrate and nitrite forms. 2.1.3.1 Total (TN), Inorganic (IN) and Total Kjeldhal Nitrogen (TKN) TKN was determined in a total organic nitrogen analyzer (Rosemount-Dohrmann DN1900) equipped with a quimio-luminiscence detector with two channels. One channel determined the Total Nitrogen (TN), by oxidation at high temperature, and the other determined the Inorganic Nitrogen (IN), by a chemical reduction. TKN was determined as the difference between TN and IN. All the nitrogen presented in the sample was catalytically oxidized to nitrous oxide (NO). Chapter 2 106 2 The process for TN determination occurred in two steps. The first step was a catalytic (Cu as catalyst) oxidation in the combustion tube at 850 °C and with pure oxygen (1 atm) as carrier gas. The second one was the chemical reduction of residual NO2 with H2SO4 at 80 ºC and catalyzed by VaCl3. For the IN determination, only the second step (chemical reduction) was used. The NO obtained in the two steps was dried and forced to react with O3 producing an unstable excited state NO2*. The change back of this oxide to its fundamental state released a proton, from which the determination of TN and IN was carried out by quimioluminiscence, using a multiplicator tube. The instrument was calibrated with a certified standard solution (KNO3, 20 mg N/L) using a response factor method. 2.1.3.2 Ammonia, method Bower, Holm-Hansen Total ammonia-nitrogen (NH4+-N) was determined spectrophotometrically by a method in which indophenol blue was produced by the reaction of ammonia with salicylate and hypochlorite, in the presence of sodium nitroprusside (Bower and Holm-Hansen, 1980). This method is safer than the phenol–hypochlorite method, 4500‐NH3 F of (APHA-AWWA-WPCF, 2005), because phenol is not used, but it may not be suitable for field determinations because of photosensitivity. The characteristic colors produced by increasing concentrations of ammonia make the assay useful for the direct, visual estimation of ammonia in culture systems. Reagents: • Reagent A: Solution of 0.28 g/L of sodium nitroprusside and 440 g/L of sodium salicylate. • Reagent B: Solution of 18.5 g/L of NaOH and 120 g/L of sodium citrate. • Reagent C: Standard commercial solution of sodium hypochlorite. • Reagent D: Solution prepared mixing 7 parts of reagent B and 1 part of reagent C. Reagent D was stable for 1 hour after preparation. Determination Procedure: • Add 120 μL of reagent A and 200 μL of reagent D to 1 mL of sample (diluted if necessary). • Store, protected from light, for more than 2 hours but less than 3 hours. • Measure it at 640 nm and compare the obtained absorbance to a calibration curve. 2.1.3.3 Ammonia, phenate method Ammonia concentration was determined following the method 4500-NH3 F. (Phenate method) (APHA-AWWA-WPCF, 2005). The method is based on the reaction of NH3 with HClO Materials and methods 107 2 and phenol, forming a strong-blue compound (indophenol) which can be colorimetrically determined using a spectrophotometer (Shimadzu UV1603, UV-Visible) at 635 nm. Reagents preparation: • Solution 1 (Phenol-nitroprusiate): 15 g of phenol and 0.05 g of sodium nitroprusiate were added to 250 mL of buffer solution. The buffer solution was prepared adding 30 g of Na3PO4·12 H2O, 30 g Na3C6H5O7·2 H2O and 3 g EDTA per liter, adjusted to pH 12. • Solution 2 (Hipochloride): 15 mL of commercial bleach were mixed with 200 mL of NaOH 1 N and filled up to 500 mL with distilled water. Determination procedure: 2.5 mL of sample (diluted if necessary to get a maximum concentration of 1 mg NH4+- N/L) were placed in a 10 mL tube. 1.0 and 1.5 mL of solution 1 and 2, respectively were added to the sample. After waiting 45 minutes at room temperature, the concentration of NH4+-N was measured in a spectrophotometer (CECIL7200) at 635 nm. The quantification was done with a 5-7 points calibration curve in the range of 0-1 mg NH4+-N/L, using NH4Cl as standard (Figure 2.1). Figure 2.1. Calibration curve for ammonia determination by the phenate method. 2.1.3.4 Nitrite Nitrite concentration was determined following the method 4500-NO2--B (Colorimetric Method) described in the Standard Methods for the Examination of Water and Wastewater (2005). Nitrite is determined through the formation of a reddish purple azo dye produced at pH y = 1.2339x -0.0408 R² = 0.9998 0.0 0.2 0.4 0.6 0.8 1.0 1.2 00.2 0.4 0.6 0.8 1 NH4+-N Absorbance Chapter 2 108 2 2.0-2.5 by coupling diazotized sulphanilamide with N-(1-napththyl)- ethylenediamine dihydrochloride (NED dihydrochloride). Reagents preparation: • Sulphanilamide: 10 g of sulphanilamide were dissolved in 100 mL of concentrated HCl and 600 mL of distilled water. After cooling, the volume was filled up to 1 L with distilled water. • NED: 0.5 g of NED were dissolved in 500 mL of distilled water. Determination procedure: A volume of 0.1 mL of each reagent was added to 5 mL of sample, diluted if necessary to fit the concentration range of the method. After waiting 20 minutes for color stabilization, the sample was measured in a spectrophotometer (CECIL-7200) at 543 nm. The quantification was done with 6-8 points calibration curve in the range of 0-0.25 mg NO2--N/L, using NaNO2 as standard (Figure 2.2). Figure 2.2. Calibration curve for nitrite. 2.1.3.5 Nitrate Nitrate concentration was determined following the method 4500‐NO3-‐B (Ultraviolet Spectrophotometric Screening Method) described in the Standard Methods for the Examination of Water and Wastewater (2005). Measurement of UV absorption at 220 nm enables rapid determination of NO3ions. Because dissolved organic matter also may absorb at 220 nm and NO3does not absorb at 275 nm, a second measurement at 275 nm is used to correct the NO3value. y = 0.2823x -0.0016 R² = 0.9995 0.00 0.05 0.10 0.15 0.20 0.25 0.30 00.2 0.4 0.6 0.8 1 NO2--N Absorbance Materials and methods 109 2 Determination procedure: A volume of 0.1 mL of HCl 1 N was added to 5 mL of sample (diluted if necessary to get a maximum concentration of NO3--N of 2.5 mg/L). Afterwards, the absorbance at 220 and 275 nm was measured in a spectrophotometer (CECIL-7200). The absorbance related to nitrate was obtained by subtracting two times the absorbance reading at 275 nm from the reading at 220 nm. The quantification was done with a 6-8 points calibration curve within the range of 0-3 mg NO3--N/L, using KNO3 as standard (Figure 2.3). Figure 2.3. Calibration curve for nitrate. 2.1.4 Inorganic ions The anions nitrite (NO2–), nitrate (NO3–), chloride (Cl–), bromide (Br–), phosphate (PO4–), sulphate (SO4–), thiosulphate (S2O3–2) and the cations lithium (Li+), sodium (Na+), ammonium (NH4+), potassium (K+), magnesium (Mg2+) and calcium (Ca2+) were determined by ion chromatography (IC) with an Advanced Compact IC system (861, Metrohm) equipped with a CO2 suppressor (MCS 853, Metrohm) and a sample processor (AG 838, Metrohm). Anions were determined with a “Metrosep A column” (250 x 4.0 mm) and a mobile phase (buffer) with 3.2 mM Na2CO3 and 1.0 mM NaHCO3 at a flow rate of 0.7 mL/min. Cations were determined with a column (250 x 4.0 mm) (Metrosep C3, Metrohm) and nitric acid 3.5 mM was used as mobile phase. The injection volume of the sample was 20 μL and data collection was done by using the Processor software IC Net 2.3. y = 4.0933x -0.057 R² = 0.997 0 0.2 0.4 0.6 0.8 1 1.2 00.05 0.1 0.15 0.2 0.25 0.3 0.35 NO3--N Abs1-2*Abs2 Chapter 2 110 2 Reagents: • Mobile phase for anions: Na2CO3 3.2 mM (339.2 mg Na2CO3 in 1000 mL of deionised water) and NaHCO3 1.0 mM (84 mg NaHCO3 in 1000 mL of distillated water). • Mobile phase for cations: Nitric acid 3.5 mM (0.243 mL of nitric acid 65% in 1000 mL of distillated water). • Standard commercial solutions for anions and cations (Fluka). Determination Procedure: Table 2.1 shows the calibration range for the different inorganic ions’ concentrations. Therefore, in some samples, a dilution with distilled water was performed in order to fit to these ranges. Table 2.1. Calibration ranges for the different inorganic ions (mg/L). Anion Low value High value Cation Low value High value Cl – 1.0 100 Li + 0.05 5 NO2– 0.05 5 Na + 1.5 150 NO3– 0.5 50 NH4+ 0.1 10 Br– 0.2 20 K+ 0.5 50 PO43– 0.5 50 Mg2+ 0.5 50 SO42– 1.5 150 Ca2+ 0.5 50 S 2 O32– 1.5 150 2.1.5 Other control parameters 2.1.5.1 pH The pH value was measured using different instruments in the laboratory and pilot plant scale reactors. In each experimental setup was specified the on-line device. The pH measurements of the liquid samples were performed with an electrode (52‐03, Crison Instruments) equipped with an automatic compensatory temperature device (21‐910‐01, Crison Instruments) and connected to a measurement instrument (pH). The sensibility of the instrument was ±1 mV, corresponding to 0.01 pH units. The electrode was calibrated at room temperature with two standard buffer solutions of pH 7.02 and 4.00. Materials and methods 111 2 2.1.5.2 Dissolved oxygen (DO) Dissolved oxygen was measured using different instruments in the laboratory and pilot plant scale reactors. The specific used device is specified in the corresponding chapter. • Dissolved oxygen pocket meter (Oxi 330i, WTW) equipped with a membrane covered galvanic dissolved oxygen sensor (CellOx® 325, WTW). • Digital multimeter (HQ40D, Hach Lange) equipped with a Luminescencebased DO probe (Intellical LDO, Hach Lange). • A Luminescence-based DO probe (LDO Process sensor, Hach Lange) connected to a controller (SC-100, Hach Lange). 2.2 BIOMASS CHARACTERIZATION 2.2.1 Solids concentrations (TSS and VSS) Total Suspended Solids (TSS) and Volatile Suspended Solids (VSS) concentrations were determined according to the methods 2540 D (Total Suspended Solids Dried at 103‐105 °C) and 2540 E (Fixed and Volatile Solids Ignited at 550 °C) reported in Standard Methods for the Examination of Water and Wastewater (APHA-AWWA-WPCF, 2005). 2.2.2 Sludge Volume Index The Sludge Volume Index (SVI) determination is defined in the Standard Methods for the Treatment of Water and Wastewater (APHA-AWWA-WPCF, 2005) as the volume in milliliters occupied by 1 g of a suspension after 30 minutes of settling. However, as suggested at the “1st IWA‐Workshop Aerobic Granular Sludge” (Munich, 2004) and by Schwarzenbeck et al., (2004) another parameter, the SVI10 (SVI after 10 minutes of settling) was used in all the chapters of this work instead of SVI30 (SVI after 30 minutes of settling) since it is more representative for granular biomass (de Kreuk et al., 2007). A low SVI30 value does not necessarily imply sludge granulation and vice versa. Nevertheless, a granular sludge bed does consolidate much faster, i.e., the terminal SVI30 is already reached after 10 minutes of settling. 2.2.3 Density of the granules The biomass density (as mass of granules per volume of granules) was determined using the method described by Beun et al., (2002) and modified in the laboratory of Environmental Engineering and Bioprocesses. First, a known amount of a homogeneous biomass sample was taken from the reactor and weighed (W1) in a tare weighed graduated cylinder (W2). Then, a known amount of liquid was removed from the sample (W3). A known volume of a dextran blue solution (1 g/L) was added to a known volume of sample of granular sludge, in a volume ratio of about 1:1. The mixture was gently mixed, and subsequently, the Chapter 2 112 2 granules were allowed to settle (W4). A known amount of the liquid above the settled granules was removed, and a sample was taken from it. This fraction (Abs1) and the original dextran blue solution (Abs0) were analyzed by a spectrophotometer at 620 nm. Subsequently the volume occupied by the biomass in the reactor sample was calculated, since dextran blue only binds to water and not to biomass. By measuring also the dry weight of the reactor sample (APHA-AWWA-WPCF, 2005) the density of the granules can be calculated as gram of biomass per liter of granule. The density was calculated from equation (2.3):   12 granule 0 42 43 1 WW VSS Abs WW · WW) Abs           (2.3) Being: ρgranule: The density of the granules (g VSS/Lgranule) VSS: Volatile Suspended Solids concentration (g/L) W1: weight of the graduated cylinder with sample (g) W2: graduated cylinder weight (g) W3: weight of the graduated cylinder with sample after removal of liquid (g) W4: weight of the graduated cylinder after dextran blue addition (g) Abs0: Absorbance of the dextran blue solution (1 g/L) Abs1: Absorbance of the sample 2.2.4 Average diameter of the granules Morphology of the granules was followed by image analysis. Images of the granular sludge were taken with a digital camera (Coolsnap, Roper Scientific Photometrics) combined to a stereomicroscope (Stemi 2000‐C, Zeiss). For digital image analysis the software Image ProPlus® was used (Figure 2.4). The procedure of average diameter determination is as follows: Materials and methods 113 2 Figure 2.4. Example of the use of Image ProPlus® software. • Definition of the range of colors corresponding to the area of interest in the image, i.e. the granules (Manual or Automatic). • Selection of the measurements of interest. • Exporting of the data of interest selected with the software (e.g., area, aspect, roundness, average diameter, etc.) to a worksheet. • Utilization of the histogram tool (data analysis tool pack) to calculate the frequency and construction of the histogram. The average diameter can be calculated from a frequency, surface or volumetric distribution. The average diameter obtained from the software corresponded to the mean feret diameter of the granules. The feret diameter was calculated as an average value from the shortest and the longest measured segments of each granule. 2.2.5 Elemental analysis The elemental analysis of the biomass was performed in order to know the composition as CnHaObNcSd. First, the sample was dried at 105 ºC for 24 hours, and then it was crushed to obtain a homogeneous powder. The quantity of sample necessary was between 1-3 mg. The elemental analysis technique is based on the complete and instantaneous oxidation (combustion) of the sample and the determination of the gases from the combustion through a thermal conductivity detector model CHNS FISONS EA 1108 (for C, H, N and S) and model CARLO ERBA EA 1108 (for oxygen). The results were expressed as percentage of compound in the sample. Chapter 2 120 2 Probe Target site 16S Probe sequence (5’→3’) % FA Target organisms Ref. a NEU653 653-670 CCC CTC TGC TGC ACT CTA 40 Most halophilic and halotolerant Nitrosomonas spp. [4] Competitor TTC CAT CCC CCT CTG CCG Nso190 189-207 CGA TCC CCT GCT TTT CTC C 55 Ammonio-oxidizing-βProteobacteria [5] Nsm156 156-174 TAT TAG CAC ATC TTT CGA T 5 Nitrosomonas spp., Nitrosococcus mobilis [5] NIT3 Competitor 1035-1052 CCT GTG CTC CAT GCT CCG CCT GTG CTC CAG GCT CCG 40 Nitrobacter spp. [6] Ntspa712 Competitor 712-732 CGC CTT CGC CAC CGG CCT TCC CGC CTT CGC CAC CGG GTT CC 35 Most members of phylum Nitrospira [7] PLA46 46-63 GAC TTG CAT GCC TAA TCC 20 Planctomycetes [8] Amx820 820-841 AAA ACC CCT CTA CTT AGT GCC C 35 Candidatus “Brocadia anammoxidans” [9] Kst157 157-174 GTT CCG ATT GCT CGA AAC 25 Candidatus Kuenenia stuttgartiensis [9] Ban162 162-179 CGG TAG CCC CAA TTG CTT 40 Candidatus Brocadia anammoxidans [9] Pae997 997-1014 TCT GGA AAG TTC TCA GCA 0 Pseudomonas spp. [10] PAR1244 1244-1262 GGA TTA ACC CAC TGT CAC C 20 Paracoccus [11] a References: [1] Amann et al., (1990); [2] Daims et al., (1999); [3] Manz et al., (1992); [4] Wagner et al., (1995); [5] Mobarry et al., (1996); [6] Wagner et al., (1996); [7] Daims et al., (2001); [8] Neef et al., (1998); [9] Schmid et al., (2001); [10] Amann et al., (1996); [11] Neef et al., (1996). 2.4 CALCULATIONS 2.4.1 SBR operational cycle analysis The activity of the different microbial populations presented in the SBR was calculated using the concentration profiles during a whole cycle of operation, based on the procedure described by Mosquera‐Corral et al., (2005b) and according to the representation of Figure 2.7. Materials and methods 121 2 Figure 2.7. Schematic representation of the nitrogen compounds concentrations during the cycle and definition of the calculated parameters. The consumption rates for ammonia (rNH4+) and for nitrogen oxides (nitrite or nitrate) (rNOx f) in the feast phase and the production rate of nitrogen oxides (nitrite or nitrate) (rNOX h) during famine phase were calculated using the equations (2.6), (2.7) and (2.8). 44 4 c 02 NH NH NH 20 CC r(t t )60     (2.6) xx x f 01 NO NO NO 10 CC r(t t )60     (2.7) xx x h 12 NO NO NO 12 CC r(t t )60     (2.8) being t0 the time at the beginning of the cycle, t1 the time at the end of the feast phase, t2 the time at the end of the famine phase and Cct the concentration of each compound (c) in a certain time (t), expressed in mg N/L. Specific rates were calculated by dividing the consumption rates by the solids concentration. 2.4.2 Estimation of the nitrogen assimilated and denitrified. The main processes for nitrogen removal in the aerobic granular SBR were nitrogen assimilation for biomass growth and nitrification‐denitrification. The sum of the amount of nitrogen removed from the liquid phase by both ways provides the value of nitrogen removed (NR). In order to discern between the percentages of nitrogen removal achieved by each of these mechanisms a nitrogen balance was performed to the reactor to determine the amount Chapter 2 122 2 of nitrogen used for growth. For each selected period, the amount of biomass produced was estimated from the biomass increase in the reactor and the amount of biomass washed out in the effluent using equation (2.9) Eff p rr W X ·V X ·Q· t    (2.9) Where ΔWP is the amount of produced biomass (g VSS), ΔXr the change of biomass concentration during each period (g VSS/L), Vr the reactor volume (L),XEff the average biomass concentration washed out in the effluent (g VSS/L), Q the flow rate (L/d) and Δt the length of each period (d). Considering a general composition of the biomass as C5H7NO2, the averaged amount of nitrogen assimilated for biomass growth (Nassimilated, g N) was calculated using equation (2.10) as: assimilated p 14 g-mol N NW 113 g-mol biomass  (2.10) Nremoved (g N) during such period of operation was calculated using equations and, being ΔN the nitrogen balance, NH4+-NInf, NO2–-NInf, and NO3–-NInf the ammonium, nitrite, nitrate (mg N/L) in the influent, and NH4+-NEff, NO2–-NEff, and NO3–-NEff the ammonium, nitrite and nitrate concentrations in the effluent respectively.     4 Inf 2 Inf 3 Inf 4 Eff 2 Eff 3 Eff N = NH -N NO -N NO -N NH -N NO -N NO -N       (2.11) removed N N·Q· t  (2.12) Nitrogen removed by denitrification (Ndenitrified) was calculated by the difference between Nremoved and Nassimilated. 2.4.3 Biomass production The growth yield of microorganisms (YObs) in aerobic granules expressed in terms of gram of biomass produced per gram of organic matter removed (biomass production) was calculated for selected operational periods. The amount of biomass produced was calculated according to equation (2.9), and the amount of organic matter removed was calculated from the experimental data obtained from the performance of the reactor in the selected period of time, as the difference between the average COD concentrations in the influent and effluent. Finally, the obtained amount of biomass is divided by the amount of COD removed according to equation (2.13). Materials and methods 123 2   p obs Inf Eff W YCOD COD ·Q· t   (2.13) Where Yobs is the growth yield of aerobic granules (g VSS /g COD), ∆WP the amount of produced biomass (g VSS) CODinf and CODeff the average COD concentration in the influent and effluent (g COD/L), Q: flow rate (L/d) and Δt: length of each period (d). 2.4.4 Nitrogen removal rates in the Anammox Processes Ammonia and nitrite oxidation rates (AOR and NOR, respectively) and nitrogen removal rate by Anammox bacteria (ANR) in g N/L·d, and Ammonia removal efficiency (AR) and Nitrogen removal efficiency (NR) in %, were estimated based on nitrogen balances and the Anammox process stoichiometry, equation (2.14), according to equations: (2.15), (2.16), (2.17), (2.18), (2.19) and (2.20). + 42 3 2 3 2 0.5 0.15 2 NH 1.32 NO 0.066 HCO 0.13 H 1.02 N 0.26 NO 0.066 CH O N 2 H O         (2.14)     4 Inf 2 Inf 3 Inf 4 Eff 2 Eff 3 Eff N = NH -N NO -N NO -N NH -N NO -N NO -N       (2.15)       4 Inf 4 Eff N NH -N NH -N 2.04 AOR = HRT    (2.16)   3 Eff 3 Inf 0.26 · N NO -N NO -N 2.04 NOR = HRT    (2.17) N ANR = HRT  (2.18) 4 Inf 4 Eff 4 Inf NH -N NH -N AR ·100 NH -N     (2.19) 4 Inf 2 Inf 3 Inf N NR ·100 NH -N NO -N NO -N    (2.20) Being HRT the hydraulic retention time (d). Chapter 2 124 2 2.5 REFERENCES Amann, R., Binder, B. J., Olson, R. J., Chisholm, S. W., Devereux, R. and Stahl, D. A. (1990). Combination of 16S ribosomal-RNA-targeted oligonucleotide probes with flow-cytometry for analyzing mixed microbial-populations. Applied and Environmental Microbiology 56(6): 19191925. Amann, R., Ludwig, W. and Schleifer, K. H. (1995). Phylogenetic identification and in-situ detection of individual microbial-cells without cultivation. Microbiological Reviews 59(1): 143-169. Amann, R., Ludwig, W., Schulze, R., Spring, S., Moore, E. and Schleifer, K. H. (1996). rRNAtargeted oligonucleotide probes for the identification of genuine and former pseudomonads. Systematic and Applied Microbiology 19(4): 501-509. APHA-AWWA-WPCF (2005). Standard methods for the examination of water and wastewater. American Public Health Association/American Water Works Association/Water Environment Federation. Washington DC, USA Beun, J. J., van Loosdrecht, M. C. M. and Heijnen, J. J. (2002). Aerobic granulation in a sequencing batch airlift reactor. Water Research 36(3): 702-712. Bower, C. E. and Holm-Hansen, T. (1980). A Salicylate–Hypochlorite Method for Determining Ammonia in Seawater. Canadian Journal of Fisheries and Aquatic Sciences 37(5): 794-798. Daims, H., Bruhl, A., Amann, R., Schleifer, K. H. and Wagner, M. (1999). The domain-specific probe EUB338 is insufficient for the detection of all Bacteria: Development and evaluation of a more comprehensive probe set. Systematic and Applied Microbiology 22(3): 434-444. Daims, H., Nielsen, J. L., Nielsen, P. H., Schleifer, K. H. and Wagner, M. (2001). In situ characterization of Nitrospira-like nitrite oxidizing bacteria active in wastewater treatment plants. Applied and Environmental Microbiology 67(11): 5273-5284. Dapena-Mora, A., Fernandez, I., Campos, J. L., Mosquera-Corral, A., Mendez, R. and Jetten, M. S. M. (2007). Evaluation of activity and inhibition effects on Anammox process by batch tests based on the nitrogen gas production. Enzyme and Microbial Technology 40(4): 859-865. de Kreuk, M., Kishida, N. and van Loosdrecht, M. C. M. (2007). Aerobic granular sludge: state of the art. Water Science and Technology 55 (8-9): 75-81. López-Fiuza, J., Buys, B., Mosquera-Corral, A., Omil, F. and Méndez, R. (2002). Toxic effects exerted on methanogenic, nitrifying and denitrifying bacteria by chemicals used in a milk analysis laboratory. Enzyme and Microbial Technology 31(7): 976-985. Manz, W., Amann, R., Ludwig, W., Wagner, M. and Schleifer, K. H. (1992). Phylogenetic oligodeoxynucleotide probes for the major subclasses of Proteobacteria - problems and solutions. Systematic and Applied Microbiology 15(4): 593-600. Mobarry, B. K., Wagner, M., Urbain, V., Rittmann, B. E. and Stahl, D. A. (1996). Phylogenetic probes for analyzing abundance and spatial organization of nitrifying bacteria. Applied and Environmental Microbiology 62(6): 2156-2162. Mosquera-Corral, A., Gonzalez, F., Campos, J. L. and Mendez, R. (2005a). Partial nitrification in a SHARON reactor in the presence of salts and organic carbon compounds. Process Biochemistry 40(9): 3109-3118. Mosquera-Corral, A., de Kreuk, M. K., Heijnen, J. J. and van Loosdrecht, M. C. (2005b). Effects of oxygen concentration on N-removal in an aerobic granular sludge reactor. Water Res 39(12): 2676-2686. Neef, A., Zaglauer, A., Meier, H., Amann, R., Lemmer, H. and Schleifer, K. H. (1996). Population analysis in a denitrifying sand filter: conventional and in situ identification of Paracoccus spp. in methanol-fed biofilms. Applied and Environmental Microbiology 62(12): 4329-4339. Neef, A., Amann, R., Schlesner, H. and Schleifer, K. H. (1998). Monitoring a widespread bacterial group: in situ detection of planctomycetes with 16S rRNA-targeted probes. Microbiology-Uk 144: 3257-3266. Materials and methods 125 2 Pijuan, M., Guisasola, A., Baeza, J. A., Carrera, J., Casas, C. and Lafuente, J. (2005). Aerobic phosphorus release linked to acetate uptake: Influence of PAO intracellular storage compounds. Biochemical Engineering Journal 26(2-3): 184-190. Schmid, M., Schmitz-Esser, S., Jetten, M. and Wagner, M. (2001). 16S-23S rDNA intergenic spacer and 23S rDNA of anaerobic ammonium-oxidizing bacteria: implications for phylogeny and in situ detection. Environmental Microbiology 3(7): 450-459. Schwarzenbeck, N., Erley, R. and Wilderer, P. A. (2004). Aerobic granular sludge in an SBRsystem treating wastewater rich in particulate matter. Water Science and Technology 49(1112): 41-46. Soto, M., Veiga, M. C., Mendez, R. and Lema, J. M. (1989). Semi-micro C.O.D. determination method for high-salinity wastewater. Environmental Technology Letters 10(5): 541-548. Strous, M., Kuenen, J. G. and Jetten, M. S. M. (1999). Key physiology of anaerobic ammonium oxidation. Applied and Environmental Microbiology 65(7): 3248-3250. Wagner, M., Rath, G., Amann, R., Koops, H. P. and Schleifer, K. H. (1995). In-situ identification of ammonia-oxidizing bacteria. Systematic and Applied Microbiology 18(2): 251-264. Wagner, M., Rath, G., Koops, H. P., Flood, J. and Amann, R. (1996). In situ analysis of nitrifying bacteria in sewage treatment plants. Water Science and Technology 34(1-2): 237-244. Chapter 3 OPERATION OF AN AEROBIC GRANULAR PILOT SCALE SBR PLANT TO TREAT SWINE SLURRY 1 Summary A pilot scale Sequencing Batch Reactor (SBR) was operated in order to remove organic matter and nitrogen from swine slurry characterized by its high variable composition. Aerobic granules successfully developed in the reactor during the first weeks of operation. The physical properties of the biomass remained rather stable during the longterm operation of the reactor (307 days), despite the high variation in the organic and nitrogen applied loading rates (OLR and NLR), which varied from 1.4 to 6.3 kg CODs/m3·d and from 0.5 to 2.5 kg N/m3·d, respectively. Furthermore, the C/N ratio of the feeding also varied in a wide range (1.9-9.4 g CODs/g N). The reactor had a good biomass retention capacity to select for granular biomass. However, its efficiency to retain the solids present in the feeding was low. Consequently, the VSS concentration in the effluent was similar to that in the influent. Aerobic granulation in SBR systems appears as an interesting alternative to treat slurry in small livestock facilities, where the implementation of anaerobic digestion systems is not a feasible option or the removal of nitrogenous compounds is required. 1 Morales N., Figueroa M., Fra-Vázquez A., Val del Río A., Campos J. L., MosqueraCorral A. and Méndez R. (2013). Operation of an aerobic granular pilot scale SBR plant to treat swine slurry. Process Biochemistry 48(8), 1216-1221. Chapter 3 3 128 OUTLINE 3.1 Introduction 129 3.2 Materials and methods 130 3.2.1 Reactor set-up and operational conditions 130 3.2.2 Analytical methods 132 3.2.3 Calculations 133 3.3 Results and discussion 134 3.3.1 Granule formation and physical properties 134 3.3.2 Organic matter and Nitrogen removal 138 3.3.3 Operational cycles 141 3.3.4 Microbial populations 144 3.4 Conclusions 148 3.5 References 149 Operation of an aerobic granular pilot scale SBR plant to treat swine slurry 129 3 3.1 INTRODUCTION Spain is the second European swine livestock producer, after Germany, and the fourth in the world (FAOSTAT, 2010). Pig population in 2011 in Spain was of around 25.6 million heads, which represented the 17% of the total production in the Europe-27 (Figure 3.1), (Eurostat, 2011). These animals produced around 25 million tones of swine slurry per year (Yagüe and Quílez, 2010). Factors such as type of livestock, diet, housing system or waste handling system affect the concentration of nutrients in pig slurry and the daily manure production values, which vary widely from herd to herd. Traditional waste treatment in farms consisted in the use of manure and slurry as nutrients for agriculture. However, the amount of nutrients that can be applied to soils is limited and depends on the soil characteristics and the current nutrient levels. For instance, the maximum load of total nitrogen to be spread by year on lands in Europe should be less than 170 kg of nitrogen per hectare, while nitrate levels in groundwater are set to a maximum of 50 mg NO3–/L or 11.3 mg N-NO3–/L (EEC, 1991). The production of biogas thought an anaerobic digestion of carbonaceous compounds for waste treatment and energy recovery has been encouraged by the European Union through energy policies (European Parliament, 2008). It can reduce volatile organic compounds emissions, control odors, mineralize nutrients, and improve its fertilizing properties and energy recovering. Nevertheless, the anaerobic digestion has the disadvantage of its low efficiency regarding the nitrogen removal, high capital investment, and requirement of specialist technical input and control. In this way, an anaerobic digestion plant with a capacity of 100000 t/year costs around 3 million euro, while it can generate 1.5 million of Nm3/year of biogas with a 65% of methane (Angulo, 2004). Centralized collective treatment facilities for small plants can solve part of the limitations, when the transport costs are affordable (Flotats et al., 2009). Nevertheless, the ammonia removal is still not solved. A valid solution when the extension of available land to absorb the nutrients is limited has to be developed. There is not a unique solution and numerous technologies have been used in order to treat the livestock waste as main treatment or pre/post-treatment, mainly with the objective of nitrogen removal based on nitrification and denitrification processes. Nowadays, Anammox treatment is also an option due to its potential of removing nitrogen from wastewater characterized by a low C/N ratio as anaerobic effluents (Vázquez-Padín et al., 2009) with a reduction of costs compared to other technologies (STOWA, 1996). Despite the promising results of this novel technology (Figueroa et al., 2012), until now, there are not so many applications at full scale. In small and isolated farms, swine waste treatment needs the application of robust, but flexible in operation, systems which can remove the organic matter content and the nutrients. Anaerobic lagoons are another option widely used to treat swine slurry, however, there are environmental and health concerns related to this technology (Vanotti et al., 2009). Chapter 3 3 136 using a synthetic wastewater with an OLR of 1.2-2.4 kg COD/m3·d, as the initial selection of biomass with good settling properties during the start-up period is skipped. A) B) C) D) Figure 3.4. Images showing the aerobic granules evolution in the SBR reactor during Stage II. A) day 41, B) day 99, C) day 120 and D) day 286. In the present work, the settling properties of biomass rapidly improved along the operation period (Figure 3.5). The SVI30 of the inoculum was about 380 mL/g TSS, while the biomass present in the reactor during the first week already had a SVI10 below 80 mL/g TSS. This parameter varied between 27 and 60 mL/g TSS throughout the remaining experimental period. The improvement of the settling properties of the biomass is one of the main characteristics of aerobic granular technology (Arrojo et al., 2004). SVI for aerobic granules is usually below 80 mL/g TSS, while the SVI for floccular biomass is above 120 mL/g TSS (Toh et al., 2003). Previously, the aerobic granules obtained by Figueroa et al. (2011), in a reactor fed with swine slurry, showed SVI10 values between 20 and 75 mL/g TSS, whereas the values of the biomass from the pilot plant SBR operated by Jungles et al. (2011) decreased from 190 to 26 mL/g TSS during the operation period. Also Isanta et al. (2012) reduced the SVI from 200 to 13-16 mL/g TSS in their pilot scale SBR reactor. In fact, Toh et al. (2003) found a relation between the granule size and the SVI values. The smallest granule size had a smallest SVI, and that value increased gradually as the granules grew bigger in size. A similar relation was found in the aerobic granules grew in Operation of an aerobic granular pilot scale SBR plant to treat swine slurry 137 3 this study (Figure 3.5 and Figure 3.6). In this way, during Stage II, average diameter and SVI decreased continuously. Later, until around day 225, the average diameter increased, and an increase in the SVI was measured. However, the SVI gradually recovered the previous values (around 30 mL/g TSS) despite the size of the granules did not decrease in the same way. The high biomass retention is one of the main advantages of aerobic granulation. Solids concentration inside the reactor reached a value up to 2-3 g VSS/L during the startup period (Figure 3.5). 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 0 2 4 6 8 10 12 14 025 50 75 100 125 150 175 200 225 250 275 300 325 VSS influent and effluent (g/L) and Diameter (mm) VSS reactor (g /L) Time (d) S I S II Figure 3.5. Solids concentration in the influent (), in the reactor ( ) and in the effluent () in g VSS/L and average diameter of the granules in mm (). The biodegradability of the aerobic granular biomass produced in the reactor was studied by Val del Río et al. (2011), with and without thermal treatment. These authors concluded that anaerobic biodegradability of aerobic granules, which was around 33%, was similar to that obtained for an activated sludge (30-50%) and demonstrated the feasibility of their anaerobic digestion. The thermal pre-treatment before the anaerobic digestion was proposed as a good option to enhance the biodegradability: between 20% at 60 ºC and 88% at 170 ºC with respect to the untreated sludge (Val del Río et al., 2011). The biomass concentration in the effluent during Stage I was always below 0.25 g VSS/L, indicating a good biomass retention capacity of the reactor, as no granules were observed in the effluent. Jungles et al. (2011) reported a maximum solids concentration in the effluent below 0.20 g VSS/L during the start up of the pilot scale SBR reactor. With the increase of the applied OLR in Stage II, the biomass concentration in the reactor reached values as high as 11-13 g VSS/L. Chapter 3 3 138 During both stages, the solids concentration in the effluent was dependent on its concentration in the influent, since the system was able to retain the biomass that grew in form of aerobic granules, but those solids that enter with the feeding were washed out. In average, the solids concentration in the effluent was around 0.05 and 0.01 g VSS/L higher than in the influent during Stages I and II, respectively. A preor posttreatment of these solids should be integrated in the overall treatment in order to achieve the limit values required for the discharge effluent quality (de Bruin et al., 2004). The effect of particulate or colloidal suspended solids in aerobic granules was studied by Schwarzenbeck et al. (2004) showing that the COD removal efficiency is strongly dependent on the particle size and that it increases with decreasing the particle size. These authors showed that particles bigger than 50 µm were removed at 40% efficiency from a SBR reactor with a Vs min of 3.0-7.5 m/h. Inizan et al. (2005) also observed that a relatively high concentration of suspended solids in the inlet were not removed from the reactor and this fact deteriorated its performance with regard to the total COD removal. However, these authors obtained good removal efficiencies when industrial wastewaters containing biodegradable CODs were treated. The average SRT value calculated in the pilot scale reactor was around 3.8 days in Stage I (Figure 3.6). This value increased in Stage II, matching up with the higher biomass concentration, and reached an average value of 11.7 d. Once the biomass stabilized around 8 g VSS/L (Figure 3.5), from day 175 on, the SRT stabilized to around 5.1 d (Figure 3.6). 0 5 10 15 20 25 30 0 10 20 30 40 50 60 70 80 025 50 75 100 125 150 175 200 225 250 275 300 325 SRT (d) SVI ( mL/g TSS) Time (d) S I S II Figure 3.6. SVI10 of the biomass in mL/g TSS ( ) and SRT of the biomass in days (□). 3.3.2 Organic matter and Nitrogen removal In Stage I, the applied OLR was of 1.91 ± 0.34 kg CODs/m3·d (Figure 3.7) to promote the formation and development of aerobic granules as this value was previously found appropriated when the reactor was fed with a synthetic feeding (Jungles et al., 2011). Then, in Stage II, the OLR was increased and varied in a wide range (minimum 1.75 and Operation of an aerobic granular pilot scale SBR plant to treat swine slurry 139 3 maximum 6.26 kg CODs/m3·d) in order to test the stability performance of the reactor. These modifications in the OLRs applied did not alter the physical properties of the granules, and the performance of the reactor in terms of removal efficiency was in average of about 73 ± 12 % in stage I and 61 ± 18 % in stage II. Average pH in the influent was 7.6, while that parameter in the effluent was 7.2 probably caused by the nitrification process and the low amount of denitrification. Ammonia stripping due to high ammonia concentration of swine slurry and the extended aeration time applied was considered negligible at these pH values. At the end of the experiment, the removal efficiency dropped dramatically due to the increase of the non-biodegradable fraction of COD present in the pig slurry fed to the reactor. A variation of the biodegradable/non-biodegradable COD ratio in the feeding was observed during the operational time, caused by the degradation of the biodegradable COD in the farm and the subsequent increase of the nonbiodegradable fraction. The presence of a high fraction of non-biodegradable organic matter provoked elevated CODs concentrations in the effluent with the consequent decrease of removal efficiency. The nonbiodegradable fraction widely varied, corresponding to up to 80% of the inlet CODs in some periods, higher than those values of 10% reported for anaerobic biodegradability (GonzálezFernández et al., 2008) and for aerobic granular treatment (Figueroa et al., 2011). These authors also found that the removal efficiency and granules properties remained stable when OLRs from 2 to 8 kg CODs/m3·d were applied. Thus, highly variable characteristics of the influent and particularly the presence of a non-biodegradable fraction provoked the maintenance of a good effluent quality in swine wastewater treatments difficult (Shin et al., 2005). 0 1 2 3 4 5 6 7 0 31 66 101 136 171 206 241 276 OLR (kg COD/m 3 ·d) Time (d) S I S II S I S II 304 Figure 3.7. OLR fed to () and removed from () the SBR in kg CODs/m3·d. Chapter 3 3 140 As a consequence of the high variability of the applied OLR and the feeding COD/N ratio, the NLR ranged between 0.3 and 1.4 kg N/m3·d (Figure 3.8). The average ammonia removal efficiency was around 56.4 ± 13.9% in Stage I and 76.6 ± 8.7% in Stage II. Ammonia was mainly oxidized to nitrite, although some amounts of nitrate were also measured in the effluent (Figure 3.9). Isanta et al. (2012) also observed a partial nitrification (80% of nitrification to nitrite) when they operated a pilot scale granular SBR reactor. These authors suggested that granules size can limit the oxygen flux towards the inner part of the granules. Having in mind that the higher the granule size, the lower the specific granular surface where nitrite oxidizing bacteria (NOB) can grow due to the stratification between heterotrophs, ammonia oxidizing bacteria (AOB) and NOB usually observed in biofilms (Terada et al., 2003). As a consequence AOB outcompete NOB due to low oxygen concentrations as suggested by Hanaki et al. (1990) in a suspended-growth reactor. 0.0 0.4 0.8 1.2 1.6 0 31 66 101 136 171 206 241 276 Load Rate (kg N/m 3 ·d) Time (d) S I S II 304 Figure 3.8. NLR in the influent () in kg N/m3·d, and Ammonia (), Nitrite () and Nitrate () loads in the effluent (accumulated values) in kg N/m3·d. Overall nitrogen removal efficiencies reached values up to 30% in Stage I (Figure 3.8), but the average nitrogen removal during that stage was about 16.4 ± 8.3%. In Stage II, the high variability of the feeding composition provoked the variability of the nitrogen removal efficiency that reached values up to 73%; however, it was highly unstable, especially after day 230 of operation. Until day 230 the average nitrogen removal efficiency was 27.4 ± 18.5%. Then, the overall nitrogen removal was almost completely lost, and nitrite accumulated in the reactor. During periods of high biomass growth, the Nremoved was used for assimilation as well as denitrified. In this way, at the beginning of Stage I from days 17 to 45, when the overall N removal was around 19%, slightly higher than the average in Stage I, the solids Operation of an aerobic granular pilot scale SBR plant to treat swine slurry 141 3 concentration increased from 1.15 to 3.06 g VSS/L and the Nassimilated accounted for 42% of the Nremoved. In stage II, from days 80 to 115, with an overall N removal about 27.3% (the average of Stage II), the Nassimilated added up to 16% of the Nremoved. In consequence, the system was not achieving a good nitrogen removal with the operation strategy tested via denitrification or simultaneous nitrification-denitrification. These results contradict previous ones obtained from laboratory scale experiments (Figueroa et al., 2011), where overall nitrogen removal reached values of 70% with denitrification and simultaneous nitrification-denitrification observed in the operational cycles. 0 50 100 150 200 250 300 350 400 025 50 75 100 125 150 175 200 225 250 275 300 325 Nitrogen (mg /L) Time (d) S I S II Figure 3.9. Concentrations of NH4+-N in the influent (), and NH4+-N (), NO2–-N () and NO3–-N ( ) in the effluent. 3.3.3 Operational cycles In order to evaluate the performance of the reactor, cycle measurements were carried out. CODs, PHA and nitrogenous compounds concentrations, pH, and DO values were measured along operational cycles, on different days of operation. The data from day 56 are shown in Figure 3.10 and Figure 3.11. The biodegradable organic matter was easily removed in the first minutes of the aeration phase in all cycles, in no more than 20-25 minutes, during the so called feast period as it can be observed from the obtained concentration profiles. During this time, a reduction in the DO concentration due to the quick organic matter oxidation was measured. Once the biodegradable organic matter was consumed, the DO concentration in the bulk liquid started to increase during the famine period, although it did not reach the saturation level due to the ammonia oxidation and endogenous respiration, while the fraction of nonbiodegradable COD remained unaltered. An increase in CODs was measured at the end of Chapter 3 3 142 each cycle, which can be due to hydrolysis processes of the colloidal particles of the swine slurry. Concentrations of PHA measured in the biomass (Figure 3.10) confirmed that bacteria stored organic matter in the form of these compounds during feast period and consumed them during first minutes of famine period. 8.8 9 9.2 9.4 9.6 9.8 10 0 50 100 150 200 250 300 350 400 020 40 60 80 100 120 140 160 180 DO (mg/L) CODs (mg/L) time (min) Figure 3.10. A) Evolution of CODs (──), DO (····) and PHA (─ ─) concentrations (mg/L) during an operational cycle in the SBR at day 56 (Stage I). The consumption rate for ammonia measured during the cycle in Stage I on day 56 was 212 mg NH4+-N/g VSS·d. In Stage II that value ranged from 44.3 to 213 mg NH4+-N/g VSS·d on days 133 and 295, respectively. Once the organic matter was depleted, nitrate and nitrite concentrations increased in the bulk liquid, but total nitrogen concentration remained almost constant. This observation suggests that simultaneous nitrificationdenitrification processes did not occur during aeration phase, contrary to that observed in other experiments (Figueroa et al., 2011, Jungles et al., 2011). In the experiments performed by these authors the denitrification process using storage compounds can justify the removal of nitrogen during famine phase as suggested by Qin et al. (2005). After the PHA consumption in the approximately first 60 minutes of famine period, PHA values continued rather constant most of the remained time of the cycle, indicating that organic matter was no longer available for denitrification. Operation of an aerobic granular pilot scale SBR plant to treat swine slurry 143 3 0 50 100 150 200 250 020 40 60 80 100 120 140 160 180 Nitrogen (mg/L) time (min) Figure 3.11. Evolution of NH4+-N (──), NO2–-N (····), NO3–-N (─ ─) and TN: NH4+-N + NO2–-N + NO3–-N (──) concentration in mg N/L during an operational cycle in the SBR on day 56 (Stage I). Figueroa et al. (2011) observed simultaneous nitrification-denitrification in their operational cycles, in spite of aeration in the 15 first minutes of the cycle, and once the organic matter was depleted it occurred using pre-accumulated poly-b-hydroxybutyric acid (PHB). Isanta et al. (2012) also observed nitrite accumulation in their SBR cycle studies, as 56% of total nitrogen at the end of the cycle was converted to nitrite and 19% to nitrate. However, these authors observed the complete nitrite and nitrate consumption through denitrification after the static feeding period, which lasted one hour, when readily degradable COD was available and without aeration. The short feeding period used in our study hindered a higher denitrification using the COD added during the feeding in a similar way to that observed by Isanta et al. (2012). On the other hand, the consumption of readily biodegradable organic matter at the beginning of the cycle and the small amount of PHA available in the granule, made simultaneous nitrification-denitrification not possible, as it happened in the reactor operated by Figueroa et al. (2011). The accumulation of non-readily biodegradable CODs can conduct to low available COD/N ratios, which are not suitable for denitrification, even via nitrite, which has a requirement of 2-4 g/g for the COD/N ratio (Mulder, 2003). In this case, a different alternative must be used. The autotrophic ammonia removal by means of Anammox based processes (Vázquez-Padín et al., 2011) can be suitable as partial nitrification was already obtained in the granular SBR reactor. Chapter 3 3 144 3.3.4 Microbial populations General probes were applied to samples collected during the reactor operation to detect the main classes of bacteria involved in the process (EUB338mix, ALF1b and BET42a probes). The hybridized bacteria belonged mainly to β-Proteobacteria subclass, and represented an important fraction in comparison with all the positives from EUB338mix probe. On the other hand, no positive results were obtained for α-Proteobacteria. Positive results were obtained from CFX1223+GNSB941 probes, specific for phylum Chloroflexi (Figure 3.12 A), and SNA probe, which targets Sphaerotilus natans within βProteobacteria (Figure 3.12 B), on samples collected during the first days of operation. Days after, a reduction of the fraction of these filamentous-shape bacteria was observed. Probably these bacteria came with the flocculent activated sludge used as inoculums, and then were washed from the system due to their poor settling properties. Positive results were obtained with NSO190 probe, designed for β-Proteobacteria (AOB), in correspondence with the detection of ammonia oxidation activity in the reactor. Individual cells of AOB were detected when the ammonia oxidation activity started (around day 15, Figure 3.13 A) while AOB cells appeared grouped into clusters when this activity increased (Figure 3.13 B). The Nsm156 probe, specific for Nitrosomonas spp., was applied and positive results can indicate the presence of Nitrosomonas communis and Nitrosomonas oligotropha (Figure 3.14), which are among the most commonly AOB species found in nitrifying WWTPs (Nielsen et al., 2009). The Ntspa712 probe was applied to test the presence of nitrite oxidizing bacteria (NOB) belonging to phylum Nitrospirae. These are the dominant NOB in the majority of nitrifying WWTPs (Daims et al., 2001). However, only a small fraction of bacteria hybridized with this probe (Figure 3.15). These results are in correspondence with the results obtained in the reactor, as ammonia was mainly oxidized to nitrite, and only a small fraction of nitrate was produced. The differences in the relative abundance of each group of bacteria (heterotrophic, AOB and NOB) can be due to the competition for the same subtract, in this case the dissolved oxygen, where NOB were the less favored. Isanta et al. (2012) observed a ratio of the AOB and NOB fractions in the range 5-8 using FISH analysis when obtained partial nitrification in their reactor. From the obtained results an evolution of the bacterial populations present in the inoculum to those observed in the granular biomass occurs. Further work is needed to establish a correlation between the microbial populations identified and the granulation process. Operation of an aerobic granular pilot scale SBR plant to treat swine slurry 145 3 A) B) Figure 3.12. FISH analysis of a biomass sample from day 2 of operation. The bar represents 10 μm. A) Filamentous shape bacteria of phylum Chloroflexi (CFX1223+GNSB941 probes: Cy3, red) and all bacteria (EUB338mix: FITC, green). B) Filamentous bacteria Sphaerotilus natans (SNA probe: Cy3, red) and all bacteria (EUB338mix: FITC, green). Chapter 6 6 248 (diameter <0.5 mm). The difference resides in the fraction of biomass subjected to aerobic conditions (Figure 6.27). It means that more than 65% of the volume of small particles was penetrated by oxygen, but only a 6% for larger granules was (Figure 6.27 A). Taking into account the granules size distribution, 92% of the overall volume of biomass was subjected to anoxic conditions, and 8% to aerobic conditions. This effect is schematized in Figure 6.28 where the aerobic and anoxic fractions of granules with two different sizes are represented. A) 40 45 50 55 60 65 70 75 80 0 10 20 30 40 50 60 70 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 Oxygen depth penetration (μm) Aerobic fraction of the granule (%) Granule Radius (mm) B) 1 10 100 0 5 10 15 20 25 30 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0 5.5 6.0 6.5 7.0 Aerobic % % Volume Diameter (mm) Figure 6.27. A) Aerobic fraction of the granule (blue) in % and oxygen depth penetration (green) in μm, in function of the granule radius (mm) for a DO concentration of 3.8 mg/L and 20 ºC ( ), and for a DO concentration of 1.35 mg/L and 15 ºC ( ). B) Size distribution of the granules during stage A-I in percentage of volume of the total distribution when the DO concentration was 4.6 mg O2/L (day 1 to 242). In each column, (■) represents the anoxic volume and (■) the aerobic volume. The line (─) represents the fraction of the granule penetrated by oxygen (subjected to aerobic conditions). Even if the DO concentration applied in the different stages was selected in order to maintain the oxygen depth penetration and consequently, the aerobic fraction constant, the variations in the size distribution, conducted to variations in the different experimental periods. I.e., in stage A-IV the penetration depth varied in the range 41-45 μm (Figure Dynamics of the CANON process treating low nitrogen load at low temperature 0F 6 249 6.27A). However, the size distribution variation during the stage (Figure 6.29) produced a progressive variation in the aerobic fraction of the biomass, from 6% at the beginning of the period (Figure 6.29 A) to 14% at the end (Figure 6.29 B), when the granules size distribution varied significantly, as bigger granules, in this case larger than 4.5 mm in diameter, were not observed. A) Diameter = 0.5 mm, oxygen depth penetration = 74.1 μm B) Diameter = 3.5 mm, oxygen depth penetration = 67.2 μm Figure 6.28. Schematic representation of the aerobic (■) and anoxic (■) fraction produced by the oxygen depth penetration, for granules with diameter 0.5 mm and 3.5 mm, in the conditions of stage A-I (20 ºC, DO = 3.8 mg O2/L). Thus, even if the oxygen penetration depth remained stable, the biomass fraction subjected to aerobic conditions increased due to the presence of smaller granules. In this way, in period 946-972 d at the beginning of the Stage A-IV, the AOR was 0.28 g N/L·d while the ANR was 0.42 g N/L·d, being the AOR/ANR ratio around 0.67 and the (AORNOR)/ANR ratio 0.60. At the end of the stage, in period 1075-1092 d, the AOR increased to 0.36 g N/L·d while ANR was around 0.38 g N/L·d. Therefore, the AOR/ANR increased to 0.95, while the (AOR-NOR)/ANR ratio was 0.66. The reduction of the Anammox activity can correspond with the loss of bigger granules, which have a larger anoxic fraction, and with the increment of percentage of small granules percentage, which have a larger aerobic fraction. The increase of the NOB activity, which reached 0.11 g N/L·d also reduced the overall nitrogen removal efficiency during this period, and avoid the nitrite accumulation in the system. In section 6.4.3, the effect of NOB development is discussed. This effect is more remarkable in small size granules. The effect of oxygen penetration depth related to the size distribution agreed with results from other works. Nielsen et al. (2005) found that aggregates smaller than 0.5 mm were specialized in aerobic ammonium oxidation, and those larger than 0.5 mm were specialized in anoxic ammonium oxidation. Vlaeminck et al. (2009) obtained aggregates with a size between 0.1 and 1.0 mm (rather floccular) specialized in aerobic ammonium oxidation and granules with 1.8 ± 0.3 mm specialized in anoxic ammonium oxidation. These Chapter 6 6 250 authors corroborated their findings through FISH analyses, which showed that Anammox bacteria were more abundant in the bigger particles while AOB dominated the flocs. Anammox bacteria grow in the more external layers of the anoxic volume of the granules, where NH4+ and NO2– are available, but protected from oxygen. Vázquez-Padín et al. (2010b) observed that Anammox bacteria were located inside the granules in the range of depths between 400 and 1000 μm, when applied DO concentration was of 6.6 mg O2/L (which would imply a wider aerobic layer than the obtained in the granules of the present study). These authors also observed that AOB and Anammox coexisted in the range 400600 μm, as AOB were still present at 600 μm, but their share decreased when the depth increased. 6.4.2.3 Biomass density effect on oxygen depth penetration An important factor in the determination of the oxygen penetration using the equation (6.14) is the biomass density of the granules. For a granule with an average diameter of 3.64 mm (the average diameter in stage A-I), the oxygen depth penetration and anoxic ratio vary as shown in Figure 6.30 as a function of the biomass density in g VSS/Lgranule. Inoculated granules in stage A-I had an average biomass density of around 45 g VSS/Lgranule and this value ranged from 45 to 90 g VSS/Lgranule in the stages at 20 ºC (A-I and A-III). These important variations in the biomass density had a limited impact on the variation of oxygen depth penetration. On the contrary, granules inoculated in stage B-I had a biomass density value of 66 g VSS/Lgranule, and after 170 days of operation under the conditions of stage B-I (DO=0.2 mg O2/L, and 50 mg NH4+-N/L) the density dropped to 23 g VSS/Lgranule. At this low biomass density, the effect over the oxygen depth penetration is more noticeable. Thus, in addition to the increase of the oxygen depth penetration due to the biomass activity decrease caused by the temperature reduction, the diminished in the biomass density also contributed to the increase in the oxygen depth penetration. This effect can be the responsible of the drop of Anammox activity in stage B-I. Dynamics of the CANON process treating low nitrogen load at low temperature 0F 6 251 A) 1 10 100 0 5 10 15 20 25 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0 5.5 6.0 6.5 7.0 Aerobic % % Volume Diameter (mm) B) 1 10 100 0 2 4 6 8 10 12 14 16 18 20 0.0 0.7 1.0 1.5 2.0 2.5 3.0 3.5 4.0 4.5 5.0 5.5 6.0 6.5 7.0 Aerobic % % Volume Diameter (mm) Figure 6.29. Columns represent the size distribution of the granules in percentage of volume of the total distribution. In each column, the red section (■) represents the anoxic volume fraction and the blue section (■) the aerobic volume fraction. The line (─) represents the fraction of the granule penetrated by oxygen (subjected to aerobic conditions). A) At the beginning of stage A-IV (days 946-970, DO=1.6 mg O2/L). B) At the end of stage A-IV (1075-1092, DO=1.5 mg O2/L) Chapter 6 6 252 0 50 100 150 200 250 1.60 1.62 1.64 1.66 1.68 1.70 1.72 1.74 1.76 1.78 1.80 010 20 30 40 50 60 70 80 90 100 110 120 130 140 150 160 Oxygen depth penetration (μm) r 0 (mm) Density (g VSS/L granule ) 20 ºC 20 ºC 15 ºC 15 ºC Figure 6.30. Anoxic radius (─) in mm and oxygen depth penetration (···) in μm as a function of the biomass density of the granule (g VSS/Lgranule), calculated for a granule with a diameter of 3.64 mm in the conditions of stage A-I (20 ºC) and A-II (15 ºC). Granules inoculated in stage B-I were cultivated previously at 24-30 ºC, fed with a solution containing an ammonia concentration of 940 mg NH4+-N/L and operated at a DO concentration of 1.8 mg O2/L. Low substrate concentration can promote the formation of looser microcolonies, with less areal cell density (the number of cells per unit area), which facilitates substrate penetration into the biofilm (Okabe et al., 2004). In the present case both substrates for AOB, ammonia and oxygen, had always low concentrations. 6.4.2.4 Ammonia oxidizing bacteria activity prediction To predict the ammonia oxidizing activity under different operational conditions is fundamental to control the processes involved in the autotrophic nitrogen removal. Since both, temperature and dissolved oxygen concentration regulate ammonia oxidizing bacteria activity, once the temperature of the reactor was decreased from 20 ºC to 15 ºC, also the DO concentration had to be decreased to avoid the deeper oxygen penetration inside the CANON granules. This deeper oxygen penetration within the granules can cause the inhibition of Anammox bacteria. To predict the AOB activity values at the different periods of operation is possible using equations (6.16) and (6.17) taking into account that the ammonia oxidation rate was limited by the internal mass transfer rate and oxygen was the limiting substrate of the process. The values measured during stage A-I were used as reference values for the stages A-II to A-IV (Table 6.5). As new granules from a pilot-scale reactor were inoculated in stages A-V, A-VI and B-I (Vázquez-Padín et al., 2012, Vázquez-Padín et al., 2013), the data from that reactor were used as reference for the aforesaid stages, taking into account its diameter, density and biomass activity values. These data were obtained from Vázquez-Padín (2013). Dynamics of the CANON process treating low nitrogen load at low temperature 0F 6 253 Table 6.5. AOB activities predicted and observed. Stage Period (days) T (ºC) DO (mg/L) Diameter (mm) Density (g VSS/ Lgranule) AOR observed (g N/ g VSS·d) AOR predicted (g N/ g VSS·d) A-I (reference) 20 3.8 3.6 54 0.041 -- A-II Complete stage 15 2.8 3.5 54 0.028 ± 0.008 0.029 (507-537) 3.8 3.5 0.034 ± 0.002 0.033 (552-623) 2.4 3.5 0.027 ± 0.009 0.027 A-III 20 1.9 3.4 70 0.031 ± 0.009 0.028 A-IV Complete stage 15 1.5 2.1 57 0.032 ± 0.011 0.034 (946-1000) 1.4 2.7 0.024 ± 0.007 0.026 (10071070) 1.4 1.9 0.035 ± 0.006 0.036 (Vázquez-Padín, 2013) Reference 29 2.1 3.6 41 0.066 ± 0.018 -- A-V Complete stage 15 1.3 3.4 36 0.029 ± 0.015 0.029 (12701340) 2.1 3.6 38 0.034 ± 0.012 0.034 (13671493) 1.0 3.0 32 0.028 ± 0.015 0.031 (Vázquez-Padín, 2013) Reference 26 0.9 2.3 41 0.035 -- A-VI Complete stage 15 0.3 1.1 41 0.026 ± 0.018 0.026 (Vázquez-Padín, 2013) Reference 23 2.0 2.5 66 0.044 -- B-I Complete stage 15 0.17 2.5 44 0.005 ± 0.001 0.007 (11-137) 0.15 2.4 50 0.006 ± 0.001 0.007 (140-185) 0.05 2.5 23 0.005 ± 0.001 0.006 The predicted values in the different stages were similar to the average AOR values observed in the reactor as it is shown in Table 6.5. The same procedure was applied for sub-periods in the stages at 15 ºC, where the activity had more variations (induced by DO concentration or variations in the average size of the granules). The predicted values in these sub-periods were even more similar to the observed ones. Even when the DO concentration was controlled in order to maintain a constant aerobic/anoxic ratio, the performance of the reactor was altered by the occurrence of the Chapter 6 6 254 nitrite oxidizing process in the equilibrium. The effect of the NOB development on the performance of the CANON system and alternatives to repress its development are discussed in the next section. 6.4.3 Nitrite Oxidizing Bacteria 6.4.3.1 NOB development Especially in stage B-I, a continuous increment in the NOB activity was detected (Figure 6.24), which contributed to the reduction in the overall nitrogen removal in the system, as these bacteria compete with Anammox bacteria for the nitrite and with AOB for the oxygen. In order to evaluate the influence of the different biomass fractions, respirometric AOB and NOB activity tests, and specific Anammox activity tests were performed to samples of the biomass from the reactor. Respirometric tests were performed to separated fractions of the granules and the floccular biomass on stage B-I, when an excessive nitrate production was detected. The measured specific Anammox activity of the overall biomass remained without significant variations during the stage (Figure 6.31). A small reduction in the Anammox activity can be inferred from the floccular sample (Figure 6.31), due to its lower anoxic fraction compared to the granular biomass. However, significant differences were found in the NOB activity tests. The NOB activity measured in the flocculent fraction was remarkably higher than the measured in the granules (Figure 6.32). The NOB activity in the floccular sludge at 15 ºC was around 0.12 g N/g VSS·d, while in the granules it was around 0.02 g N/g VSS·d. This can indicate that NOB are preferably in the flocculent fraction of the biomass. Winkler et al. (2011) determined that a mix of AOB, NOB and Anammox formed the white and the smaller granules observed in their reactor, while larger granules were dominated by AOB and Anammox. Activity tests performed by these authors to biomass of their reactor showed similar ammonium oxidizing capacity for both biomass fractions. On the contrary, higher nitrite oxidizing capacity was measured for biomass which had a larger fraction of smaller white granules. Dynamics of the CANON process treating low nitrogen load at low temperature 0F 6 255 0 0.05 0.1 0.15 0.2 0.25 0.3 10 15 20 25 30 35 Activity (g N/g VSS·d) Temperature (ºC) Figure 6.31. Anammox activity of the overall biomass on days: 1 (●), 69 ( ), 91 () and 167 ( ), and of the flocculent biomass on day 120 (), in g N/g VSS·d at different temperatures in stage B-I. 0.00 0.05 0.10 0.15 0.20 10 15 20 25 30 35 NOB Activity (g N/g VSS· d) Temperature (ºC) Figure 6.32. NOB activity of the overall biomass ( ) on day 105, of the granules (●) and flocculent biomass () on day 120 in g N/g VSS·d, at different temperatures during stage B-I. NOB would be preferentially accumulated in smaller granules due to the higher availability of oxygen compared with larger granules. Gilbert et al. (2013) observed that the NOB activity decreased with increasing aggregate size for floccular biomass. Third et al. (2001) observed that NOB bacteria were present in their CANON biomass, even if the NOB activity test failed in their detection. While high temperature and ammonia concentration in excess were applied to their reactor, the nitrate production was near the stoichiometric of the Anammox process. However, a gradual increase in the NOB activity of the biomass was Chapter 6 6 256 detected by these authors when the reactor was operated under ammonia concentration limitation. Volcke et al. (2012) determined that, in a single granular sludge reactor, NOB are present in the smallest granules (and for the same reason in the floccular fraction), while they are outcompeted by Anammox bacteria in larger granules. Gilbert et al. (2013) concluded that NOB can be inhibited in the system, but no extincted, even if the NOB activity is negligible. Furthermore, the FISH technique shows the presence of these microorganisms. 6.4.3.2 NOB repression The continuous repression of the NOB bacteria appears as a key factor in order to achieve high nitrogen removal efficiency in CANON systems operated at low temperatures and ammonia loads. NOB activity in stage B-I was partially reduced after the increasing of the effluent withdrawal (increase of the settling velocity from 0.1 to 0.3 m/h), which mainly removed the floccular biomass from the system, and with the substitution of floccular biomass by re-inoculated granules. However, this strategy needs the availability of CANON granules developed in another reactor treating high loaded nitrogen streams. Bin et al. (2011) observed in nitrifying granules that oxygen penetration was not restricted in the granules of less than 0.6 mm particle diameter. However, in the larger granules (d > 0.9 mm), a smaller aerobic volume fraction and inhibition of NOB growth can be achieved. According to these authors, in small granules, the location of the boundaries between AOB and NOB biomass species were distinct due to the limited existing space provided by granules for the growth of both microorganisms. The research conducted by Volcke et al. (2010) suggested that NOBs are favored to grow in small granules because of their high aerobic fraction. According to these authors, completely autotrophic nitrogen removal by the granular sludge can be achieved while NOB are outcompeted by Anammox bacteria, by controlling the DO concentration in the bulk liquid in a determined range. This optimal range of bulk DO concentration is broader for larger particles and for increasing influent ammonium concentrations. Consequently, systems with big granules are easier to control and more robust towards disturbances in bulk oxygen concentration than those with small ones. However, the granules size also has an effect over AOB, as activity and abundance of these bacteria decrease with increasing the aggregate size (Vlaeminck et al., 2010). Different strategies have been proposed by different authors with the objective of outselecting the NOB and/or favor the AOB and Anammox selection. At high temperatures and high ammonia loads this objective has been successfully achieved by the use of these mechanisms: the AOB growth rate is larger than the NOB one at high temperatures; the inhibition by FA and/or FNA is larger for NOB than for AOB (Anthonisen et al., 1976); DO affinity is larger for AOB than NOB, at least at high temperatures (Blackburne et al., 2008). Nevertheless, batch oxygen half saturation (KO2) test performed more recently by Akintayo Dynamics of the CANON process treating low nitrogen load at low temperature 0F 6 257 (2012) suggested that NOB had a higher affinity for oxygen at lower DO than AOB at 15 ºC. Wett et al. (2013) observed that the KO2 of NOB decreased during the operation of their plant when the DO concentration was between 0.06 and 0.3 mg/L, while that value for the AOB remained stable. Then, when operating a reactor at low temperatures and nitrogen concentrations other selection mechanisms should be explored. In this way, Winkler et al. (2013) concluded that for CANON granular sludge, mainly the diameter of the granular biomass and not the density differences must be used for the segregation of the biomass. Winkler et al. (2011) proposed the feeding of nitrite and ammonium during a CANON start up process at mesophilic temperatures in order to promote the growth of bigger granules, which are dominated by Anammox, and to favor the biomass segregation. Biomass selection based only on the density of the cells of the microorganisms is not viable, as the density of the NOB is higher than that of the Anammox bacteria (Winkler et al., 2013). These authors stated that larger granules can achieve greater settling velocities, and consequently, smaller granules can accumulate in the top of the reactor and will be easily removed. These smaller granules are enriched in NOB. A strategy based on the larger density of the granules was followed by Wett et al. (2013) in the Strass WWTP, where Anammox granules produced from sludge liquor treatment were seeded to the mainstream. Then, a hydrocyclone classifier selected for the high-density sludge fraction in order to retain these Anammox enriched granules. These authors also observed an excessive NOB activity in an Anammox process applied to the water line that conducted to an accumulation of nitrate and to a reduction in the overall nitrogen removal. They did not achieve a NOB repression as long a low DO concentration was applied (around 0.1 mg O2/L). These authors suggested the application of rapid transitions from high to low DO levels in order to avoid NOB development. This strategy is based on the lag-time of the NOB behind the AOB when the system alternates between anoxic/anaerobic and aerobic conditions (Yoo et al., 1999). This lag time was suggested by Turk and Mavinic (1986), however, these authors talked about several hours for the duration of such lag time. Yao et al. (2013) proposed the addition of hydrazine (N2H4), a fundamental intermediate in the biochemical pathway of Anammox bacteria (Kartal et al., 2011), to inhibit the NOB activity and enhance the performance of a CANON SBR reactor operated at 31 ºC. The hydrazine addition reduced at the same time the NO3– production below the stoichiometric ratio. According to these authors, the electrons released from the oxidation of additional N2H4, which substituted the electrons from NO2– oxidation to NO3–, replenished the consumption of Anammox anabolism. The N-N bond in hydrazine is catalyzed in the Anammox metabolism by the hydrazine synthase, which low activity possibly explains the slow growth rates and long doubling times of the Anammox bacteria (Kartal et al., 2012). However, the applicability of these alternative is complicated, as in the work of Yao et al. (2013), hydrazine in a concentration of around 7.5% of the ammonia influent concentration Chapter 6 6 264 Yoo, H., Ahn, K.-H., Lee, H.-J., Lee, K.-H., Kwak, Y.-J. and Song, K.-G. (1999). Nitrogen removal from synthetic wastewater by simultaneous nitrification and denitrification (SND) via nitrite in an intermittently-aerated reactor. Water Research 33(1): 145-154. CONCLUSIONES GENERALES La presente tesis se centra en el estudio de algunas de las nuevas tecnologías desarrolladas en los últimos años en el ámbito del tratamiento de aguas residuales. Actualmente se tienen en cuenta nuevos conceptos para afrontar las nuevas limitaciones ambientales, económicas y sociales, al enfrentarse al crecimiento de la población, al cambio climático, a la escasez de agua, etc. Los nuevos diseños deben ser realizados usando la sostenibilidad como factor de referencia, ya que, los sistemas empleados en la actualidad, están basados en un consumo intensivo de energía y en la utilización de productos químicos, lo que los convierte en sistemas totalmente insostenibles. La aplicación de las tecnologías estudiadas en los capítulos de esta tesis puede dar lugar a sistemas de tratamiento de aguas residuales más eficientes y sostenibles. De este modo, los requerimientos de terreno de las Estaciones Depuradoras de Aguas Residuales (EDAR) se pueden reducir gracias a la aplicación de la tecnología de granulación aerobia a sistemas que operen tanto en discontinuo como en continuo; la recuperación de nutrientes de las aguas residuales puede lograrse mediante el tratamiento separado de orina y por stripping de amonio; y la eficiencia energética de las EDAR puede ser mejorada si los procesos Anammox se aplican directamente a la línea de agua en las EDAR. Las principales conclusiones de cada capítulo de esta tesis se detallan a continuación: Capítulo 3: Operación de una planta de granulación aerobia a escala piloto SBR para el tratamiento de purines porcinos. En este capítulo se estudió la tecnología basada en biomasa granular aerobia para su aplicación como sistema de eliminación de nitrógeno y materia orgánica en pequeñas granjas porcinas, usando un reactor secuencial (secuencing batch reactor, SBR) a escala de planta piloto. El reactor presenta una alta relación entre su altura y diámetro, y una alta capacidad de retención de biomasa, lo cual permite reducir la superficie requerida para su instalación. Conclusiones generales 266 Después de 9 días de operación se formaron los primeros gránulos aerobios en el reactor. Dichos gránulos además mantuvieron sus propiedades a pesar de las variaciones en las características de la alimentación. De este modo, el diámetro medio de los gránulos se estabilizó alrededor de 3 mm, con un índice volumétrico de lodos (IVL10) menor a 50 mL/g SST. A pesar de que el reactor mostró una buena capacidad de retención de la biomasa formada en el reactor, los sólidos presentes en el purín de cerdo no eran bien retenidos en el sistema. Por tanto, se requiere la integración de un sistema de preo post-tratamiento de dichos sólidos en el tratamiento global para lograr cumplir los valores requeridos para la descarga del efluente. La concentración de sólidos en el reactor varió entre los 5 y los 12 g SSV/L. Esta concentración de biomasa es más alta que el valor estándar medido en los reactores de lodos activos. El hecho de obtener una alta concentración de biomasa en el reactor biológico permite tratar una alta carga, lo que permite reducir el volumen total del reactor. Los gránulos aerobios soportaron adecuadamente las variaciones producidas en la velocidad de carga orgánica (VCO), en la velocidad de carga nitrogenada (VCN), y en la relación entre la demanda química de oxígeno (DQO) y el N de la alimentación. Estas variaciones se suelen encontrar en los efluentes industriales. La eficiencia de eliminación de materia orgánica no se vio afectada por las fluctuaciones en la VCO aplicada, sino por la fracción no biodegradable de la materia orgánica presente en los purines porcinos. La carga de amonio se oxidó principalmente a nitrito, la eficacia de eliminación de amonio fue de aproximadamente un 76%. Sin embargo prácticamente no se observó denitrificación durante la experimentación, tal como revelan los análisis de los ciclos de operación. La evolución de la población microbiana se siguió aplicando análisis de Hibridación in situ con Fluorescencia (FISH). Las poblaciones de bacterias identificadas evidenciaron una evolución desde el lodo inoculado hasta las poblaciones que forman los gránulos aerobios. En el inóculo se detectaron principalmente organismos filamentosos, que posteriormente fueron eliminados del sistema al tiempo que se desarrollaban los gránulos. Las poblaciones microbianas nitrificantes estaban compuestas fundamentalmente por miembros de Nitrosomonas spp. como bacterias oxidantes de amonio, y de una menor cantidad de bacterias oxidantes de nitrito pertenecientes al filo Nitrospirae, en correspondencia con la acumulación de nitrito que se observó en el sistema. Capítulo 4: Desarrollo de biomasa de tipo aerobia granular en un reactor continuo agitado. Se operó un reactor continuo de tanque agitado (RCTA) con el objetivo de definir las condiciones de operación más adecuadas para la obtención del crecimiento de biomasa aerobia en forma de gránulos. Los agregados microbianos se formaron en un reactor con una geometría similar a la de los reactores de lodos activos usados en las EDAR, que normalmente tienen una relación entre su altura y su diámetro de entorno a 1. Los gránulos Conclusiones generales 267 aerobios se obtienen generalmente en reactores secuenciales SBR, donde la selección de biomasa se logra gracias a los ciclos de operación. Sin embargo, en el reactor en continuo operado en este capítulo, la selección se logró en el tubo de salida del efluente, donde se controló la velocidad ascensional del agua. Con este diseño, las partículas que tenían una velocidad de sedimentación menor que la velocidad ascensional fijada en el tubo de salida eran lavadas del reactor, mientras que la biomasa con buenas propiedades de sedimentación eran retenidas en el sistema. La formación de gránulos aerobios en el reactor operado en continuo se logró cuando se aplicó un tiempo de retención hidráulico (TRH) de 1 hora, y una velocidad ascensional límite en el tubo de descarga del efluente de 10 m/h. Los valores de diámetro medio y de velocidad de sedimentación de los gránulos que se obtuvieron en el reactor RCTA fueron similares a los que presentan los gránulos formados en reactores secuenciales. De este modo, el diámetro medio de los gránulos fue de 6,8 mm y la velocidad de sedimentación varió en un rango de 36-48 m/h. Sin embargo, el valor de IVL fue de unos 127 mL/g SST, menos favorable que el que presentan los gránulos formados en reactores SBR. De todos modos, el valor de IVL de los gránulos formados en continuo sigue siendo mejor que el que se obtiene en los reactores convencionales de lodos activos. Cuando el TRH se fijó en 6 y 3 horas, se desarrollaron grandes cantidades de bacterias filamentosas en el reactor, y no se logró la formación de gránulos aerobios. Por otra parte, cuando el TRH aplicado fue de 1 hora, las bacterias tenían fundamentalmente forma de bacilos, mientras las bacterias de forma filamentosa estaban ausentes. Transformar un sistema en continuo para operar en las condiciones de un SBR para obtener gránulos aerobios puede ser dificultoso. Por tanto, la posibilidad de formar gránulos aerobios en un reactor operado en continuo abre una nueva perspectiva a la aplicación de la tecnología granular aerobia para la mejora de EDAR ya en operación. Capítulo 5: Recuperación de N y P de orina mediante precipitación de estruvita, seguido de un tratamiento combinado con sobrenadante de un digestor anaerobio mediante un sistema de stripping de amonio a escala industrial. En este capítulo, se utilizó un reactor de stripping a escala real equipado con un prestripper de CO2 para recuperar amonio de una corriente rica en nitrógeno, y de este modo producir un fertilizante, el sulfato de amonio. En el mismo estudio, se empleó un sistema a escala de planta piloto para recuperar fósforo de orina recogida separadamente. Se añadió óxido de magnesio al tanque de orina, con el objetivo de eliminar más del 95% del fósforo disuelto en la orina, mediante la precipitación y posterior separación de estruvita. Los cristales de estruvita obtenidos en el reactor tenían tamaños medios de entre 42 y 80 µm. Conclusiones generales 268 A pesar de esto, el sistema de recuperación de sólidos durante el pre-tratamiento de la orina necesita ser mejorado para alcanzar una mayor eficacia de recuperación de nutrientes. Además, el sistema de recogida separada de orina debe ser conectado al sistema de stripping, de modo que se puedan combinar ambos procesos y optimizar el sistema de recuperación de nutrientes. Los resultados preliminares obtenidos del reactor a escala real de stripping de amonio, tratando el sobrenadante de un digestor anaerobio conjuntamente con la orina recogida separadamente, muestran la viabilidad de este tratamiento combinado para la recuperación de nitrógeno y fósforo de las aguas residuales. La adición de un 10% en volumen de orina tratada al sobrenadante alimentado al sistema de tratamiento por stripping, produce un incremento del 40% en la concentración de amonio. A pesar de que las eficiencias obtenidas en esta unidad fueron menores que las alcanzadas durante las simulaciones, se alcanzó un incremento en la producción de sulfato amónico del 36 y 56% durante los dos experimentos llevados a cabo a escala real en una EDAR municipal. Además, los operadores de la EDAR no informaron de incidencias relacionadas con la adición de orina tratada al sistema de stripping. Sin embargo, la duración de los experimentos fue demasiado corta como para descartar esta posibilidad completamente. El tratamiento de corrientes con altas concentraciones de nitrógeno, como la orina recogida separadamente y el sobrenadante de un digestor anaerobio, reduce los requerimientos de los procesos biológicos de nitrificación/desnitrificación en el reactor de lodos activos. De este modo, se logran ahorros en cuanto a energía de aireación, y se evita el uso de materia orgánica para el proceso de desnitrificación heterótrofa. Capítulo 6: Dinámica de un proceso CANON tratando bajas cargas de nitrógeno a baja temperatura. Se operó un sistema de eliminación autótrofa de nitrógeno basado en el proceso CANON (Completely Autotrophic Nitrogen Removal Over Nitrite) a 15 y 20 ºC alimentado con concentraciones moderadas y bajas de nitrógeno, para estudiar la viabilidad de la aplicación de procesos basados en la oxidación anaerobia de amonio (Anammox) a la línea de agua de EDAR municipales. La aplicación de dichos procesos abre la posibilidad de eliminar los compuestos nitrogenados con menores requerimientos de energía y reactivos químicos. Además, mayores cantidades de materia orgánica estarían disponibles para la producción de metano, ya que la materia orgánica no sería necesaria para el proceso de desnitrificación heterótrofa. El sistema CANON operó en condiciones estables cuando la temperatura era de 20 ºC y la carga de amonio moderada. Bajo estas condiciones operacionales se obtuvo un crecimiento neto de la biomasa, y su concentración alcanzó los 12 g SSV/L. El sistema fue incluso capaz de recuperar la pérdida de concentración de biomasa observada al operar a 15 ºC una vez que la temperatura de operación se devolvió a 20 ºC. Las eficiencias de Conclusiones generales 269 eliminación de amonio y nitrógeno alcanzadas operando a 20ºC fueron de alrededor del 80 y 60%, respectivamente. Cuando el reactor se operó a 15 ºC se observó una pérdida progresiva de la concentración de biomasa, especialmente cuando la concentración de amonio simulaba aquella que se puede obtener en la línea de agua de una EDAR. Consecuentemente se detectó una pérdida en la actividad de eliminación de nitrógeno y una reducción en la eficiencia de eliminación. La pérdida de biomasa observada obligó a usar tiempos de sedimentación más altos en el reactor SBR, con el objetivo de alcanzar mejores tasas de retención de biomasa. Sin embargo, con estas condiciones, se observó el desarrollo de las bacterias oxidantes de nitrito (BON), la acumulación de partículas floculentas y la reducción en la densidad de biomasa de los gránulos. A la hora de controlar el desempeño de las bacterias oxidantes de amonio (BOA) y de las Anammox en un proceso CANON que opere a baja temperatura tratando corrientes con bajas concentraciones de nitrógeno se deben tener en cuenta tres factores cruciales: 1) alcanzar una alta retención de biomasa; 2) obtener un equilibrio entre las actividades de las bacterias BOA y de las bacterias Anammox; y 3) evitar el desarrollo de bacterias BON en la biomasa. Cuando la temperatura de operación se reduce, será necesario mantener bajas concentraciones de oxígeno disuelto, ajustadas a los cambios en la distribución de tamaños y de densidad de biomasa de los gránulos. Estas bajas concentraciones de oxígeno evitan una excesiva penetración del oxígeno en el interior de los gránulos, la cual facilitaría el desarrollo de las bacterias BON y la inhibición de las bacterias Anammox. Las medidas de actividad de las bacterias BON fueron más altas en las partículas pequeñas que en los gránulos más grandes. Se sugiere el uso de re-inoculaciones periódicas, así como el control de la fracción floculenta de la biomasa, para evitar el excesivo desarrollo de BON y para mantener la estabilidad del sistema CANON. CONCLUSIÓNS XERAIS A presente tese céntrase no estudo dalgunhas das novas tecnoloxías desenvolvidas nos últimos ano no eido do tratamento de augas residuais. Actualmente, téñense en conta novos conceptos para afrontar as novas limitacións ambientais, económicas e sociais, ó enfrontarse ó crecemento da poboación, ó cambio climático, á escaseza de auga, etc. Os novos deseños deberanse realizar usando a sostibilidade como factor de referencia, xa que os sistemas empregados na actualidade están baseados no uso intensivo de enerxía e son dependentes do uso de produtos químicos, o que os converte en sistemas totalmente insostibles. A aplicación das tecnoloxías estudadas nos capítulos da presente tese poden dar lugar a sistemas de tratamento de augas residuais máis eficientes e sostibles. Deste xeito, os requirimentos de terreo das Estacións Depuradoras de Augas Residuais (EDAR) pódense reducir grazas á aplicación da tecnoloxía de granulación aerobia a sistemas que operen tanto en descontinuo como en continuo; a recuperación de nutrientes das augas residuais pódese acadar mediante o tratamento separado de urina e por stripping de amonio; a eficiencia enerxética das EDAR pode ser mellorada se os procesos Anammox se aplican directamente á liña de auga nas EDAR. As principais conclusións de cada capítulo desta tese descríbense a continuación: Capítulo 3: Operación dunha planta de granulación aerobia SBR a escala piloto para o tratamento de xurros porcinos. Neste capítulo estudouse a tecnoloxía baseada na biomasa granular aerobia para a súa aplicación como sistema de eliminación de nitróxeno e materia orgánica en pequenas granxas porcinas, empregando un reactor secuencial (secuencing batch reactor, SBR) a escala de planta piloto. O reactor presenta unha alta relación entre a súa altura e o seu diámetro, e unha alta capacidade de retención de biomasa, o cal permite reducir a superficie requirida para a súa instalación. Conclusións xerais 272 Despois de 9 días de operación formáronse os primeiros gránulos aerobios no reactor. Os devanditos gránulos ademais mantiveron as súas propiedades a pesar das variacións nas características da alimentación. Deste xeito, o diámetro medio dos gránulos estabilizouse ó redor de 3 mm, cun índice volumétrico de lamas (IVL10) menor a 50 mL/g SST. A pesar de que o reactor mostrou unha boa capacidade de retención da biomasa formada no reactor, os sólidos presentes nos xurros de porco non eran ben retidos no sistema. Polo tanto, requírese a integración dun sistema de preou posttratamento destes sólidos no tratamento global para lograr cumprir os valores requiridos para a descarga do efluente. A concentración de sólidos no reactor variou entre os 5 e os 12 g SSV/L. Esta concentración de biomasa é máis alta que o valor estándar medido nos reactores de lamas activas. O feito de obter unha alta concentración de biomasa no reactor biolóxico permite tratar unha alta carga, o que permite diminuír o volume total do reactor. Os gránulos aerobios soportaron de xeito adecuado as variacións producidas na velocidade de carga orgánica (VCO), na velocidade de carga nitroxenada (VCN), e na relación entre a demanda química de osíxeno (DQO) e o N da alimentación. Estas variacións adoitan atoparse nos efluentes industriais. A eficiencia de eliminación de materia orgánica non se viu afectada polas flutuacións da VCO aplicada, senón pola fracción non biodegradable da materia orgánica presente nos xurros porcinos. A carga de amonio oxidouse principalmente a nitrito, e a eficacia de eliminación de amonio foi de aproximadamente un 76%. Sen embargo, practicamente non se observou denitrificación durante a experimentación, tal como revelan as análises dos ciclos de operación. A evolución da poboación microbiana seguiuse mediante a aplicación de análises de Hibridación in situ con Fluorescencia (FISH). As poboacións de bacterias identificadas evidenciaron unha evolución dende o lodo inoculado ata as poboacións que forman os gránulos aerobios. No inóculo detectáronse principalmente organismos filamentosos, que posteriormente foron eliminados do sistema ó tempo que se desenvolvían os gránulos. As poboacións microbianas nitrificantes estaban compostas fundamentalmente por membros de Nitrosomonas spp. como bacterias oxidantes de amonio, e dunha menor cantidade de bacterias oxidantes de nitrito pertencentes ó filo Nitrospirae, en correspondencia coa acumulación de nitrito que se observou no sistema. Capítulo 4: Desenvolvemento de biomasa de tipo aerobia granular nun reactor continuo axitado. Operouse un reactor continuo de tanque axitado (RCTA) co obxectivo de definir as condicións de operación máis axeitadas para a obtención do crecemento de biomasa aerobia en forma de gránulos. Os agregados microbianos formáronse nun reactor cunha xeometría similar á dos reactores de lamas activas empregados nas EDAR, que normalmente teñen unha relación entre a súa altura e o seu diámetro de arredor de 1. Os gránulos aerobios obtéñense xeralmente en reactores secuenciais SBR, onde se acada a selección da biomasa Conclusións xerais 273 grazas ós ciclos de operación. Non obstante, no reactor en continuo operado neste capítulo, a selección acadouse no tubo de saída do efluente, onde se controlou a velocidade ascensional da auga. Con este deseño, as partículas que tiñan unha velocidade de sedimentación menor que a velocidade ascensional fixada no tubo de saída eran lavadas do reactor, mentres que a biomasa con boas propiedades de sedimentación era retida no sistema. A formación de gránulos aerobios no reactor operado en continuo acadouse cando se aplicou un tempo de retención hidráulico (TRH) de 1 hora, e unha velocidade ascensional límite no tubo de descarga do efluente duns 10 m/h. Os valores de diámetro medio e de velocidade de sedimentación dos gránulos obtidos no reactor RCTA foron similares ós que presentan os gránulos formados en reactores secuenciais. Deste xeito, o diámetro medio dos gránulos foi de 6,8 mm e a velocidade de sedimentación variou nun rango de 36-48 m/h. Non obstante, o valor de IVL foi duns 127 mL/g SST, menos favorable có que presentan os gránulos formados en reactores SBR. De todas as maneiras, o valor de IVL dos gránulos formados en continuo segue sendo mellor có que se obtén nos reactores convencionais de lamas activas. Cando o TRH se fixou en 6 e 3 horas, formáronse grandes cantidades de bacterias filamentosas no reactor, e non se acadou a formación de gránulos aerobios. Por outra parte, cando o TRH aplicado foi de 1 hora, as bacterias tiñan fundamentalmente forma de bacilos, mentres que as bacterias de forma filamentosa estaban ausentes. Transformar un sistema en continuo para operar nas condicións dun SBR para obter gránulos aerobios pode ser dificultoso. Polo tanto, a posibilidade de formar gránulos aerobios nun reactor operado en continuo abre unha nova perspectiva á aplicación da tecnoloxía granular aerobia para a mellora de EDAR xa en operación. Capítulo 5: Recuperación de N e P de urina mediante precipitación de estruvita, seguido dun tratamento combinado con sobrenadante dun dixestor anaerobio mediante un sistema de stripping de amonio a escala industrial. Neste capítulo utilizouse un reactor de stripping a escala real equipado cun pre-stripper de CO2 para recuperar amonio dunha corrente rica en nitróxeno e, deste xeito, producir un fertilizante, o sulfato de amonio. No mesmo estudo, empregouse un sistema a escala de planta piloto para recuperar fósforo de urina recollida separadamente. Engadiuse óxido de magnesio ó tanque de urina, co obxectivo de eliminar máis do 95% do fósforo disolto na urina, mediante a precipitación e posterior separación de estruvita. Os cristais de estruvita obtidos no reactor tiñan tamaños medios de entre 42 e 80 µm. A pesar disto, o sistema de recuperación de sólidos durante o pre-tratamento da urina precisa ser mellorado para alcanzar unha maior eficacia de recuperación de nutrientes. General conclusions 280 Chapter 6. Dynamics of CANON process treating low nitrogen load at low temperature. A Completely Autotrophic Nitrogen Removal Over Nitrite (CANON) reactor was operated at 15 and 20 ºC and fed with moderated and low concentrations of ammonia in order to test the viability of the application of the ANaerobic AMMonium OXidation (Anammox) based processes to the main line of urban WWTPs. The application of such processes opens the possibility of removing the nitrogenous compounds with less energy and chemical requirements. Furthermore, larger amounts of organic matter would be available for methane production, as this organic matter is not necessary for the heterotrophic denitrification process. The CANON system operated under stable conditions at 20 ºC treating moderate ammonia loads. Under these operational conditions there was a net growth of the biomass and its concentration in the reactor increased up to 12 g VSS/L. Even more, the system was able to recover the biomass concentration loss observed at 15 ºC when the temperature was increased to 20ºC. Ammonia and nitrogen removal efficiencies of around 80 and 60% respectively, were achieved. A progressive biomass loss was observed when the reactor was operated at 15 ºC, especially when the ammonia concentration simulated that of the water line in a WWTP and a consequent loss in the nitrogen removal activity and reduction of the removal efficiency were detected. This biomass loss forced to use a high settling time in the SBR, in order to achieve higher biomass retention. With these conditions, the development of nitrite oxidizing bacteria (NOB), the accumulation of flocculent particles and the loss of biomass density in the granules were observed. Three main factors should be pointed out as crucial to control the performance of the AOB and Anammox bacteria in a CANON system operated at low temperature and nitrogen load: 1) to achieve a high biomass retention; 2) to achieve an equilibrium between the AOB and Anammox activities and 3) to avoid the NOB development in the biomass. Low Dissolved Oxygen (DO) concentrations, adjusted to the granule size distribution and density of the granules variation, were needed to maintain when the temperature decrease. These low DO concentrations avoid an excessive oxygen depth penetration in the granules, which facilitate the development of NOB and inhibition of Anammox. NOB activity measured was larger for small particles than for large granules. The use of periodical re inoculations and the control of the floccular biomass fraction are suggested to avoid the excessive development of NOB and maintain the stability of the CANON system. LIST OF ACRONYMS AND SYMBOLS Abs Absorbance AD Anaerobic Digester Anammox Anaerobic Ammonium Oxidation ANR Nitrogen Removal Rate by Anammox bacteria g N/L·d, kg N/m3·d AOB Ammonia Oxidizing Bacteria AOR Ammonium Oxidation Rate g N/L·d, kg N/m3·d AR Ammonia Removal efficiency % ATES Aquifer Thermal Energy Storage AUFB Aerobic Upflow Fluidized Bed reactor AUSB Aerobic Upflow Sludge Blanket reactor b Endogenous respiration rate d-1 BAF Biological Aerated Filters BAS Biofilm Airlift Suspension Reactor C Carbon CANON Completely Autotrophic Nitrogen Removal Over Nitrite cap per capita CAS Conventional Activated Sludge ce cattle equivalent CGSFDMBR Continuous-flow bioreactor with aerobic granular sludge and self-forming dynamic membrane CHF Swiss Franc COD Chemical Oxygen Demand mg/L, g/L COD/N Chemical Oxygen Demand to Nitrogen ratio C s Concentration of substance S g/L, mg/L CSTR Continuous Stirring Tank Reactor List of acronyms and symbols 282 CULTAN Controlled Uptake Long Term Ammonium Nutrition fertilization CW Constructed Wetlands Cy3 Cyanine 3 Cy5 Cyanine 5 D Diameter mm, cm DAPI 4',6‐DiAmidino‐2‐Phenylindole DEMON Deammonification DO Dissolved Oxygen concentration mg O 2 /L D s Diffusivity coefficient of substance S m2/d EAWAG Swiss Federal Institute of Aquatic Science and Technology EBPR Enhanced Biological Phosphorus Removal EEA European Environment Agency EEC European Community EGSB Expanded Granular Sludge Blanket reactor EPA United States Environmental Protection Agency EPS Extracellular Polymeric Substances (exopolysaccharides) EU European Union F Formamide f b Biodegradable fraction of the biomass F/M Food to Microorganism ratio FA Free Ammonia g N/L FAS Ferrous Ammonium Sulphate FISH Fluorescent in situ hybridization FITC Fluorescein IsoThioCyanate FNA Free Nitrous Acid g N/L FOG Fats, Oil and Grease GC Gas Chromatography GHG Greenhouse Gas GSBR Granular Sequencing Batch Reactor H/D Height to Diameter ratio H c Henry’s law constant for compound c HR Heterotrophic denitrification Rate g COD/L·d List of acronyms and symbols 283 HRT Hydraulic Retention Time d, h IC Internal Circulating reactor IC Inorganic Carbon g/L IFAS Integrated Fixed Film Activated Sludge IN Inorganic Nitrogen IWA International Water Association K s Half saturation constant of substance S g/L MAP Magnesium Ammonium Phosphate MBBR Moving Bed Biofilm Reactor MBR Membrane Biological Reactor MDG Millennium Development Goal MFC Microbial Fuel Cells N Nitrogen NASA Nitrifying Activated Sludge Airlift reactor ND Not detected NLR Nitrogen Loading Rate g N/L·d, kg N/m3·d NOB Nitrite Oxidizing Bacteria NOR Nitrite Oxidation Rate g N/L·d, kg N/m3·d NR Nitrogen Removal NRR Nitrogen Removal Rate g N/L·d, kg N/m3·d OLAND Oxygen-Limited Autotrophic NitrificationDenitrification OLR Organic Loading Rate g COD/L·d, kg N/m3·d P Phosphorus p.e. Population equivalent PBS Phosphate Buffer Solution PH2MV Poly-Hydroxy-2-MethylValerate PHA Poly‐Hydroxy‐Alkanoates PHB Poly-Hydroxy-Butyrate PHV Poly-Hydroxy-Valerate PLC Programmable Logic Controller Q Flow rate L/d, Nm3/h R Ideal gas coefficient RBC Rotating Biofilm Contactor reactor List of acronyms and symbols 284 rNH 4 + Consumption rate for ammonia mg N/L·h rNO x f Consumption rate for nitrogen oxides in the feast phase mg N/L·h rNO X h Production rate of nitrogen oxides during famine phase mg N/L·h rRNA Ribosomal Ribonucleic Acid S Stripping factor SAA Specific Anammox Activity g N/g VSS·d SBR Sequencing Batch Reactor SCADA Supervisory Control And Data Acquisition SEM Scanning Electron Microscope SHARON Single reactor system for High‐activity Ammonia Removal over Nitrite SRT Solids Retention Time d SVI n Sludge Volume Index, after n minutes of settling mL/g TSS T Temperature ºC, K t Time TC Total Carbon g/L TF Trickling or Percolating Filter TKN Total Kjeldahl Nitrogen g N/L TN Total Nitrogen g N/L TOC Total Organic Carbon g/L Tris Tris(hydroxymethyl)aminomethane TSS Total Suspended Solids g/L U.N. United Nations UASB Upflow Anaerobic Sludge Blanket reactor UCBR Ultra-Compact Biofilm Reactor USA / U.S. United States of America UV Ultraviolet V Volume L VER Volumetric Exchange Ratio % VFD Variable Frequency Drives VOC Volatile Organic Compounds Vs min Minimum settling velocity for the biomass to be retained in the reactor m/h List of acronyms and symbols 285 v set Settling velocity m/h VSS Volatile Suspended Solids g/L v up Upflow velocity m/h W Weight WFD Water Framework Directive WWTP Wastewater Treatment Plant X Eff (or SSVEff) Biomass washed out in the effluent Y Yield coefficient removed g VSS g Substrate ΔN Nitrogen balance Δt Length of period d ΔW P Amount of produced biomass in a certain period g VSS ΔX r The change of biomass concentration during a certain period g VSS/L μ Biomass growth rate d-1 ρ Density g/L Sub index: air Air assimilated Assimilated biomass biomass denitrified Denitrified Eff Effluent G Gas Phase granule Granule Inf Influent L Liquid Phase obs Observed r Reactor recovered Recovered removed Removed s Soluble t Total LIST OF PUBLICATIONS International journal publications: Jungles, M. K., Figueroa, M., Morales, N., Val del Río, A., da Costa, R. H. R., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2011). Start up of a pilot scale aerobic granular reactor for organic matter and nitrogen removal. Journal of Chemical Technology and Biotechnology 86(5): 763-768. Val del Río, A., Morales, N., Isanta, E., Mosquera-Corral, A., Campos, J. L., Steyer, J. P. and Carrère, H. (2011). Thermal pre-treatment of aerobic granular sludge: Impact on anaerobic biodegradability. Water Research 45(18): 6011-6020. Vázquez-Padín, J. R., Fernández, I., Morales, N., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2011). Autotrophic nitrogen removal at low temperature. Water science and technology 63(6): 1282-1288. Isanta, E., Suárez-Ojeda, M. E., Val del Río, Á., Morales, N., Pérez, J. and Carrera, J. (2012). Long term operation of a granular sequencing batch reactor at pilot scale treating a low-strength wastewater. Chemical Engineering Journal (Lausanne) 198–199(0): 163-170. Val del Río, Á., Morales, N., Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2012). Effect of coagulant-flocculant reagents on aerobic granular biomass. Journal of Chemical Technology & Biotechnology 87(7): 908-913. Morales, N., Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2012). Aerobic granular-type biomass development in a continuous stirred tank reactor. Separation and Purification Technology 89(0): 199-205. Val del Río, A., Morales, N., Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2013). Effects of the cycle distribution on the performance of SBRs with aerobic granular biomass. Environmental Technology 34(11): 1463-1472. Morales, N., Boehler, M., Buettner, S., Liebi, C. and Siegrist, H. (2013). Recovery of N and P from Urine by Struvite Precipitation Followed by Combined Stripping with Digester Sludge Liquid at Full Scale. Water 5(3): 1262-1278. Morales, N., Figueroa, M., Fra-Vázquez, A., Val del Río, A., Campos, J. L., MosqueraCorral, A. and Méndez, R. (2013). Operation of an aerobic granular pilot scale SBR plant to treat swine slurry. Process Biochemistry 48(8): 1216-1221. List of publications 288 Vázquez-Padín, J. R., Morales, N., Gutiérrez, R., Fernández, R., Rogalla, F., Barrio, J. P., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2013). Implications of full scale implementation of an anammox based process as post-treatment of a municipal anaerobic sludge digester operated with co-digestion. Water science and Technology (Accepted) DOI: 10.2166/wst.2013.795. Spanish journal publications: Figueroa, M., Morales, N., Val del Río, A., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2011). Tratamiento de la fracción líquida de purín de cerdo en reactores granulares aerobios SBR. InfoEnviro 66: 108-111. Figueroa, M., Val del Río, A., Morales, N., Mosquera-Corral, A., Méndez, R. and Campos, J. L. (2011). Anammox: Hacia una eliminación sostenible del amonio en los efluentes de digestores anaerobios de purines. Revista Técnica de Medio Ambiente 24(151): 84-90. Vázquez-Padín, J.R., Morales, N., Gutiérrez, R., Fernández, R., Rogalla, F., MosqueraCorral, A., Campos, J.L., and Méndez, R. (2013). Bacterias implicadas en la eliminación de nitrógeno en retornos de fango de EDAR ─ Desarrollo y escalado del sistema ELAN® de eliminación autótrofa de nitrógeno. IX Jornadas de Transferencia de Tecnología sobre Microbiología del Fango Activo. Editorial GBS. Book Chapters: Figueroa, M., Val del Río, A., Morales, N., Campos, J. L., Mosquera-Corral, A. and R., M. (2009). Nitrogen removal in aerobic granular systems. In: Environmental Technologies to Treat Nitrogen Pollution, Principles and Engineering. Cervantes, F. J. (ed), IWA Publishing. London: 373-401. Mosquera-Corral, A., Figueroa, M., Morales, N., Val del Río, A., Campos, J.L. y Méndez, R. (2009). Tecnologías basadas en biomasa granular aerobia. En: Tecnologías Avanzadas para el Tratamiento de Aguas Residuales. 48-68. Editorial Lápices 4. Santiago de Compostela. Mosquera-Corral, A., Figueroa, M., Morales, N., Val del Río, A., Campos, J. L. and Méndez, R. (2010). Chapter 2: Aerobic granulation technology. In: Innovative Technologies for Urban Wastewater Treatment Plants. Omil, F. et al.. (ed), Editorial Lápices 4. Santiago de Compostela. Mosquera-Corral, A., Figueroa, M., Morales, N., Val del Río, A., Campos, J. L. and Méndez, R. (2012). Chapter 2: Aerobic granulation technology. In: Innovative Technologies for Urban Wastewater Treatment Plants (2nd Edition). Omil, F. et al.. (ed), Editorial Lápices 4. Santiago de Compostela. Campos, J. L., Figueroa, M., Morales, N., Fajardo, C., Vázquez-Padín, J. R., MosqueraCorral, A. and Méndez, R. (2012). Advanced Systems for Nitrogen Removal from Effluents List of publications 289 Produced in the Fish Canning Industry. In: Agricultural Research Updates. Volume 2. Hendriks, B. P. (ed), Nova Science Publishers 395-414 Conference Proceedings: Morales, N., Figueroa, M., Val del Río, A., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2008). Start up and operation of a pilot scale aerobic granular SBR. The Third International Meeting on Environmental Biotechnology and Engineering (3IMEBE). Palma de Mallorca (Spain). Morales, N., Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2010). Aerobic Granular Biomass in a Continuous Stirred Tank Reactor. Sustainable Solutions For Small Water and Wastewater Treatment Systems S2Small 2010. Girona (Spain). Jungles, M. K., Figueroa, M., Morales, N., Vázquez-Padín, J.R., da Costa, R. H. R., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2010). Start up of a pilot scale aerobic granular reactor. Sustainable Solutions For Small Water and Wastewater Treatment Systems S2Small 2010. Girona (Spain). Vázquez-Padín, J.R., Figueroa, M., Morales, N., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2010). Autotrophic nitrogen removal in granular reactors. Conferencia Nacional de Jóvenes Profesionales del Agua de España (Spain National Young Water Professionals Conference, Spain NYWPC), Barcelona (Spain). Vázquez-Padín, J.R., Fernández, I., Morales, N., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2010). Autotrophic nitrogen removal at low temperature. IWA World Water Congress and Exhibition. Montreal (Canada). Val del Río, A., Morales, N., Figueroa, M., Palmeiro, T., Mosquera-Corral, A., Campos, J. L. Suárez-Ojeda, M.E., Isanta, E. and Méndez, R. (2011). Potential anaerobic biodegradability of aerobic granular sludge. Water and Industry 2011. IWA Specialist Conference Chemical Industries. Valladolid (Spain). Val del Río, A., Morales, N., Figueroa, M., Mosquera-Corral, A., Campos, J. L. and Méndez, R. (2011). Effect of the Length of the Famine Period on Mature Aerobic Granular Biomass. 8th IWA Leading-Edge Conference on Water and Wastewater Technologies (LET 2011). Amsterdam (The Netherlands). Morales, N., Figueroa, M., Val del Río, A., Campos, J. L., Mosquera-Corral, A. and Méndez, R. (2011). Treatment of Swine Slurry in an Aerobic Granular Pilot-Scale Reactor. 8th IWA Leading-Edge Conference on Water and Wastewater Technologies (LET 2011), Amsterdam (The Netherlands). De Florio, L., Morales, N., Díez, R., Esteban, L., Rodriguez, L. and Tejero, I. (2011). Principles and basic technical devices for an Ecological Sanitation. In proceeding of: Industrial Ecology, ECOMONDO 2011, Rimini (Italy). Böhler, M., Büttner, S., Morales, N., Liebi, C., Schachtler, M. and Siegrist, H. (2012). Recovery of nutrients from ammonia rich sludge liquids and urine for the production of fertilizer