Assessment of the nitritation and anammox processes for mainstream wastewater treatment

TESE DE DOUTORAMENTO ASSESSMENT OF THE NITRITATION AND ANAMMOX PROCESSES FOR MAINSTREAM WASTEWATER TREATMENT Alba Pedrouso Fuentes ESCOLA DE DOUTORAMENTO INTERNACIONAL PROGRAMA DE DOUTORAMENTO EN ENXEÑARÍA QUÍMICA E AMBIENTAL SANTIAGO DE COMPOSTELA 2019 DECLARACIÓN DO AUTOR/A DA TESE ASSESSMENT OF THE NITRITATION AND ANAMMOX PROCESSES FOR MAINSTREAM WASTEWATER TREATMENT Dna. Alba Pedrouso Fuentes Presento a miña tese, seguindo o procedemento axeitado ao Regulamento, e declaro que: 1) A tese abarca os resultados da elaboración do meu traballo. 2) De selo caso, na tese faise referencia ás colaboracións que tivo este traballo. 3) A tese é a versión definitiva presentada para a súa defensa e coincide coa versión enviada en formato electrónico. 4) Confirmo que a tese non incorre en ningún tipo de plaxio doutros autores nin de traballos presentados por min para a obtención doutros títulos. En Santiago de Compostela., 7 de Outubro de 2019 Asdo. Alba Pedrouso Fuentes AUTORIZACIÓN DO DIRECTOR / TITOR DA TESE ASSESSMENT OF THE NITRITATION AND ANAMMOX PROCESSES FOR MAINTREAM WASTEWATER TREATMENT Dna. Anuska Mosquera Corral, Profesora Titular de Universidade de Enxeñaría Química Dna. Mª Ángeles Val del Río, Profesora Axudante Doutora de Enxeñaría Química e D. Ramón Méndez Pampín, Catedrático de Enxeñaría Química INFORMA/N: Que a presente tese, correspóndese co traballo realizado por Dna. Alba Pedrouso Fuentes, baixo a nosa dirección, e a utorizo a súa presentación, considerando que reúne os r equisitos esixidos no R egulamento de Estudos de Doutoramento da USC, e que como directores desta non incorre nas causas de abstención establecidas na Lei 40/2015. En Santiago de Compostela., 7 de Outubro de 2019 Asdo. Anuska Mosquera Corral Asdo. Mª Angeles Val del Río Asdo. Ramón Méndez Pampín Table of contents 1.1. Shift from wastewater treatment to water reuse and resource recovery 32 1.1.1. Water scarcity 32 1.1.2. Domestic wastewater reuse as a pool of resources 34 1.1.2.1. Reclaimed wastewater reuse 34 1.1.2.2. Energy recovery 39 1.1.2.3. Nutrients recovery 40 1.1.3. Nitrogen removal in municipal wastewater treatment: current status 41 1.1.3.1. Conventional activated sludge systems 44 1.1.4. Innovative wastewater treatment plants 44 1.1.4.1. Organic matter removal (A-Stage) 45 1.1.4.2. Nitrogen removal (B-Stage) 47 1.1.4.3. Decentralised systems 47 1.2. Nitrogen cycle and opportunities for municipal wastewater treatment 48 1.2.1. Nitrification (nitritation and nitratation) 49 1.2.2. Complete ammonia oxidising (comammox) bacteria 50 1.2.3. Heterotrophic denitrification (denitratation and denitritation) 51 1.2.4. Anaerobic ammonia oxidation: anammox 52 1.2.5. Other autotrophic denitrification processes 54 1.2.5.1. Methane autotrophic denitrification 54 1.2.5.2. Sulphur autotrophic denitrification 55 1.3. Towards efficient wastewater treatment: nutrient removal alternatives 57 1.3.1. Partial nitritation and anammox processes application in WWTPs 59 1.4. Application of PN/AMX processes at mainstream conditions 61 List of acronyms and symbols 1 Resumo 5 Resumen 17 Chapter 1. Introduction 29 Table of contents ii 1.4.1. Factors affecting NOB suppression 61 1.4.1.1. Aerobic sludge retention time 62 1.4.1.2. Dissolved oxygen concentration 62 1.4.1.3. Alternating anoxic and aerobic conditions 63 1.4.1.4. Aeration time control 63 1.4.1.5. Dissolved oxygen to total ammonium nitrogen concentration ratio 64 1.4.1.6. pH value with free nitrous acid and free ammonia concentrations 64 1.4.2. One-stage PN/AMX configuration at mainstream conditions 66 1.4.3. Two-stage configuration 68 1.4.3.1. Nitritation process 69 1.4.3.2. Anammox process 71 1.4.4. Main challenges of autotrophic nitrogen removal in mainstream 71 1.4.4.1. Low temperature 72 1.4.4.2. Low concentrations and high flows 72 1.4.4.3. Presence of organic matter 73 1.4.4.4. Wastewater alkalinity limitations 73 1.4.4.5. Fluctuation of wastewater characteristics 73 1.4.4.6. Source separation wastewater management 74 1.5. Aims and scope 74 1.6. References 77 2.1. Analysis of the liquid phase 98 2.1.1. Carbon compounds 98 2.1.1.1. Chemical oxygen demand 98 2.1.1.2. Dissolved total, organic and inorganic carbon 101 2.1.2. Nitrogen compounds 102 2.1.2.1. Total nitrogen 102 2.1.2.2. Ammonium 103 2.1.2.3. Nitrite 104 2.1.2.4. Nitrate 105 2.1.3. Inorganic ions 107 2.1.4. Alkalinity 108 2.1.5. Other control parameters 109 2.1.5.1. pH 109 2.1.5.2. Dissolved oxygen 109 2.1.5.3. Temperature 109 2.2. Biomass characterisation 109 Chapter 2. Materials and methods 95 Table of contents iii 2.2.1. Solid concentration 110 2.2.2. Solid concentration as biofilm 111 2.2.3. Sludge volume index 111 2.2.4. Density of the granules 112 2.2.5. Average diameter of the granules 113 2.2.6. Specific bacterial activities by batch tests 114 2.2.6.1. Specific anammox activity assays 114 2.2.6.2. Specific heterotrophic denitrifying activity assays 117 2.2.6.3. Specific aerobic activities by respirometric assays 117 2.3. Microbial population identification 120 2.3.1. Fluorescence in situ hybridisation 120 2.3.2. 16S rRNA gene-based amplicon analysis (Illumina®) 125 2.4. Calculations 128 2.4.1. Free ammonia and free nitrous acid calculation 128 2.4.2. Statistical analysis 128 2.4.3. Nitrogen transformation ratios and rates in the nitritation process 129 2.4.4. Nitrogen removal rates in the anammox process 129 2.4.5. Mass balances in one-stage partial nitritation-anammox systems 130 2.4.6. Nitrogen removal rates in one-stage PN/AMX system 133 2.5. References 133 3.1. Introduction 137 3.2. Objectives 139 3.3. Materials and Methods 140 3.3.1. Reactor setup and operation 140 3.3.2. Ex-situ specific activity tests in batch mode 142 3.3.3. Analytical methods 143 3.3.4. Calculations 143 3.4. Results and discussion 144 3.4.1. Performance of the PN/AMX processes 144 3.4.2. Biomass and involved activities 147 Chapter 3. PN/AMX system robustness under repeated starvation and reactivation periods for blackwater treatment 135 Table of contents iv 3.4.3. Treatment of blackwater via PN/AMX processes 150 3.4.4. PN/AMX system robustness under repeated starvation/reactivation periods 151 3.5. Conclusions 153 3.6. References 153 4.1. Introduction 159 4.2. Objectives 161 4.3. Materials and Methods 161 4.3.1. Reactor setup and operational conditions 161 4.3.2. Ex-situ specific activity tests in batch mode 163 4.3.3. Analytical methods 163 4.3.4. Calculations 163 4.4. Results and discussion 164 4.4.1. Hydraulic load effect on AOB and NOB competition 164 4.4.2. Effectiveness of SRT to achieve the nitritation process 169 4.4.3. FA and FNA concentrations contribution to decouple AOB and NOB rates 170 4.4.4. Evolution of the AOB and NOB rates and its practical implications 171 4.5. Conclusions 172 4.6. References 172 5.1. Introduction 179 5.2. Objectives 180 5.3. Materials and Methods 181 5.3.1. Reactor setup and operation 181 5.3.2. Batch specific activity tests 183 5.3.3. Analytical methods 183 5.3.4. Calculations 184 5.4. Results 185 5.4.1. Nitritation reactor operation 185 Chapter 4. Is the hydraulic load a parameter useful to uncouple AOB and NOB activities? 157 Chapter 5. Nitrite oxidising bacteria suppression based on in situ free nitrous acid production 177 Table of contents v 5.4.2. Long-term effect of sodium azide on AOB 191 5.4.3. Biomass characteristics and NOB populations 191 5.5. Discussion 194 5.5.1. FNA accumulation as a strategy to obtain stable nitritation 194 5.5.2. Influence of the inlet ammonium to inorganic carbon ratio 197 5.5.3. Nitrite accumulation activation 200 5.5.4. Effective NOB washout 202 5.6. Conclusions 203 5.7. References 204 6.1. Objectives 213 6.2. Materials and Methods 214 6.2.1. Reactor setup and operational conditions 214 6.2.2. Biomass batch specific activity tests 216 6.2.3. Analytical methods 216 6.2.4. Calculations 217 6.3. Results 217 6.3.1. Nitritation establishment by natural in situ FNA accumulation 217 6.3.2. Performance of the nitritation and organic matter oxidation 221 6.3.3. Oxidation processes restoration after anoxic starvation 222 6.3.4. Optimisation of the operational cycle to treat very low nitrogen concentrations 223 6.4. Discussion 226 6.4.1. Nitritation process treating municipal wastewater in the presence of organic matter 226 6.4.2. Robustness and feasibility of the in situ FNA inhibitory concentration strategy 227 6.5. Conclusions 231 6.6. References 232 Chapter 6. Performance of nitritation and organic matter oxidation processes with municipal wastewater 209 Table of contents vi 7.1. Introduction 239 7.2. Objectives 241 7.3. Materials and Methods 242 7.3.1. Reactor description and operating conditions 242 7.3.2. Microbial activity batch tests 245 7.3.3. Response surface methodology 245 7.3.4. Analytical methods 247 7.3.5. Calculations 247 7.4. Results and discussion 248 7.4.1. Anammox process establishment and maintenance 248 7.4.1.1. Anammox reactor start-up 248 7.4.1.2. Anammox process performance at decreasing alkalinity concentrations 251 7.4.1.3. Dependence of the anammox activity with temperature 253 7.4.2. Anammox process performance treating nitritified municipal wastewater 256 7.4.3. Biomass retention 258 7.4.4. Effect of pH, COD and temperature over the specific maximum anammox activity 260 7.5. Conclusions 265 7.6. References 266 8.1. Introduction 273 8.2. Objectives 275 8.3. Materials and Methods 275 8.3.1. IFAS pilot plant setup 275 8.3.2. IFAS operational conditions 277 8.3.3. Analytical methods 278 8.3.4. Microbial activity batch tests 279 8.3.5. Identification of microbial populations 279 8.3.6. Calculations 280 8.4. Results 280 Chapter 7. Assessment of the anammox process performance 237 Chapter 8. Performance of one-stage partial nitritation-anammox processes in an IFAS system 271 Resumo No eido da economía circular, as augas residuais xa non son consideradas residuos que teñen que ser tratados, senón unha fonte de auga para o reemprego, de enerxía, de nutrientes e de outros recursos que poden e deben ser valorizados. A agricultura é o sector cun maior consumo de auga representando o 69 % do total da auga doce extraída mundialmente. Cando a auga residual se reemprega, esta usase principalmente para fertilizar e irrigar neste sector, supoñendo un 32 % do total de auga reempregada. Ademais de constituír unha fonte alternativa de auga, as augas residuais proporcionan nutrientes (nitróxeno e fósforo) reducindo a necesidade de aplicar fertilizantes químicos. Sen embargo, a escorrentía de nitratos derivada da aplicación excesiva de nutrientes na agricultura constitúe un dos maiores problemas de calidade de auga a nivel mundial. De feito, a lexislación sobre o reemprego de augas establece que o seu contido de nutrientes debe avaliarse axeitadamente tendo en conta o equilibrio dos mesmos no solo. Este feito, xunto coa gran variabilidade estacional da demanda de auga na agricultura, fan que sexa necesario desenvolver alternativas para a eliminación de nitróxeno que permitan outros usos da auga ou, en última instancia, a súa descarga ao medio acuático. Na actualidade, a eliminación de nitróxeno non se aplica aínda en todas as estacións depuradoras de augas residuais (EDAR) sendo unha práctica común nas EDAR de gran tamaño, pero apenas aplicada nas EDAR de pequenos núcleos de poboación (sistemas descentralizados). Na lexislación sobre tratamento de augas residuais de moitos países non se establecen límites para a eliminación de nitróxeno como é o caso da maioría dos países do Centro e Sur de América, mentres que en Europa tan só é obrigatorio nas EDAR que tratan cargas de máis de 10.000 habitantes equivalentes. Ademais da carencia de imposicións legais, a elevada demanda de enerxía asociada aos procesos convencionais de eliminación de nitróxeno limitou a súa implementación nas EDAR. O tratamento de augas residuais representa o 42 % do consumo enerxético do sector da auga nos países desenvolvidos. Con todo, as Resumo 6 augas residuais conteñen máis enerxía que a que potencialmente fai falta para o seu tratamento. O nitróxeno é habitualmente eliminado nos sistemas convencionais de lamas activas mediante a combinación dos procesos de nitrificación-desnitrificación que requiren unha elevada cantidade de enerxía para airear e materia orgánica que non se valoriza. A optimización dos procesos de eliminación de nitróxeno mediante procesos autótrofos podería converter ás EDAR en produtoras netas de enerxía e facilitar o reemprego da auga facendo un uso máis eficiente dos recursos presentes na auga residual. Entre os diferentes procesos biolóxicos de eliminación de nitróxeno, a combinación dos procesos de nitritación parcial e anammox (NP/AMX) ten a vantaxe de reducir o consumo de enerxía un 60 %, xa que tan só é necesario oxidar a metade do contido amoniacal a nitrito e, ao ser procesos completamente autótrofos, producen menos lamas e non precisan materia orgánica. Polo tanto, a totalidade da materia orgánica pode transformarse en biogás maximizando a recuperación de enerxía. O obxectivo principal de esta tese é avaliar a aplicación dos procesos de NP/AMX para a eliminación autótrofa do nitróxeno presente na liña principal das EDAR. Primeiro, estudouse a aplicación dos procesos NP/AMX nunha etapa para o tratamento das augas negras orixinadas nun sistema descentralizado con redes separativas coma unha alternativa axeitada para as EDAR de pequeno tamaño (Capítulo 3). Logo, avaliouse a viabilidade de aplicar os procesos baseados en NP/AMX para o tratamento de augas residuais municipais caracterizadas por ter concentracións de nitróxeno máis baixas (< 50 mg N/L), baixa temperatura (< 25 °C) e pola grande variabilidade nas súas características. Estudos previos sinalaron a supresión das bacterias oxidantes de nitrito (BON) como un dos principais retos para a aplicación dos procesos NP/AMX na liña principal das EDAR. A obtención do proceso de nitritación estudouse mediante dúas estratexias diferentes: diminuír o tempo de residencia hidráulico (TRH) nun quimiostato (Capítulo 4) ou inhibir as BON mediante a acumulación in situ de ácido nitroso libre (ANL) tratando tanto auga residual sintética (Capítulo 5) como auga residual municipal (Capítulo 6). A continuación, avaliouse a viabilidade e a estabilidade do proceso anammox operado a baixa temperatura (Capítulo 7). Ademais, explorouse a aplicación dos procesos NP/AMX alimentados con augas residuais municipais a escala piloto empregando dúas configuracións de reactor diferentes: un reactor híbrido de lamas activas de Resumo 7 leito fixo (IFAS, do inglés Integrated Fixed-biofilm Activated Sludge) (Capítulo 8), onde os procesos de NP/AMX ocorren simultaneamente, e nunha configuración de dúas etapas onde ambos procesos están separados en dous reactores distintos (Capítulo 9). Finalmente, discútense as conclusións máis relevantes obtidas así como as principais limitacións identificadas (Capítulo 10). Nas seguintes seccións, amósase un resumo dos contidos principais de cada un dos capítulos da presente tese. Capítulo 1. Introdución No Capítulo 1 amósase unha visión global do novo paradigma do tratamento de augas residuais nas que a escaseza de auga promove o reemprego das mesmas unha vez tratadas e unha mellor xestión dos recursos durante o seu tratamento para valorizar a enerxía. Ademais, inclúese unha revisión do estado actual da investigación e aplicación dos procesos de NP/AMX así como os factores principais que inflúen na selección das BON. Tamén se sinalan os principais desafíos para a aplicación dos procesos de NP/AMX tendo en conta as características principais das augas residuais na liña principal das EDAR. Para rematar, preséntanse os obxectivos principais e específicos da presente tese. Capítulo 2. Materiais e Métodos No Capítulo 2 descríbense con detalle os materiais e métodos empregados durante o traballo experimental descrito na presente tese. Estes inclúen as análises da fase líquida e sólida, así como os métodos para determinar as actividades específicas das diferentes poboacións bacterianas. Ademais, neste capítulo tamén se presenta a descrición minuciosa dos cálculos empregados nos diferentes capítulos para a determinación, entre outros, das velocidades de transformación e a eficiencia dos procesos biolóxicos. As descricións específicas dos reactores e montaxes experimentais amósanse nos capítulos correspondentes. Capítulo 3. Robustez do sistema de NP/AMX baixo períodos repetidos de parada e reactivación para o tratamento das augas negras No Capítulo 3, avaliouse a aplicación dun sistema en unha etapa cos procesos simultáneos de NP/AMX para tratar, a temperatura ambiente (14 - 21 °C), augas negras dixeridas anaeróbicamente procedentes dun sistema de tratamento de augas residuais descentralizado con redes separativas. As augas residuais procedían dun edificio de oficinas e a principal característica deste tipo de sistemas Resumo 8 descentralizados é que deben facer fronte á falta de dispoñibilidade de augas residuais para tratar de xeito continuado. Así, operouse un reactor secuencial (SBR, do inglés Sequencing Batch Reactor), de 4 L, aplicando períodos repetidos de parada, durante as noites e as fins de semana, e posterior reactivación. A pesar da baixa temperatura, as concentracións moderadas de nitróxeno total (NT, 120 mg NT/L) e as paradas regulares, logrouse manter o proceso de eliminación de nitróxeno estable durante 100 días. Tras a implementación dunha fase anóxica para refinar a calidade do efluente utilizando a materia orgánica residual (100 mg/L como demanda química de osíxeno (DQO)), obtivéronse eficiencias de eliminación de nitróxeno do 95 % acadando velocidades de eliminación de nitróxeno (VEN) de aproximadamente 66 mg NT/(L·d). Ademais, o sistema NP/AMX recuperou inmediatamente (nun día) a súa capacidade de eliminación de nitróxeno despois de estar parado durante 15 días simulando un período de vacacións. O sistema mostrou unha excelente capacidade de retención de biomasa obtendo unha concentración de sólidos en suspensión volátiles (SSV) no efluente de entre 7 e 16 mg SSV/L. Durante todo o período de operación, a actividade das BON foi suprimida con éxito. Por outra banda, observouse a segregación das diferentes poboacións bacterianas sendo as bacterias oxidantes de amonio (BOA) e as heterótrofas (tanto aerobias como desnitrificantes) máis activas na biomasa floculenta mentres que a biomasa granular estaba enriquecida en bacterias anammox. Con respecto á calidade do efluente, este contiña baixas concentracións de nitróxeno (≤ 10 mg NT/L), de materia orgánica (≤ 30 mg DQO/L) e de sólidos (≤ 20 mg SST/L) cumprindo cos límites de descarga establecidos na Directiva europea de tratamento de augas residuais urbanas (91/271/CEE) e os requisitos mínimos de calidade para o reemprego da auga definidos pola Unión Europea (UE) en TA(2019)0071. Capítulo 4. É a carga hidráulica un parámetro útil para desacoplar as actividades BOA e BON? Neste capítulo, estudouse a selección con fins de enriquecemento das BOA sobre as BON de acordo coas súas velocidades de crecemento. Para iso, operouse un reactor continuo de tanque axitado (CSTR, do inglés Continuos Stirred Tank Reactor), de 7.15 L a 16 ± 1 °C, reducindo progresivamente o TRH dende 7,0 a 1,5 días durante 158 días e alimentado con auga residual sintética (50 mg NT/L). O reactor inoculouse con lamas activas nitrificantes. A acumulación de nitrito activouse a valores de TRH de 4,6 ± 0,2 días, pero o nitrato seguía a ser o principal produto de oxidación do Resumo 9 amonio representado o 80 %, mesmo cando o TRH se reduciu a 1,5 días. A pesar de que non se logrou establecer o proceso de nitritación con éxito, esta experiencia demostrou que a aplicación dunha elevada carga hidráulica podería brindar unha vantaxe competitiva das BOA sobre as BON. A relación entre as velocidades de reacción de ambos grupos de nitrificantes aumentou progresivamente alcanzando valores de 1,5 g N-NH4+consumido/g N-NO2-consumido. Por tanto, foi posible activar a acumulación de nitrito pero non obter un lavado eficaz das BON, que requiriría a combinación da aplicación da carga hidráulica con outros parámetros para acadar e manter con éxito o proceso de nitritación. Capítulo 5. Supresión das bacterias oxidantes de nitrito baseada na produción in situ de ácido nitroso libre No Capítulo 5 propúxose e avaliouse unha nova estratexia para establecer o proceso de nitritación. A maior sensibilidade das BON con respecto ás BOA fronte ó ANL permite promover o proceso de nitritación mediante a exposición das BON a concentracións de ANL superiores a 0,02 mg N-HNO2/L. As concentracións inhibitorias de ANL xeráronse dentro do reactor mediante o proceso de oxidación de amonio por si mesmo, aproveitando a produción de nitrito e a diminución do pH asociada ao consumo de alcalinidade durante a nitritación. Esta alternativa simple explorouse nas condicións da liña principal de augas, tratando auga residual simulada (50 mg NT/L) a 16 ° C, nun reactor SBR de 2 L, durante 690 días. Así, se obtivo e mantivo estable a longo prazo o proceso de nitritación acadando concentracións de ANL de ata 0,06 mg N-HNO2/L. Dado que o lodo inoculado tiña actividades BOA e BON similares, probáronse dous estimuladores de acumulación de nitrito: azida de sodio e nitrito. A análise microbiolóxica revelou que as poboacións de BON (sendo dominante o xénero Nitrospira) foron eliminadas do reactor. O éxito da estratexia de acumulación de ANL in situ baséase na relación entre as concentracións de nitróxeno e carbono inorgánico (CI) das augas residuais. Por iso, avaliáronse diferentes proporcións desta relación (entre 0,5 - 1,0 g N/g CI), observando que cando é menor a 0,6 g N/g CI, as concentracións de ANL non son suficientes para inhibir as NOB. Baixo estas condicións non inhibitorias, as BON demoraron aproximadamente 40 días en desenvolver unha actividade significativa. Este longo período amosa a robustez da estratexia baseada no ANL en canto ao eficaz lavado de BON. Durante a operación do reactor de nitritación, non se observou ningún efecto negativo sobre a actividade BOA asociado ás concentracións de ANL Resumo 10 acadadas ou do baixo pH, obtendo valores de actividade de ata 400 mg N-NH4+/(g SSV·d). Capítulo 6. Comportamento dos procesos de nitritación e oxidación de materia orgánica tratando augas residuais municipais No Capítulo 6, inoculouse un segundo SBR (de 2 L) para levar a cabo o proceso de nitritación empregando un lodo enriquecido en BOA e con actividade insignificante de BON. O proceso de nitritación estableceuse e mantivose con éxito mediante a estratexia de acumulación in situ de ANL tratando auga sintética (50 mg NT/L) a 15 ± 1 °C. Logo, a operación continuouse pero tratando augas residuais municipais despois do tratamento primario (que contiñan 18 - 55 mg N-NH4+/L e unha concentración de carbono orgánico total (COT) de 20 - 50 mg COT/L), para estudar a viabilidade do proceso de nitritación baseado na estratexia do ANL. A acumulación de nitrito mantívose en valores de practicamente o 100 % do amonio oxidado e a estabilidade do proceso de nitritación sostívose a longo prazo (354 días dos cales 191 tratou auga municipal), mediante a acumulación in situ de ANL en concentracións entre 0,02 e 0,20 mg N-HNO2/L. Ademais, aínda que se alcanzaron valores mínimos de pH de 5,6, estes non produciron ningún efecto adverso sobre o proceso de nitritación nin sobre a actividade específica das BOA que se mantivo ao redor de 200 mg N-NH4+/(g SSV·d). Independentemente da presenza de materia orgánica, que tamén fora sinalada previamente como un reto para a aplicación dos procesos de NP/AMX en condicións da liña principal de augas, o proceso de nitritación permaneceu estable e o 80 % da materia orgánica foi eliminada no mesmo reactor. Este feito confire a esta estratexia unha gran vantaxe, xa que permite ser máis flexible coas eficiencias de eliminación requiridas nas unidades anteriores para a eliminación de materia orgánica. Por outra banda, observouse que para facer fronte ás flutuacións das concentracións de nitróxeno e materia orgánica das augas residuais, requírese a optimización da duración do ciclo de operación do SBR. Por tanto, a duración variable do ciclo proporcionaría unha operación máis axeitada evitando períodos inactivos cando o amonio se consume por completo ao tratar auga residual máis diluída. Capítulo 7. Avaliación do comportamento do proceso anammox O Capítulo 7 baséase no estudo da viabilidade da aplicación do proceso anammox en condicións da liña principal de augas das EDAR así como o seu Resumo 11 comportamento a baixa temperatura a longo prazo. Con este obxectivo, inoculouse un SBR (cun volume de 5 L) con lodo granular procedente dun reactor ELAN® (ELiminación Autótrofa de Nitróxeno) que trata augas residuais con concentracións elevadas de nitróxeno en condicións mesófilas e expúxose directamente ás concentracións baixas de nitróxeno (50 mg NT/L) e baixa temperatura (15 °C) características da liña principal das EDAR. Os resultados demostraron que o período de aclimatación progresiva ás novas condicións non sería necesario, acurtando os períodos de arranque, xa que se obtiveron resultados de eficacias de eliminación similares aos publicados por outros autores que realizaron unha adaptación lenta da biomasa ás novas condicións. Ademais, demostrouse a estabilidade do proceso anammox a longo prazo (485 días) tratando tanto augas residuais sintéticas con alcalinidades decrecentes así como con auga residual municipal nitritificada. Esta última consistía no efluente do reactor de nitritación descrito no Capítulo 6 mesturado con auga municipal sen tratar no devandito reactor, para axustar a relación nitrito/amonio necesaria para o proceso anammox en 1,2 g N-NO2-/g NNH4+. A pesar de que no caso da operación con auga real, a alimentación do reactor anammox tiña concentracións de alcalinidade baixas (2 ± 1 mg CI/L) e valores de pH ao redor de 6,2, estes non afectaron á estabilidade do proceso. Por outra banda, a presenza de concentracións residuais de materia orgánica da auga residual municipal tras o tratamento primario contribuíu a un aumento da eficiencia de eliminación de nitróxeno dende valores en torno ao 80 % a 90 %, acadando concentracións de NT no efluente inferiores a 10 mg NT/ L (límite de vertedura na Unión Europea nas áreas sensibles para grandes depuradoras). A biomasa anammox retívose satisfactoriamente no sistema e a súa actividade específica aumentou ao longo da operación dende 53 ± 11 mg N/(g SSV·d), no inóculo, ata valores máximos de 78 ± 8 mg N/(g SSV·d) ao final da operación con auga sintética. Con todo, durante o período de operación con auga municipal observouse un descenso da actividade a valores medios de 58 ± 6 mg N/(g SSV·d), probablemente debido ao desenvolvemento de bacterias heterótrofas desnitrificantes. Considerando a actividade específica anammox e a concentración de biomasa no reactor, o sistema tiña unha capacidade de eliminación potencial maior á VEN tratada (40 mg NT/(L·d)) indicando que esta estaba limitada pola carga aplicada. Por outra banda, estudouse o efecto do pH (6 - 8), temperatura (15 - 30 °C) e concentración COT (0 - 75 mg COT/ L) sobre a actividade específica anammox en ensaios en descontinuo. Os resultados indican Resumo 12 que o pH é o parámetro ensaiado con maior efecto sobre a actividade e que o rango de pH óptimo varía coa temperatura. Capítulo 8. Comportamento dos procesos de nitritación parcial-anammox nunha etapa nun sistema IFAS No Capítulo 8 preséntase o reactor IFAS como unha configuración prometedora para levar a cabo os procesos de NP/AMX nunha etapa dado á súa capacidade de segregar as diferentes poboacións bacterianas en distintas fraccións de biomasa (biomasa en suspensión e biopelícula). O reactor IFAS a escala piloto (200 L) alimentouse de xeito continuo con auga residual municipal pretratada anaeróbicamente e operouse a temperaturas decrecentes dende 21 a 15 ° C. A pesar de que a alimentación contiña unha DQO relativamente alta con relacións DQO/N de 2.5 ± 0.3 g DQO/g N, os balances de materia indican que o proceso anammox é o maior responsable da eliminación de nitróxeno obtida. Durante a operación obtivéronse eficiencias de eliminación de nitróxeno de 72 ± 11 % e valores medios de VEN de 37 ± 3 mg NT/(L·d) a 15 °C. Con respecto á distribución das poboacións principais, observouse unha segregación sendo as BOA as máis abundantes na biomasa en suspensión mentres que as bacterias anammox se atopaban principalmente na biopelícula formada sobre os soportes. Con todo, a abundancia das BON era similar en ambas fraccións de biomasa e, a pesar de que a súa actividade específica diminuíu durante a operación do reactor, non foron eliminadas por completo do sistema. Adicionalmente, o patrón de aireación intermitente empregado no reactor IFAS permitiu a limitación da actividade das BON dentro do sistema observándose unha actividade no reactor de en torno ao 10 - 20 % da actividade máxima. Con respecto á calidade do efluente, obtivéronse concentracións de nitróxeno de 12 ± 5 mg NT/ L, o que demostra que se require unha maior optimización do sistema para cumprir cos límites de vertedura establecidos pola lexislación europea para as áreas sensibles: 15 mg NT/L para as EDAR que tratan cargas de entre 10.000 e 100.000 habitantes equivalentes e 10 mg NT/ L para as de maior tamaño. Capítulo 9. Aplicación dun sistema de nitritación parcial e anammox en dúas etapas a escala piloto Os prometedores resultados obtidos a escala laboratorio coa configuración de NP/AMX en dúas etapas conduciron a probar esta configuración por primeira vez a Resumo 13 escala piloto (600 L cada unidade) como se describe no Capítulo 9. Neste caso, a planta piloto instalouse nunha EDAR municipal onde operaba un reactor aerobio de lodos activos de alta carga para a eliminación de materia orgánica (sen afectar á concentración de nitróxeno). O sistema NP/AMX operouse sen control de temperatura (variando entre 11 e 28 °C) e facendo fronte ás mesmas flutuacións nas características da auga residual ás que estaba sometida a EDAR municipal. O efluente do reactor aerobio de alta carga caracterizábase por ter concentracións de amonio entre 22 e 63 mg N-NH4+/L, que se oxidou parcialmente a nitrito na unidade de nitritación parcial. Durante toda a operación non se observou actividade das BON que estaban inhibidas mediante a acumulación de ANL in situ (0,015 - 0,200 mg NHNO2/L). Para obter unha relación nitrito/amonio no efluente de 1,2 g N-NO2-/ g NNH4+, axeitada para alimentar ao reactor anammox en serie, implementouse unha estratexia de control para a etapa de aireación baseándose no valor de pH no interior do reactor de nitritación. Deste xeito, o reactor puido soportar mellor as flutuacións nas características das augas residuais. Por outra banda, no reactor anammox alcanzouse unha eficiencia de eliminación de nitróxeno do 80 % cunha VEN de 99 ± 26 mg NT/(L·d), limitada principalmente polo TRH imposto previamente na unidade de NP para lograr unha adecuada retención de biomasa. De feito, a principal limitación atopada nesta experiencia foi a retención de biomasa nesta unidade. Con respecto á calidade do efluente, obtívose unha concentración de 12 ± 3 mg NT/ L. Con todo, é necesario sinalar que o efluente do reactor aerobio de alta carga da EDAR alimentado ao sistema NP/AMX xa contiña aproximadamente 5 mg N-NO3-/L de nitrato. El nitrato non é eliminado no sistema NP/AMX, limitando por tanto a súa eficacia total de eliminación de nitróxeno. A optimización do sistema aerobio de alta carga da EDAR, mediante a redución do tempo de retención celular para limitar o proceso de nitrificación nesta unidade e evitar a formación de nitrato, fomentaría a obtención dun efluente de maior calidade na saída do sistema NP/AMX, cumprindo cos límites de vertedura establecidos comunmente na lexislación. Finalmente, realizouse un estudo dos potenciais aforros comparando a combinación do sistema aerobio de alta carga e a eliminación de nitróxeno mediante NP/AMX cun sistema convencional de lodos activos (a tecnoloxía máis comunmente empregada para a eliminación de nitróxeno nas EDAR). Os resultados amosaron que a implementación dos procesos de eliminación autótrofa de nitróxeno permite duplicar a recuperación de enerxía da planta xa que máis materia orgánica é tratada no dixestor anaerobio de lodos producindo máis biogás. Ademais, a eficiencia enerxética da planta Resumo 14 aumenta do 21 % cando se aplica o sistema convencional de lodos activos ao 57 % coa aplicación dos procesos de NP/AMX, e os custos de operación tamén se reducen á metade. Con todo, a produción de lamas que deben ser xestionadas como residuo é menor no escenario convencional xa que o reactor aerobio de alta carga ten unha produción elevada de lamas incrementando os custos asociados á súa xestión. Capítulo 10. Conclusións xerais e futuros desafíos para a eliminación autótrofa de nitróxeno na liña principal de augas No capítulo 10 preséntase unha discusión xeral co obxectivo de integrar os principais resultados e conclusións máis relevantes obtidos durante a realización desta tese. En xeral, os resultados desta tese contribúen a incrementar a comprensión do proceso de NP/AMX e proporcionan información interesante sobre a viabilidade de aplicar os procesos de NP/AMX para eliminar de xeito autótrofo o nitróxeno a baixa temperatura. Así téntase responder a algunhas das principais cuestións que fican abertas neste eido: i) É posible aplicar sistemas baseados no proceso anammox en sistemas de tratamento de augas descentralizados? ii) É posible suprimir a actividade das BON en condicións da liña principal de augas? iii) Son capaces as bacterias anammox de tratar cargas suficientemente altas de xeito estable para a eliminación de nitróxeno na liña de augas? iv) Como afecta a composición da auga residual municipal ao comportamento dos procesos NP/AMX? v) Cal é a configuración de reactor máis eficaz para a implementación dos procesos de NP/AMX para o tratamento de augas domésticas? Demostrouse que o tratamento das augas negras mediante un sistema NP/AMX nunha etapa é viable, probablemente porque a súa composición é relativamente máis alta e estable en comparación coas augas residuais municipais, unha vantaxe para a estabilidade do proceso de NP/AMX. Os sistemas descentralizados, especialmente os de pequena escala, poderían enfrontarse a períodos de ausencia de augas residuais para tratar, pero comprobouse que as paradas que se ocasionarían no sistema de tratamento non impiden a aplicación dos procesos de NP/AMX. Así, poderíase incrementar o número de pequenas instalacións con sistemas de eliminación de nitróxeno para promover o reemprego Resumen 21 desde 7,0 a 1,5 días durante 158 días alimentado con agua residual sintética (50 mg NT/L). El reactor se inoculó con lodos activos nitrificante. La acumulación de nitrito se activó cuando el TRH se estableció en valores de 4,6 ± 0,2 días, pero el nitrato seguía siendo el principal producto de oxidación del amonio representado el 80 % incluso cuando el TRH se redujo a 1,5 días. A pesar de que no se logró establecer el proceso de nitritación con éxito, se demostró que la aplicación de un elevado estrés hidráulico podría brindar una ventaja competitiva de las BOA sobre las BON. La relación entre las velocidades de reacción de ambos grupos de nitrificantes aumentó progresivamente alcanzando valores de 1,5 g N-NH4+consumido/g N-NO2-consumido. Por lo tanto, fue posible activar la acumulación de nitrito pero no obtener un lavado de las BON eficaz, que requeriría la combinación de la aplicación del estrés hidráulico con otros parámetros para lograr y mantener con éxito el proceso de nitritación. Capítulo 5. Supresión de las bacterias oxidantes de nitrito basada en la producción in situ de ácido nitroso libre En el Capítulo 5 se propuso y evaluó una nueva estrategia para establecer el proceso de nitritación. El hecho de que las BON sean más sensibles que las BOA al ANL permite promover el proceso de nitritación mediante la exposición de las BON a concentraciones de ANL superiores a 0,02 mg N-HNO2/L. Las concentraciones inhibitorias de ANL se generaron dentro del reactor mediante el proceso de oxidación de amonio por sí mismo, aprovechando la producción de nitrito y la disminución del pH asociada al consumo de alcalinidad durante la nitritación. Esta alternativa simple se exploró en las condiciones de la línea principal de aguas, tratando agua residual simulada (50 mg NT/L) a 16 °C en un reactor SBR de 2 L durante 690 días. Así, se obtuvo y mantuvo estable a largo plazo el proceso de nitritación alcanzado concentraciones de ANL de hasta 0,06 mg N-HNO2/L. Debido a que el lodo inoculado tenía actividades BOA y BON similares, se probaron dos estimuladores de acumulación de nitrito: azida de sodio y nitrito. El análisis microbiológico reveló que las poblaciones de BON (siendo el género Nitrospira el dominante) fueron eliminadas del reactor. El éxito de la estrategia de acumulación de ANL in situ se basa en la relación entre las concentraciones de nitrógeno y carbono inorgánico (CI) de las aguas residuales. Por eso se evaluaron diferentes proporciones de esta relación (0,5 - 1,0 g N/g CI), observando que cuando es menor a 0,6 g N/g CI, las concentraciones de ANL no son suficientes para inhibir las NOB. Bajo estas condiciones no inhibitorias, las BON demoraron aproximadamente 40 días en Resumen 22 desarrollar una actividad significativa. Este largo período muestra la robustez de la estrategia basada en el ANL en cuanto al eficaz lavado de BON. Durante la operación del reactor de nitritación, no se observó ningún efecto negativo asociado a las concentraciones de ANL alcanzadas o del bajo pH sobre la activad BOA, obteniendo valores de actividad de hasta 400 mg N-NH4+/(g SSV·d). Capítulo 6. Comportamiento de los procesos de nitritación y oxidación de materia orgánica tratando aguas residuales municipales En el Capítulo 6, se inoculó un segundo SBR (de 2 L) para llevar a cabo el proceso de nitritación utilizando un lodo enriquecido en BOA y con actividad insignificante de BON. El proceso de nitritación se estableció y mantuvo con éxito mediante la estrategia de acumulación in situ de ANL tratando agua sintética (50 mg NT/L) a 15 ± 1 °C. Luego, la operación continuó pero tratando aguas residuales municipales después del tratamiento primario (que contenían 18 - 55 mg N-NH4+/L y concentración de carbono orgánico total (COT) de 20 - 50 mg COT/L), para estudiar la viabilidad del proceso de nitritación basado en la estrategia del ANL. Se observó una acumulación de nitrito de prácticamente el 100 % del amonio oxidado y la estabilidad del proceso de nitritación se mantuvo a largo plazo (354 días de los cuales 191 trató agua municipal), mediante la acumulación de ANL en concentraciones variando entre 0,02 y 0,20 mg N-HNO2/L. Además, aunque se alcanzaron valores mínimos de pH de 5,6, estos no produjeron ningún efecto adverso sobre el proceso de nitritación ni sobre la actividad específica de las BOA que se mantuvo alrededor de 200 mg N-NH4+/(g SSV·d). Independientemente de la presencia de materia orgánica, que también había sido señalada como un reto para la implementación de los procesos de NP/AMX en condiciones de la línea principal de aguas, el proceso de nitritación se mantuvo estable y el 80 % de la materia orgánica fue eliminada en el mismo reactor. Este hecho confiere a esta estrategia una gran ventaja, ya que permite ser más flexible con las eficiencias de eliminación requeridas en las unidades anteriores para la eliminación de materia orgánica. Se observó que, para hacer frente a las fluctuaciones de las concentraciones de nitrógeno y materia orgánica de las aguas residuales, se requiere la optimización de la duración del ciclo de operación del SBR. Por lo tanto, la duración variable del ciclo proporcionaría una operación más adecuada evitando períodos inactivos cuando el amonio es consumido por completo al tratar agua residual diluida. Resumen 23 Capítulo 7. Evaluación del comportamiento del proceso anammox El Capítulo 7 se basa en el estudio de la viabilidad de aplicar el proceso anammox en condiciones de la línea principal de aguas de las EDAR, así como su comportamiento a baja temperatura a largo plazo. Con este objetivo, se inoculó un SBR (con un volumen de 5 L) con lodo granular procedente de un reactor ELAN® (ELiminación de Nitrógeno Autótrofa) que trata aguas residuales con concentraciones elevadas de nitrógeno en condiciones mesofílicas y se expuso directamente a las concentraciones bajas de nitrógeno (50 mg NT/L) y baja temperatura (15 °C) características de la línea principal de las EDAR. Los resultados demostraron que el período de aclimatación progresiva a las nuevas condiciones no sería necesario, acortando los períodos de arranque, ya que se obtuvieron eficacias de eliminación similares a los publicadas por otros autores realizando una adaptación lenta de la biomasa a las nuevas condiciones. Además, se demostró la estabilidad del proceso anammox a largo plazo (485 días) tratando tanto aguas residuales sintéticas con alcalinidades decrecientes, así como con agua residual municipal nitritificada. Esta última consistía en el efluente del reactor de nitritación descrito en el Capítulo 6 mezclado con agua municipal sin tratar en dicho reactor, para ajustar la relación nitrito/amonio necesaria para el proceso anammox en 1,2 g N-NO2-/g N-NH4+. A pesar de que, en el caso de la operación con agua real, la alimentación del reactor anammox tenía concentraciones de alcalinidad bajas (2 ± 1 mg CI/L) y valores de pH alreador de 6,2, estos no afectaron a la estabilidad del proceso. Por otra parte, la presencia de concentraciones residuales de materia orgánica del agua residual municipal tras el tratamiento primario contribuyó a un aumento de la eficiencia de eliminación de nitrógeno desde valores alrededor del 80 % a 90 %, alcanzando concentraciones de NT en el efluente inferiores a 10 mg NT/L (límite de vertido en la Unión Europea en las áreas sensibles para grandes depuradoras). La biomasa anammox se retuvo satisfactoriamente en el sistema y su actividad específica aumentó a lo largo de la operación desde 53 ± 11 mg N/(g SSV·d), en el inóculo, hasta valores máximos de 78 ± 8 mg N/(g SSV·d) al final de la operación con agua sintética. Sin embargo, durante el período de operación con agua municipal se observó un descenso de la actividad a valores medios de 58 ± 6 mg N/(g SSV·d), probablemente debido al desarrollo de bacterias heterótrofas desnitrificantes. Considerando la actividad específica anammox y la concentración de biomasa en el reactor, el sistema tenía una capacidad de eliminación potencial mayor a la VEN Resumen 24 tratada (40 mg NT/(L·d) indicando que ésta estaba limitada por la carga aplicada. Finalmente, se estudió el efecto del pH (6 - 8), temperatura (15 - 30 °C) y concentración de materia orgánica (0 - 75 mg COT/L) sobre la actividad específica anammox en ensayos en discontinuo. Los resultados indican que el pH es el parámetro ensayado con mayor efecto sobre la actividad y que el rango de pH óptimo varía con la temperatura. Capítulo 8. Comportamiento de los procesos de nitritación parcial-anammox en una etapa en un sistema IFAS En el Capítulo 8 se presenta el reactor IFAS como una configuración prometedora para llevar a cabo los procesos de NP/AMX en una etapa debido a su capacidad de segregar las diferentes poblaciones bacterianas en distintas fracciones de biomasa (biomasa en suspensión y biopelícula). El reactor IFAS a escala piloto (200 L) se alimentó continuamente con agua residual municipal pretratada anaeróbicamente y se operó a temperaturas decrecientes desde 21 a 15 °C. A pesar de que la alimentación contenía una DQO relativamente alta con relaciones DQO/N de 2.5 ± 0.3 g DQO/g N, los balances de materia indican que el proceso anammox es el mayor responsable de la eliminación de nitrógeno obtenida. Durante la operación se obtuvieron eficiencias de eliminación de nitrógeno de 72 ± 11 % y valores medios de VEN de 37 ± 3 mg NT/(L·d) a 15 °C. Con respecto a la distribución de las poblaciones principales, se observó una segregación siendo las BOA más abundantes en la biomasa en suspensión mientras que las bacterias anammox se encontraban principalmente en la biopelícula formada sobre los soportes. Sin embargo, la abundancia de las BON era similar en ambas fracciones de biomasa y, a pesar de que su actividad específica disminuyó durante la operación del reactor, no fueron eliminadas por completo del sistema. Adicionalmente, el patrón de aireación intermitente empleado en el reactor IFAS permitió la limitación de la actividad de las BON dentro del sistema observándose una actividad en el reactor de en torno al 10 - 20 % de la actividad máxima. Con respecto a la calidad del efluente, se obtuvieron concentraciones de nitrógeno de 12 ± 5 mg NT/L, lo que demuestra que se requiere una mayor optimización del sistema para cumplir con los límites de vertido establecidos por la legislación europea para las áreas sensibles: 15 mg NT/L para las EDAR que tratan cargas de entre 10.000 y 100.000 habitantes equivalentes y 10 mg NT/L para las de mayor tamaño. Resumen 25 Capítulo 9. Implementación de un sistema de nitritación parcial y anammox en dos etapas a escala piloto Los prometedores resultados obtenidos a escala laboratorio con la configuración de NP/AMX en dos etapas condujeron a probar esta configuración por primera vez a escala piloto (600 L cada unidad) como se describe en el Capítulo 9. En este caso, la planta piloto se instaló en una EDAR municipal donde operaba un reactor aerobio de lodos activos de alta carga para la eliminación de materia orgánica (sin afectar a la concentración de nitrógeno). El sistema NP/AMX se operó sin control de temperatura (variando entre 11 y 28 °C) y haciendo frente a las mismas fluctuaciones en las características del agua residual a las que estaba sometida la EDAR municipal. El efluente del reactor aerobio de alta carga se caracterizaba por tener concentraciones de amonio entre 22 y 63 mg N-NH4+/L, que se oxidó parcialmente a nitrito en la unidad de nitritación parcial. Durante toda la operación no se observó actividad de las BON que estaban inhibidas mediante la acumulación de ANL in situ (0,015 - 0,200 mg N-HNO2/L). Para obtener una relación nitrito/amonio en el efluente de 1,2 g N-NO2-/g N-NH4+, adecuada para alimentar al reactor anammox en serie, se implementó una estrategia de control para terminar la etapa de aireación basándose en el valor de pH en el interior del rector de nitritación. De este modo, el reactor pudo soportar mejor las fluctuaciones en las características de las aguas residuales. Por otra parte, en el reactor anammox se alcanzó una eficiencia de eliminación de nitrógeno del 80 % con una VEN de 99 ± 26 mg NT/(L·d), limitada principalmente por el TRH impuesto previamente en la unidad de NP para obtener una adecuada retención de biomasa. De hecho, la principal limitación encontrada en esta experiencia fue la retención de biomasa en esta unidad. Con respecto a la calidad del efluente, se obtuvo una concentración de 12 ± 3 mg NT/L. Sin embargo, es necesario señalar que el efluente del reactor aerobio de alta carga de la EDAR alimentado al sistema NP/AMX ya contenía aproximadamente 5 mg N-NO3-/L de nitrato. El nitrato no es eliminado en el sistema NP/AMX, limitando por tanto su eficacia total de eliminación de nitrógeno. La optimización del sistema aerobio de alta carga de la EDAR, mediante la reducción del tiempo de retención celular para limitar el proceso de nitrificación en esta unidad, fomentaría la obtención de un efluente de mayor calidad cumpliendo con los límites de vertido establecidos comúnmente en la legislación. Finalmente, se realizó un estudio de los potenciales ahorros comparando la combinación del sistema aerobio de alta carga y la Resumen 26 eliminación de nitrógeno mediante NP/AMX con un sistema convencional de lodos activos (la tecnología más comúnmente empleada para la eliminación de nitrógeno en las EDAR). Los resultados demuestran que la implementación de los procesos de eliminación autótrofa de nitrógeno permite duplicar la recuperación de energía de la planta ya que más materia orgánica es tratada en el digestor anaerobio de lodos produciendo más biogás. Además, la eficiencia energética de la planta aumenta del 21 % con el sistema convencional de lodos activos al 57 % con la aplicación de los procesos de NP/AMX, y los costes operacionales también se reducen a la mitad. Sin embargo, la producción de lodos que deben ser gestionados como residuo es menor en el escenario convencional ya que el reactor aerobio de alta carga tiene una producción elevada de lodos incrementando los costos asociados a su gestión. Capítulo 10. Conclusiones generales y futuros desafíos para la eliminación autótrofa de nitrógeno en la línea principal de aguas En el capítulo 10 se presenta una discusión general con el objetivo de integrar los principales resultados y conclusiones más relevantes obtenidos durante la realización de esta tesis. En general, los resultados de esta tesis contribuyen a incrementar la comprensión del proceso de NP/AMX y proporcionan información interesante sobre la viabilidad de aplicar los procesos de NP/AMX para eliminar autotróficamente el nitrógeno a baja temperatura. Así se intenta responder a algunas de las principales preguntas que permanecen abiertas en este campo: i) ¿Es posible aplicar los sistemas basados en el proceso anammox en sistemas de tratamiento de aguas descentralizado? ii) ¿Es posible suprimir la actividad de las BON en condiciones de la línea principal de agua? iii) ¿Son capaces las bacterias anammox de tratar cargas suficientemente altas de forma estable para la eliminación de nitrógeno en la línea de aguas? iv) ¿Cómo afecta la composición del agua residual municipal al comportamiento de los procesos de NP/AMX? v) ¿Cuál es la configuración de reactor más eficaz para la implementación de los procesos de NP/AMX para el tratamiento de aguas domésticas? Se demostró que el tratamiento de las aguas negras mediante un sistema NP/AMX en una etapa es viable, probablemente porque su composición es Resumen 27 relativamente más alta y estable en comparación con las aguas residuales municipales, una ventaja para la estabilidad del proceso de NP/AMX. Los sistemas descentralizados, especialmente los de pequeña escala, podrían enfrentarse a períodos de ausencia de aguas residuales para tratar, pero se ha comprobado que las paradas que se ocasionarían en el sistema de tratamiento no impiden la aplicación de los procesos de NP/AMX. Así, se podría incrementar el número de pequeñas instalaciones con sistemas de eliminación de nitrógeno para promover su reutilización. Por otra parte, el sistema IFAS permite segregar las diferentes poblaciones microbianas en las fracciones de biomasa en función de su velocidad de crecimiento. Las BON podrían eliminarse del sistema mediante un control adecuado del tiempo de retención celular incrementado la eficacia de eliminación de nitrógeno. Además, se han obtenido resultados especialmente relevantes y novedosos en el marco de la aplicación de los procesos de NP/AMX en dos etapas. Primero se propuso una nueva estrategia operativa basada en la acumulación de ANL in situ a escala laboratorio y tratando agua residual sintética. Una vez entendido el mecanismo de inhibición de las BON, se validó la estrategia tratando agua residual urbana con presencia de materia orgánica. Luego, se eliminó el nitrógeno en reactores anammox en los que se maximiza la eliminación autótrofa de nitrógeno al ser un reactor exclusivamente anóxico con ausencia de materia orgánica, porque es oxidada en la unidad NP previa. Se obtuvieron eficacias de eliminación entre 80 y 90 % obteniendo efluentes de alta calidad. Así, el sistema de NP/AMX en dos etapas permite alcanzar eficacias de eliminación obteniendo efluentes de alta calidad con estrategias de operación simples que no requieren sistemas de control complejos lo que favorece su implementación. Por último, en este capítulo se discuten las principales limitaciones para la implementación de los procesos de NP/AMX en la línea principal de las EDAR a baja temperatura prestando especial atención a las propuestas de trabajo futuro. Chapter 1 Introduction SUMMARY Efficient nitrogen removal from sewage is crucial since wastewater treatment plants (WWTPs) are facing more and more stringent effluent quality requirements. Nitrogen is conventionally removed via biological nitrification-denitrification processes. These processes are large energy-consumers and organic matterdemanding as aeration and organic carbon need to be provided for nitrification and denitrification, respectively. For these reasons, new technologies are under development and study to improve nitrogen removal from wastewater without compromising the WWTP energy efficiency. Among them the nitritationdenitritation, partial nitritation-anammox (PN/AMX) or those processes based on microorganisms that use methane as electron donor stand out. In particular, the implementation of the PN/AMX processes in the mainstream is expected to transform the WWTP into an energy producer facility instead of the current energy sink, while the produced effluent fulfils the nitrogen discharge limits fixed by the European Union. In the present chapter, the scope and motivations of the thesis are thoroughly explained and the existent alternatives to biologically remove nitrogen from municipal wastewater analysed. Special attention is paid to the application of the PN/AMX processes for the treatment of low-strength wastewater at low temperature. Finally, the main objectives pursued in this thesis are presented. Chapter 1 30 OUTLINE 1.1.2.1. Reclaimed wastewater reuse 34 1.1.2.2. Energy recovery 39 1.1.2.3. Nutrients recovery 40 1.1.3.1. Conventional activated sludge systems 44 1.1.4.1. Organic matter removal (A-Stage) 45 1.1.4.2. Nitrogen removal (B-Stage) 47 1.1.4.3. Decentralised systems 47 1.2.5.1. Methane autotrophic denitrification 54 1.2.5.2. Sulphur autotrophic denitrification 55 1.4.1.1. Aerobic sludge retention time 62 1.4.1.2. Dissolved oxygen concentration 62 1.4.1.3. Alternating anoxic and aerobic conditions 63 1.4.1.4. Aeration time control 63 1.1. Shift from wastewater treatment to water reuse and resource recovery 32 1.1.1. Water scarcity 32 1.1.2. Domestic wastewater reuse as a pool of resources 34 1.1.3. Nitrogen removal in municipal wastewater treatment: current status 41 1.1.4. Innovative wastewater treatment plants 44 1.2. Nitrogen cycle and opportunities for municipal wastewater treatment 48 1.2.1. Nitrification (nitritation and nitratation) 49 1.2.2. Complete ammonia oxidising (comammox) bacteria 50 1.2.3. Heterotrophic denitrification (denitratation and denitritation) 51 1.2.4. Anaerobic ammonia oxidation: anammox 52 1.2.5. Other autotrophic denitrification processes 54 1.3. Towards efficient wastewater treatment: nutrient removal alternatives 57 1.3.1. Partial nitritation and anammox processes application in WWTPs 59 1.4. Application of PN/AMX processes at mainstream conditions 61 1.4.1. Factors affecting NOB suppression 61 Introduction 37 Table 1.1. Example of reference documents and regulations in force of water reuse worldwide and corresponding established nitrogen limits (elaborated from Paranychianakis et al. (2015) and cited legal documents). Territory Uses Agriculture Urban Industrial Environmental Groundwater recharge (non potable) Indirect potable reuse USA [1]* N.S. N.S. N.S. N.S. N.S. N.S. California [1] 10 mg N/L Florida [1] Soil balance N.S N.S 6 mg N/L, 4 mg NH4/L < 12 mg N/L 10 mg N/L Arizona [1] 10 mg N/L 10 mg N/L N.S. N.R. N.S. 10 mg N/L North Carolina [1] 6 mg NH3-N/L N.S. N.S. N.S N.R. N.R. Australia [2]* Soil balance N.S. N.S. N.S. 0.5 mg NH3/L, 50 mg NO3/L and 3 mg NO2/L Israel [3] < 25 mg N/L and 10 mg NH3-N/L European Union [4] N.S. N.R. N.R. N.R. N.S. N.R. Italy [5] 35 mg N/L N.R. N.R. N.R. 15 mg N/L; 2 mg NH4+/L N.R. Cyprus [6] 15 mg N/L Portugal [7] N.S. N.R. N.R. N.R. N.R. N.R. Spain [8] 10 mg N/L N.S. N.S. 10 mg N/L and 25 mg N/L France [9] N.S. N.R. N.R. N.R. N.R. N.R. Greece [6] 30 mg N/L 15 mg N/L, 2 mg NH4+/L N.R. N.S.: not specified values for nitrogen; N.R.: non-regulated use; USA: United States of America. *National reference document for reuse guidance but not legally binding as actual regulation occurs at state level. References: [1] EPA/600/R-12/618; [2] NRMMC-EPHC, 2006 and 2008; [3] Israel Ministry of Environmental Protection (2010); [4] P8_TA(2019)0071; [5] D152/2006; [6] Paranychianakis et al. (2015); [7] NP 4434/2005 and ReciRAR 2/2007; [8] RD 1620/2007; [9] JORF 0201/31.08.10/n 34 page 15828. Chapter 1 38 The Nitrates Directive aims to protect water sources which are or may become exposed to nitrogen pollution from agricultural activities. This directive considers the treated wastewater containing nitrogen as fertiliser and limits its disposal in the socalled nitrate vulnerable zones (a specific type of eutrophic sensitive area) that are those at risk of exceeding 50 mg NO3-/L (Figure 1.4). Moreover, 16 member States apply a more stringent treatment than the secondary one in their whole territory which is considered a sensitive area (article 3.5 of UWTD). In total sensitive areas amount to 75 % of the entire EU territory (Figure 1.4). A) B) Figure 1.4. A) Nitrate vulnerable zones (■) and areas applying the article 3.5 of the Urban Wastewater Treatment Directive (i.e. more stringent treatment in the whole territory) (■). B) Percentage of monitoring points with annual average nitrate concentrations above 25 mg/L: ■ 75 - 100 %, ■ 50 - 75 %, ■ 25 - 50 %, ■ 15 - 25 %, ■ 5 - 15 % and ■ 0 - 5 %. Reprinted from Water portal, European Commission. Despite the benefits of recycling water, there are several concerns related to its associated environmental and health risks. If it is not properly managed, irrigation induces nitrate run-off and ultimate soil acidification. Moreover, the presence of inorganic compounds such as metals or salts might limit the acceptability of reclaimed water for several reuse applications like irrigation, industrial uses as salts might damage the machinery or groundwater recharge varying the composition of the receiving water body (Rahman et al. 2016, UN-Water, 2018). The main problem due to salinity, when the desired use is irrigation, is the loss of soil structure accelerating the salinisation within the root zone. According to the FAO, 20 % of the Introduction 39 irrigated land is salt affected, mainly located in the USA, Australia, China, India, Argentina and Central Asia (UN-Water, 2018). Over the last decade, salinisation of water bodies increased due to the leaching of salts accumulated in solids, production of brackish drainage, industrial activity, intrusion of seawater in aquifers (due to the excessive water extraction) and the salts added by urban uses. Different tertiary treatment steps, including desalinisation and reverse osmosis, are required to enable the water reuse or discharge. Other concern is the presence of complex toxic compounds (more frequent if wastewater from some industries is collected), heavy metals, or substances of emerging concern that are not removed in the WWTPs and therefore, they could be accumulated in reclaimed water (UN-Water, 2018). 1.1.2.2. Energy recovery Wastewater plays a significant role in the water-energy nexus. Water cycle needs approximately 4 % of the total energy consumed worldwide (UN-Water, 2018). Wastewater is a source of chemical, thermal and hydraulic energy that can be recovered as biogas, heating and electricity. In developed countries, 42 % of electricity consumption in the water sector is used for wastewater treatment. By 2040 it is estimated that water collection and treatment will require 60 % more electricity. If energy is properly valorised, wastewater treatment could produce 30 - 100 TWh by 2040 (UN-Water, 2018) converting the WWTPs into net energyproducing facilities (Garrido et al. 2013). Among the possible options, the chemical energy recovery as biogas is the most popular alternative since organic matter removal takes place simultaneously. Anaerobic wastewater digestion is only applied in temperate climates, but in many cases, the produced biogas is not recovered as energy but released to the environment (Chernicharo et al. 2015, Noyola et al. 2012, Stazi and Tomei 2018). In cold climates, the conventional activated sludge (CAS) is the most widely applied technology. In these WWTPs primary and secondary sludge are anaerobically digested to produce biogas. Nevertheless, energy recovery is limited to 10 - 50 % (Burton et al. 2014, Longo et al. 2016). For this reason, sewage sludge pretreatments are applied to improve anaerobic biodegradability such as the thermal hydrolysis (TH), mechanical disintegration, application of acidic or alkaline reagents (Carrère et Chapter 1 40 al. 2010, Zhen et al. 2017). The TH is a well-established technology with commercial solutions (e.g., CambiTHPTM and Biothelys®) applied worldwide to enhance the methane yield from approximately 30 to 80 % depending on the operating conditions (Nordin et al. 2018, Sapkaite et al. 2017, Zhen et al. 2017). Another alternative to further increase energy generation in WWTPs is to use co-digestion that consists in mixing the sewage sludge with other organic-rich wastes (e.g., animal manure, food-processing waste, grease, microalgae) increasing the biogas production (Elalami et al. 2019, Guven et al. 2019). The addition of new wastes must be appropriately assessed considering its composition since besides organic matter other substances such as nitrogen would be supplied too (RodriguezVerde et al. 2018, Taboada-Santos et al. 2019a). Moreover, the digested solids might be also valorised as a soil conditioner and fertiliser for agriculture or further treated off-site in centralised plants through incineration for energy recovery (Guven et al. 2019, UN-Water, 2018). Innovative technologies such as microbial fuel cell (MFC) or microbial electrolysis cell (MEC) allow converting the wastewater organic matter content either directly into electrical energy or indirectly through hydrogen gas production, respectively. Combined with anaerobic digestion (AD) or constructed wetlands they are also expected to increase their efficiency. These technologies are being tested at pilot-scale and face limitations related to large investment costs or upscale technical issues (Guadarrama-Pérez et al. 2019, Katuri et al. 2019, Pham et al. 2006, Zhen et al. 2017). Other alternatives like the use of algae or phototrophic purple bacteria (PPB) are also under consideration. The phototrophic systems use energy from sunlight reducing the chemical oxygen demand (COD) required for biological nitrogen removal. Moreover, PPB assimilates approximately 16 g COD/g N and produces hydrogen from organic matter or other value-added products such as polyhydroxyalkanoates or proteins (Vasiliadou et al. 2018). Algae are used to harvest the organic matter and nutrients on their biomass obtaining a concentrated stream. Algae biomass is then digested producing biogas. 1.1.2.3. Nutrients recovery One of the main identified problems for resource recovery from domestic wastewater is the fact that it is a diluted stream which involves the treatment of high-water volumes with low concentrations of valorising compounds (Liu et al. Introduction 41 2018b, Winkler and Straka 2019). Thus, diluted wastewater streams containing ammonium and phosphorus concentrations below 100 mg NH4+-N/L and 40 mg P/L, respectively, are preferable submitted to biological nutrient removal treatments for their better cost-effectiveness compared to resource recovery processes (Mulder 2003). Although there are other alternatives, for example, P chemical precipitation is technically feasible and produces a phosphorus-rich sludge that might be used as a soil amendment (Zhou et al. 2018). However, the N recovery technologies are not competitive with those used in the chemical industry for nitrogen-based fertiliser production (Winkler and Straka 2019). Among the proposed technologies, other examples are the promising results obtained by MEC or the adsorption of ammonium and phosphorus onto zeolite surface obtaining a solid that might be used as soil fertiliser (Winkler and Straka 2019, You et al. 2019). 1.1.3. Nitrogen removal in municipal wastewater treatment: current status Wastewater treatment level and regulations concerning N compounds highly vary within countries (Figure 1.5). With this in mind, the UN analysed 275 national standards from 100 countries and almost 80 % established discharge limits for total nitrogen (TN). Nitrogen discharge limits varied from 80 mg TN/L (mainly in Latin America) to values as low as 2 mg NH4+-N/L allowed in Switzerland. Nevertheless, in African or Asian countries standards are often not accomplished (UN-Water, 2018). Figure 1.5. Total nitrogen (TN) standard limits for treated wastewater discharge in different countries: limit lower than 15 mg TN/L (■), regulated but at values between 20 and 80 mg TN/L (■), no regulated (■) and no data (■). Compilation based on the revision of country policies. Chapter 1 42 In Latin America, by 2015, less than 60 % of the population was connected to sewer systems and nearly all urban wastewater was discharged without any treatment (UN-Water, 2018). The principal applied treatment technologies are the stabilisation pounds (80 % in number of facilities and treated flow), CAS and up-flow anaerobic sludge blanket (UASB) reactors (Noyola et al. 2012). Nowadays, in Chile, Brazil, Mexico and Uruguay at least 50 % of the used water is treated being wastewater treatment less likely in other countries (UN-Water, 2018). Nevertheless, regulations established high TN concentrations limits (40 - 80 mg TN/L) to be reached before discharge into the environment (Figure 1.5). In the EU, an average 83 % of the total wastewater is treated in WWTPs whereas 5 % is treated in individual and other appropriate systems (IAS) and the remaining 12 % is not treated (European Environment Agency, 2017). In small agglomerations (< 2,000 p.e., population equivalent) 34 % of the water remains untreated being clear that this is a niche for decentralised wastewater treatment systems. The big agglomerations (> 100,000 p.e) account for only 2 % of the total municipal WWTPs but these facilities treat 51 % of the total load (as p.e.) (Figure 1.6). Figure 1.6. Distribution of the percentage of wastewater treated in municipal WWTPs (■), individual and other appropriate systems (IAS, ■) and untreated (■) according to the agglomeration population size in EU. Elaborated with data from European Environment Agency (2017). The wastewater treatment level depends on the size of the WWTPs, the larger the WWTPs the more intensive the applied treatment (Figure 1.6 and Figure 1.7.A). The UWTD regulated the wastewater treatment of agglomerations from 2,000 p.e or more that should be subjected wastewater to secondary treatment. This directive established TN discharge limits to sensitive areas for WWTPs with capacities of 81 % 6 % 13 % 2,000 - 10,000 p.e 57 % number plants 11 % total load 62 % 3 % 34 % < 2,000 p.e 12 % number plants 1 % total load 95 % 3 % 2 % 10,000 -100,000 p.e 29 % number plants 37% total load 98 % 2 % 1 % > 100,000 p.e 2 % number plants 51% total load Introduction 43 treatment over 10,000 p.e.: between 10,000 and 100,000 p.e. the discharge standard is 15 mg TN/L while it is reduced to 10 mg TN/L for loads above 100,000 p.e. Alternatively a reduction of the 70 - 80 % of the applied nitrogen load might be requested. A) B) Figure 1.7. Wastewater maximum treatment level: primary treatment, secondary treatment, phosphorus (P) removal, nitrogen (N) removal, N+P removal and N+P removal plus more stringent treatment. A) Depending on the wastewater treatment plant size in population equivalent (p.e.) and B) In different countries of which the names are coded according to the ISO 3166-1. Note that to obtain the total percentage of nitrogen removal the three last categories are considered. Elaborated from data from European Environment Agency (2017). Regarding nutrients, N is only removed in 30 % of the small WWTPs while 97 % of the biggest ones already remove it. It is worth to note that advanced treatments (e.g., ozonation, ultraviolet) were considered combined with nutrients removal in Figure 1.7, but the use of tertiary treatments occurs in other WWTPs with lighter 0 10 20 30 40 50 60 70 80 90 100 < 2,000 2,000 - 10,000 10,000 - 100,000 >100,000 Treatment level (%) Plant load (p.e) 0 10 20 30 40 50 60 70 80 90 100 AT BE BG CH CY CZ DE EE ES FI FR GB GR HR HU IE IS IT LT LU LV MT NL NO PL PT RO SE SI SK Treatment level (%) Country Primary Secondary P removal N removal NP removal NP+Other Chapter 1 44 treatment. Moreover, the treatment level differs also depending on the geographical location and while N is removed in the 98 % of the WWTP in The Netherlands and Latvia, its removal amounts to less than 10 % in countries such as Iceland, Norway and United Kingdom (Figure 1.7.B). Thus, improvements for the treatment level applied are still required. In addition, effluent discharge limits worldwide are expected to become more stringent and will drive future innovations due to the increasing relevance of indirect water reuse (UN-Water, 2018). For this reason, most of the recent studies focus on the development of new technologies producing effluents with improved quality, which help to increase the economic and environmental sustainability of WWTPs where they are implemented. 1.1.3.1. Conventional activated sludge systems In CAS systems the COD contained in wastewater is converted to sludge and CO2 in a similar approximate ratio. For removing nitrogen, CAS systems are operated at long sludge retention times (SRTs), diminishing the sludge biodegradability. When the nitrification-denitrification processes take place the energy and resource demand, together with the operational costs and WWTP carbon footprint, augmented (Siegrist et al. 2008). The high energy consumption for aeration and the excessive produced waste activated sludge, due to the heterotrophic bacteria growth, are the main drawbacks of the CAS systems. Indeed, the aeration in the biological reactors and the sludge handling units (dewatering, thickening…) are the major energy consumers using 50 - 60 % and 15 - 25 % of the total energy consumption, respectively (Liu et al. 2018b, Wang et al. 2019). Thus, new alternatives are under study to improve the WWTPs from an economical and environmental sustainable point of view. 1.1.4. Innovative wastewater treatment plants The simultaneous increase of the energy recovery and the reduction of aeration energy consumption in large WWTPs is faced nowadays by augmenting the COD directed to AD. With this objective, a stage for organic matter removal (A-stage) followed by another stage for N removal (B-stage) are proposed to replace the CAS systems. Introduction 45 When dealing with small agglomerations decentralised wastewater treatment systems become an attractive alternative to be applied diminishing the pumping costs, and enabling the energy and nutrient recovery (WWAP, 2017) since raining water is not collected (or in a lesser extent) (Verstraete and Vlaeminck 2011). 1.1.4.1. Organic matter removal (A-Stage) The energy contained in wastewater can be recovered either by direct anaerobic wastewater digestion (Chernicharo et al. 2015) or by pre-concentrating the COD in the sludge, facilitating the energy recovery in anaerobic sludge digesters, requiring less energy for aeration and reducing the sludge production (Guven et al. 2019). COD capture from municipal wastewater is possible by applying chemically enhanced primary treatment (CEPT) or high rate activated sludge (HRAS) processes. In the A-Stage, only 10 to 30 % of the N is removed by microbial growth (De Graaff et al. 2016, Guven et al. 2017, He et al. 2016a) and it varies with the influent biological oxygen demand (BOD5) to nitrogen ratio(BOD5/N), SRT and the amount of N in recycling streams from solids dewatering after anaerobic digestion (Burton et al. 2014). • High Rate Activated Sludge The aerobic HRAS process consists in applying high loading rates of 2 g BOD5/(g VSS·d) at short SRT values of 0.2 - 4.0 days and short hydraulic retention time (HRT) of 0.5 - 4.0 h. In the HRAS process, COD is adsorbed onto the activated sludge flocs resulting in COD removal efficiencies of 50 - 70 %, minimising its conversion to CO2 and consequently reducing the energy consumption for aeration (Jimenez et al. 2015, Meerburg et al. 2015). The SRT and the concentration of the COD fractions are the main parameters determining COD biosorption. Besides soluble COD (sCOD), particulate COD (pCOD) is also removed and reached efficiency increases with the lengthening of SRT (Jimenez et al. 2015). In HRAS, a COD-rich sludge is produced with a volatile fraction of 78 %, higher than the one from the sludge generated in the CAS systems (58 %). Thus, methane production yield from the former is expected to be larger than that of the latter. Nevertheless, the higher COD diversion to sludge in HRAS systems and the high water content in sludge may complicate its dewatering process (Cagnetta et al. 2019). Chapter 1 46 • Chemically enhanced primary treatment During the CEPT coagulants are added to enhance the removal efficiency of primary settling tanks from 50 % of total suspended solids (TSS) and 30 % of pCOD to up to 90 % TSS and 75 % COD (Guven et al. 2019). Besides pCOD, up to 55 % of sCOD can also be removed (Guven et al. 2019, Taboada-Santos et al. 2019b). Additionally, when aluminium or ferric salts are used, up to 80 % of the phosphorus is chemically removed (Burton et al. 2014, Guven et al. 2019). CEPT is characterised by low energy requirements, simple operation and maintenance, high treatment capacity and removal efficiencies, and the ability to cope with overloads better than CAS systems (Murugesan et al. 2014). The used salts exert no adverse effect on methane production during anaerobic sludge digestion but provoke the increase in sludge production and consequently in the operational costs (Guven et al. 2019). An additional issue is the alkalinity depletion (Equation 1.1). Large amounts of coagulant dosages may result in acidic pH, which can damage downstream processes (Burton et al. 2014) limiting, for example, the nitrification process efficiency. 2 FeCl3+3 Ca(HCO3)2 ⇄2 Fe(OH)3+3 CaCl2+6 CO2 Eq. 1.1 • Anaerobic digestion (AD) of wastewater Direct anaerobic treatment of diluted municipal wastewater leads to removal efficiencies ranging from 65 to 80 % of COD and 75 to 95 % of BOD5 at HRT values of 6 - 10 hours (Chernicharo et al. 2015). Anaerobic treatments are more efficient than aerobic ones when treating municipal wastewater characterised by medium to high organic matter content (> 1 g COD/L). The effluent quality from anaerobic wastewater digesters is worse than those from aerobic treatments, in terms of COD and suspended solids (Stazi and Tomei 2018). Nevertheless, compactness of the technologies, minimal energy requirements and low sludge production make AD attractive and competitive in climate regions with temperatures above 20 °C (Chernicharo et al. 2015). AD feasibility to treat low strength wastewater at psychrophilic conditions is still under evaluation (Arias et al. 2018, Silva-Teira et al. 2017, Stazi and Tomei 2018, Zhou et al. 2018). Besides methane (CH4) and CO2, mineralised compounds such as ammonium, phosphate and hydrogen sulphide (H2S) are produced. A primary concern towards its applicability is the potential stripping of H2S and CH4 dissolved in the anaerobic digester effluent contributing to bad Introduction 53 1.2.4. Anaerobic ammonia oxidation: anammox The anaerobic ammonium oxidation (anammox; AMX) bacteria are chemolithoautotrophic microorganisms which anoxically oxidise ammonium, using nitrite as electron acceptor, to nitrogen gas, producing residual amounts of nitrate in the anabolism. The anammox process was thermodynamically predicted by Broda (1977) but it was in 1990s when Mulder et al. (1995) experimentally proved its existence in a denitrifying pilot plant. Anammox stoichiometry was first proposed by Strous et al. (1999) (Equation 1.9) and recalculated later by Lotti et al. (2014a) (Equation 1.10). 𝑁𝐻4 ++1.32 𝑁𝑂2 −+0.066 𝐻𝐶𝑂3 −+0.13 𝐻+ →1.02 𝑁2+0.26 𝑁𝑂3 −+0.066 𝐶𝐻2𝑂0.5𝑁0.15 +2.03 𝐻2𝑂 Eq. 1.9 𝑁𝐻4 ++1.146 𝑁𝑂2 −+0.071 𝐻𝐶𝑂3 −+0.057 𝐻+ →0.986 𝑁2+0.161 𝑁𝑂3 −+0.071 𝐶𝐻1.74𝑂0.31𝑁0.20 +2.002 𝐻2𝑂 Eq. 1.10 The amount of nitrate produced represents 11.2 % (Strous et al. 1999) or 7.5 % (Lotti et al. 2014a) of the TN converted depending on the stoichiometry used, limiting the NRE of the system. The anammox process has a low impact on wastewater alkalinity and the low biomass yield, ranging from 0.07 to 0.13 g VSS/g NH4+-N, is an advantage from the sludge handling point of view but it requires to operate the anammox systems at long SRT values and long start-up periods. One of the main advantages of the anammox process is that COD is not needed for N removal. Furthermore, hydrazine (N2H4) and nitric oxide (NO) are metabolic intermediates of the process but N2O is not produced potentially decreasing the GHGs emissions originated in WWTPs (Campos et al. 2016, Castro-Barros et al. 2015, Kampschreur et al. 2008). Anammox bacteria can reduce nitrate first to nitrite and then to ammonia using formate, acetate or propionate as electron donors (Kartal et al. 2007). However, the presence of COD usually promotes the heterotrophic denitrifying bacteria overgrowth (Castro-Barros et al. 2017, Xu et al. 2015). Anammox bacteria are ubiquitous in WWTPs and natural environments such as marine sediments. Anammox bacteria belong to the phylum Planctomycetes and overall ten anammox species have been described within five different genera: Candidatus “Kuenenia” (represented by K. stuttgartiensis sp.), Candidatus “Brocadia” (with the species B. anammoxidans, B. fulgida and B. sinica), Candidatus Chapter 1 54 “Anammoxoglobus” (A. propionicus sp.), Candidatus “Jettenia” (J. asiatica sp.), Candidatus “Scalindua” (with the species S. brodae, S. sorokinii, S. wogneri and S. profunda). All of them were enriched from WWTP sludge samples except C. “Scalindua” that comes from marine sediments. C. “Kuenenia” and C. “Brocadia” are the most commonly found genera in enrichments from WWTPs. The average optimal temperature and pH values have been found to be 35 °C and 8, respectively. The anammox bacteria are characterised by low productivity with typical doubling times as long as 11 days (Strous et al. 2002). More recently, a faster-growing anammox culture with doubling times of 3 days was also obtained at the optimal conditions (Lotti et al. 2015a) but in practical applications, they are still considered as slowgrowing microorganism. One of the most relevant features of anammox bacteria is its high affinity for both substrates, ammonium and nitrite, being the affinity coefficients under 0.10 mg N/L (Strous et al. 1999). On the contrary, anammox bacteria sensitiveness is widely reported (Jin et al. 2012) being affected by numerous compounds such as DO concentration (Seuntjens et al. 2018), COD (Giustinianovich et al. 2016), nitrite (or its protonated form, free nitrous acid (FNA)) (Fernández et al. 2012, Lotti et al. 2012, Puyol et al. 2014b), ammonium concentration (or it unionized form free ammonia (FA)) (Fernández et al. 2012, Puyol et al. 2014a), pH (Daverey et al. 2015), temperature (Dosta et al. 2008, Lotti et al. 2015c) or salts (Dapena-Mora et al. 2007, Fajardo et al. 2014). Considerable differences were found in the inhibition thresholds probably due to the anammox different enrichment degree of the performed studies. Thus, an assessment of the wastewater composition is recommended before the anammox implementation. 1.2.5. Other autotrophic denitrification processes Other appealing autotrophic denitrification processes might occur using the CH4 or H2S (produced during the AD) as alternative inexpensive electron donors (Chernicharo et al. 2015). 1.2.5.1. Methane autotrophic denitrification Methane oxidation linked to denitrification can occur either through aerobic or anaerobic processes. In the aerobic methane oxidation coupled to denitrification (AMO-D) process, aerobic methanotrophs oxidise CH4 into products such as Introduction 55 methanol, acetate and carbohydrates, that can be used by the heterotrophic denitrifiers to reduce nitrite (Equation 1.11) and/or nitrate (Equation 1.12) to dinitrogen gas (Burton et al. 2014). Methanotrophs are strictly aerobic bacteria and use IC for microbial growth whereas CH4 is used as carbon and energy source. 3 𝐶𝐻4+ 3 𝑂2+ 4 𝑁𝑂2 − + 4 𝐻+→2 𝑁2+3 𝐶𝑂2+8 𝐻2𝑂 Eq. 1.11 5 𝐶𝐻4+5 𝑂2+ 4 𝑁𝑂3 − + 4 𝐻+→2 𝑁2+ 5 𝐶𝑂2+5 𝐻2𝑂 Eq. 1.12 In anoxic conditions, Raghoebarsing et al. (2006) reported the first enriched culture of nitrite/nitrate dependent anaerobic methane oxidation (N-damo) microorganisms and since then, it became a hot topic in wastewater treatment research. During the N-damo process, CH4 is oxidised to CO2 coupled to both nitrate (performed by archaea, Equation 1.13) and nitrite (performed by bacteria, Equation 1.14) reduction and using IC for biosynthesis. The most studied N-damo archaea is Candidatus “Methanoperedens nitroreducens” sp. (Haroon et al. 2013) while Ndamo bacteria belong to the NC10 phylum being C. “Methylomirabilis oxyfera” and C. “Methylomirabilis sinica” the most researched species (Ettwig et al. 2010, He et al. 2016b). 2 𝐶𝐻4+ 8 𝑁𝑂3 −→8 𝑁𝑂2 − + 2 𝐶𝑂2+4 𝐻2𝑂 Eq. 1.13 3 𝐶𝐻4+8 𝑁𝑂2 −+8 𝐻+→4 𝑁2+3 𝐶𝑂2+10 𝐻2𝑂 Eq. 1.14 As N-damo bacteria have long doubling times of up to two weeks, the existence of efficient biomass retention systems is crucial (Ettwig et al. 2009). Moreover, Ndamo bacteria have the advantage of avoiding the production of N2O since this is not an intermediate in their metabolic pathway (Ettwig et al. 2010). The anoxic N-damo processes make more efficient use of the methane for N removal requiring 0.7 g CH4/g NO3--N whereas AMO-D process uses 1.4 g CH4/ g NO3--N and requires energy for aeration. Municipal wastewater has approximately 50 mg TN/L, thus, a concentration of 35.5 mg CH4/L would be required for N removal. However, CH4 concentrations in the anaerobic municipal wastewater digester ranged from 12 to 20 mg CH4/L limiting the NRE. Despite the high interest attracted by N-damo, few is known about them and its prevalence in natural environments. Kampman et al. (2012) incorporated the Ndamo process as an UASB digester system and a nitritation reactor to treat municipal wastewater and managed to remove 38 mg TN/(L·d), although the contribution of Chapter 1 56 the N-damo to the TN removal is unclear. Different co-cultures were obtained, mostly combined with anammox bacteria where relatively high nitrogen removal rates (NRRs) were obtained like: 1,030 mg TN/(L·d) at 35 °C (Xie et al. 2017) and 275 mg TN/(L·d) at 21 °C (Xie et al. 2018) in membrane biofilm reactors or 126 mg TN/(L·d) obtained by Allegue et al. (2018) at 28 °C in a membrane bioreactor. Moreover, in cultures where anammox and N-damo bacteria and archaea coexist, anammox bacteria tend to outcompete the N-damo bacteria that progressively disappear (Hu et al. 2015). This fact also happens under nitrite limiting conditions, since anammox bacteria present affinities for nitrite higher than N-damo microorganisms (van Kessel et al. 2018). Thus, most of the studies ended up with the use of a combined N-damo and anammox culture to achieve an excellent effluent quality (Hu et al. 2015, Liu et al. 2019, Stultiens et al. 2019, van Kessel et al. 2018, Xie et al. 2018). 1.2.5.2. Sulphur autotrophic denitrification The sulphur oxidising bacteria (SOB) reduced the nitrate and/or nitrite to dinitrogen gas using sulphur compounds (S2O32-, S2-, S0, S4O62-, SO32-) as electron donors and IC as carbon source for bacterial biosynthesis. In the case of sulphide, the dissolved form abundances changes according to the pH value varying from hydrogen sulphide (H2S; pka=7.0), bisulphide (HS-; pka=12.9) and sulphide (S2-) ions. The total sulphide concentration depends mainly on the sulphate content of the raw wastewater that varies according to the geographical location (Sánchez-Ramírez et al. 2015) and can reach values up to 87 mg S/L when municipal wastewater is treated (Delgado Vela et al. 2015). This compound was found inhibitory for the biological processes but SOB have high tolerance (Fajardo et al. 2012, Lu et al. 2018). Thus, SOB might improve other biological removal process performances by removing the toxic sulphide concentrations from the media (Cui et al. 2019b, Di Capua et al. 2019). Stoichiometric reactions are complex since different sulphur compounds can be used and produced from these reactions. As an example, Equations 1.15 and 1.16 show the complete oxidation of HSto produce sulphate (SO42-) and in Equations 1.17 and 1.18 the HSis oxidised to elemental sulfur (S0), using in both cases either nitrate or nitrite as electron acceptor (Burton et al. 2014, Di Capua et al. 2019). During this process, intermediate compounds are usually detected and mixed final products are Introduction 57 obtained (Cai et al. 2008, Campos et al. 2019, Di Capua et al. 2019). In fact, different HSoxidation degrees are expected depending on the relative S/N molar ratio (Dolejs et al. 2015, Liu et al. 2017a). It was also reported that under sulphur limiting conditions (S/N ratio < 6.51 g S/g N), nitrate reduction will be only carried out to nitrite instead to N2 (Oh et al. 2000). 5 𝐻𝑆−+8 𝑁𝑂3 −+3 𝐻+→ 5 𝑆𝑂4 2− +4 𝑁2+4 𝐻2𝑂 Eq. 1.15 3 𝐻𝑆−+8 𝑁𝑂2 −+5 𝐻+→ 3 𝑆𝑂4 2− +4 𝑁2+4 𝐻2𝑂 Eq. 1.16 5 𝐻𝑆−+2 𝑁𝑂3 −+7 𝐻+→ 5 𝑆0+ 𝑁2+6 𝐻2𝑂 Eq. 1.17 3 𝐻𝑆−+2 𝑁𝑂2 −+5 𝐻+→ 3 𝑆0+ 𝑁2+4 𝐻2𝑂 Eq. 1.18 SOB populations are widespread (Lu et al. 2018). Some of these bacteria only use sulphur compounds as electron donor such as Thiobacillus denitrificans sp., Thiobacillus thiophilus sp. or Sulfurimonas denitficans sp. whereas others can use either organic matter or sulphur compounds like Paracoccus sp. (with P. denitrificans, P. ferrooxidans and P. pantotrophus species) (Di Capua et al. 2019). Other advantages of this process are the SOB low biomass yield that ranged from 0.15 to 0.6 g VSS/g N (Cui et al. 2019b, Di Capua et al. 2019) and the significant lower N2O production compared with the heterotrophic denitrification process (Campos et al. 2019, Cui et al. 2019b, Fajardo et al. 2014). However, the associated high alkalinity consumption and the possible release of sulphate or other intermediate compounds to the environment difficult its implementation, especially in acidic wastewater streams (Cui et al. 2019b, Di Capua et al. 2019). 1.3. Towards efficient wastewater treatment: nutrient removal alternatives New technologies, to achieve economic and environmentally sustainable wastewater treatment, are developed which minimise energy consumption, are reliable and cheap. Figure 1.9 shows the combination of different biological processes to optimise the nitrogen removal from wastewater. Traditional nitrification-denitrification (N-HDN) is an energy-intensive nitrogen removal pathway mainly due to electricity consumption for aeration (4.57 kg O2/kg N) and organic matter use as electron donor (2.86 kg COD/kg N) that cannot be valorised as biogas. The combined nitritation and denitritation processes (N-HDN via Chapter 1 58 nitrite) reduce by 40 % the required COD and by 25 % the energy for aeration. However, 1.71 g COD/g N is still required for N removal (Figure 1.10). Figure 1.9. Comparison of different nitrogen removal processes combinations involving ammonium oxidising bacteria (AOB), nitrite oxidising bacteria (NOB), heterotrophic denitrifying bacteria (HDN) and anammox (AMX) bacteria. The application of autotrophic N removal process enables more sustainable use of wastewater resources maximising the energy recovery from the wastewater COD and reducing the sludge production (Kartal et al. 2010). The combination of the PN/AMX processes requires 40 % of the aeration energy (only half of the incoming N is oxidized to nitrite by AOB), produces less than 89 % of the sludge and it allows to remove the N by whole autotrophically process (Figure 1.10) (Morales et al. 2015).Thus, all the COD contained in the wastewater is driven for biogas. The SOB and N-damo based processes are interesting since methane and sulphur dissolved in the liquid media are removed improving the wastewater reuse possibilities and decreasing the operational costs (reducing the consumption of COD). However, the limited concentrations of CH4 and H2S in the anaerobically digested municipal wastewater define their potential contribution to the removal of N by these routes and the predominance of the occurring anammox process. N-damo Partialnitritation –anammox + denitratation(PN/AMX+denit) 1.05 NH4+ 0.25 NO3Nitritation AOB AMX Anammox Denitratation 0.95 NH4+ 0.98 N2 0.25 NO21.05 NO2HDN Partialnitritation –anammox (PN-AMX) 1.14 NH4+ 0.22 NO3Nitritation AOB AMX Anammox 0.86 NH4+ 0.89 N2 1.14 NO2Nitritation –denitritation(N-HDN vianitrite) 2 NH4+ Nitritation AOB N2 2 NO2Denitritation HDN Nitrification –denitrification(N-HDN) 2 NH4+ Nitritation AOB N2 2 NO2Denitritation HDN 2 NO3Nitratation NOB HDN Denitratation Introduction 59 and SOB processes might contribute to polish the effluent from anammox based systems by reducing the produced nitrate into nitrite that could be consumed by the anammox process increasing the nitrogen removal efficiency (PN/AMX+denit in Figure 1.9) (Hu et al. 2015, van Kessel et al. 2018, Xie et al. 2018). For example, Stultiens et al. (2019) observed that all the nitrate produced by anammox bacteria was reduced by N-damo archaea and consequently 70 % of nitrite was removed by anammox. Indeed, the coexistence and interaction between the anammox bacteria, N-damo bacteria and archaea and SOB were reported in several studies (Chen et al. 2016, Langone et al. 2014, Pelaz et al. 2018a, Xie et al. 2018). The denitratation process for anammox effluents polishing might be also performed using residual COD requiring approximately 5 % of the total COD present in the wastewater (Figure 1.9 and Figure 1.10). Figure 1.10. Comparison of oxygen consumption for the biological reaction (■, in kg O2/kg N), organic matter use as electron donor (■, in kg COD/kg N) and alkalinity as inorganic carbon (IC) (□, in kg IC/kg N) in the different nitrogen removal combined processes: complete nitrification-dentirification (H-DN), nitritation-denitritation (H-DN nitrite), partial nitritationanammox (PN/AMX) and PN/AMX plus denitratation (PN/AMX+denit). Note that for PN/AMX+denit, hetetrotrophic denitration was considered to estimate the consumptions. In general, nitrite production from AOB is the limiting step to implement the PN/AMX processes. Some authors also proposed the combination of nitrificationdenitratation-anammox processes instead of nitritation (Du et al. 2019), but from an energetic point of view this combination is less favourable than PN/AMX requiring 2.51 kg O2/kg N (to oxidise 55 % of ammonium to nitrate) and 0.63 kg COD/N (for nitrate reduction into nitrite). These features reduce its attractive but it might be a 4.57 3.43 1.90 1.71 2.86 1.71 00.23 0.80 1.16 0.94 0.76 0 1 2 3 4 5 N-HDN N-HDN (nitrite) PN/AMX PN/AMX+denit kg/kg Nconverted Chapter 1 60 solution when the complete NOB suppression is not feasible since the denitratation process is presumably easier to control (Du et al. 2019). 1.3.1. Partial nitritation and anammox processes application in WWTPs The successful PN/AMX implementation relies on favouring the AOB and anammox bacteria growth while outcompeting the undesired NOB. The application of the PN/AMX processes can be performed in two different system configurations: (a) two reactors located in series, where the partial nitritation (PN) takes place in the first aerobic unit and the anammox (AMX) process in the second anoxic unit; (b) a single reactor where both processes occur simultaneously. In the two-reactor configuration, respective optimal operational conditions of both processes are imposed in the corresponding unit. In the one-stage configuration, both processes co-occur by controlling parameters like the DO concentration, reaction time (aerobic and anoxic alternative periods) and/or the distribution of the microorganisms in the reactor by the use of biomass that grows as biofilm. Inside the biofilms, different environments exist, in such a way that AOB grow in the outer layers consuming the oxygen and creating anoxic conditions in the deeper zones of the biofilm for the anammox bacteria (Agrawal et al. 2018). The single reactor configuration is the most frequently applied at full-scale as it requires less complex control systems and lower investment costs (Lackner et al. 2014). The most common application of PN/AMX systems is for the treatment of the supernatant of the anaerobic sludge digester in municipal WWTPs (i.e., sidestream) (Figure 1.11) representing the 75 % of the total plants (Cao et al. 2017, Lackner et al. 2014). This stream amounts to 10 - 20 % of the TN load entering into the WWTPs. The implementation of the PN/AMX processes in the sidestream allows for reducing by approximately 26 % the energy requirements for aeration, producing 18 % more biogas since less organic matter is consumed for denitrification, and increasing by 17 % the solids derived from the primary settling (Morales et al. 2015). Nowadays, the challenge for the PN/AMX processes application at sidestream conditions is related to the implementation of sludge pre-treatments to increase methane production, being the TH the most common method. When a TH process is applied, the obtained digestate is more concentrated in ammonia, but it also contained other compounds such as recalcitrant organic matter, volatile fatty acids Introduction 61 or others still undefined that might affect the PN/AMX processes performance (Figdore et al. 2011, Han et al. 2017, Zhang et al. 2016). Information about the applicability of the PN/AMX processes treating the TH-pretreated AD digestate lacks and more research about this topic is needed. Figure 1.11. Scheme for efficient WWTPs: in the mainstream (⸺) with different alternatives (1, 2 and 3) for the A-stage and with a two-stage or hybrid PN/AMX configuration for the Bstage; while in the sidestream (⸺) a single-stage PN/AMX is proposed. Moreover, 80 - 90 % of the TN load entering the WWTPs is still treated in the CAS system. For the implementation of the PN/AMX processes in the mainstream of municipal WWTPs, the COD needs to be removed in a previous A-stage as it was discussed previously in Section 1.1.4 (Figure 1.11). For source-separation decentralised systems, the anaerobic wastewater digestion is highly recommended. In these installations, wastewater is expected to be more concentrated and the anaerobic sludge digester implementation would be not feasible due to the high investment cost due to the relatively low sludge production to be valorised. 1.4. Application of PN/AMX processes at mainstream conditions The application of anammox based process for the treatment of the mainstream in the WWTPs has been revealed as one of the most promising alternatives for improving the municipal WWTPs energy efficiency enabling to reach the WWTPs self-sufficiency (Garrido et al. 2013, Kartal et al. 2010). The mainstream of a WWTP is characterised by low nitrogen concentration (< 50 mg TN/L), low temperature (< 25 °C) and high variable composition (Cao et al. 2017). Although the application of the mainstream PN/AMX processes has been proposed almost 20 years ago (Kartal et al. 2010, Siegrist et al. 2008), the challenge of attaining stable nitrite pathway performances has hindered its implementation (Agrawal et al. 2018). Pretreatment 1.Anaerobic wastewater digestion 2.High rate activated sludge 3.Chemical enhanced pretreatment Partial nitritation Anammox Partial nitritation Anammox Digested sludge Effluent Anaerobic sludge digester Sludge pretreatment Sewage Sludge Chapter 1 62 1.4.1. Factors affecting NOB suppression NOB suppression is crucial to obtain stable PN/AMX processes performances as they compete with AOB for oxygen and with anammox bacteria for nitrite. In the following section, factors that selectively inhibit or limit NOB growth are summarised. A combination of different factors is usually applied in order to guarantee the NOB suppression (Agrawal et al. 2018). 1.4.1.1. Aerobic sludge retention time At high temperature, AOB grow faster than NOB and therefore if short SRT values are applied NOB are washout from the system. However, at low temperature, NOB growth rates are higher than those from AOB being unfeasible to achieve the nitritation process by solely shortening the SRT (Figure 1.12.A, Jubany et al. 2008, Regmi et al. 2014). 1.4.1.2. Dissolved oxygen concentration Traditionally the application of low DO concentration (< 1.5 mg O2/L) was considered crucial for establishing the nitritation process since AOB have higher oxygen affinity than NOB at sidestream conditions (Blackburne et al. 2008). However, this behaviour is not observed at low temperature since the dominant genus of NOB shifts from Nitrobacter to Nitrospira. Nitrospira spp., as k-strategist, present oxygen affinities higher than AOB (Figure 1.12.B, Ma et al. 2016, Regmi et al. 2014). Even though, under DO limiting concentrations (< 0.2 mg O2/L) stable NOB suppression was obtained but probably due to the combination with other factors (Akaboci et al. 2018, Laureni et al. 2019, Yang et al. 2017). The suitable DO concentration and the oxygen affinities highly vary within systems (Cao et al. 2017) due to the oxygen mass transfer resistance, enrichment degree, existence of biofilm biomass and operational conditions. Introduction 69 2013, Morales et al. 2016) and pilot-scale (Hoekstra et al. 2019, Lotti et al. 2015b, Pedrouso et al. 2018, Seuntjens et al. 2016). The ability to decouple the SRT of different populations (i.e., hybrid systems) arose as an exciting approach to improve the process stability by selectively washing out the NOB. Malovanyy et al. (2015) obtained a NRE of 52 % treating a NLR of 100 mg TN/(L·d) in an integrated fixed-bed activated sludge (IFAS) system at 25 °C fed with anaerobically pretreated wastewater. However, nitrate was still observed in the pilot-scale IFAS systems (Han et al. 2016, Malovanyy et al. 2015). More recently and posterior to the research study performed in this thesis, Laureni et al. (2019) operated a laboratory-scale unit (12 L) and achieved stable PN/AMX processes performance with a NRR of 80 mg TN/(L·d) in a sequencing batch reactor (SBR) IFAS system. This system treated aerobically pretreated municipal wastewater at 15 °C by controlling the SRT of the flocculent sludge and suppressing the NOB activity in the biofilm by limiting the DO concentration. The achieved NRE of 88 ± 5 % and residual TN concentrations of 3 mg TN/L constitute the best performance and produced effluent quality reported so far. The process performances at pilot-scale facing the wastewater composition fluctuations should be assessed. Moreover, Li et al. (2019) reported in a PN/AMX IFAS system (at 30 °C), where granules and flocculent biomass coexisted, the significant reduction of NOB activity but also AOB one by reducing the SRT from 30 to 20 days. Thus, further optimisation of the applied SRT-control strategy is required. 1.4.3. Two-stage configuration Due to the observed difficulties in implementing the PN/AMX processes at mainstream conditions with the one-stage configuration, some research studies were recently focused on splitting the system into two stages optimising both processes separately (Cao et al. 2017). With this configuration, in the aerobic nitritation unit, the remaining organic matter coming from the A-stage will be aerobically oxidised decreasing the potential for heterotrophic denitrifying bacteria development and fostering the autotrophic N removal. Then, in the anoxic unit, the anammox biomass retention can be maximised since no biomass out selection is required. Furthermore, the anammox activity is expected to be promoted as neither COD nor DO will be present in the liquid media. However, scarce information is Chapter 1 70 available, in general, for the operation of mainstream PN/AMX processes, and particularly in the two-stage configuration, treating municipal wastewater (Agrawal et al. 2018, Cao et al. 2017, Pedrouso et al. 2018). In the following sections, a brief description of the main reported studies (still at laboratory scale) are presented. 1.4.3.1. Nitritation process Different strategies for the achievement of the nitritation process were tested but in most of the studies treating municipal wastewater, the long-term stable nitrite accumulation was an issue. The major part of the reported successful strategies relies on the treatment of the nitritation sludge in an external unit by exposing it to inhibitory conditions for NOB. Cui et al. (2019a) proposed to treat the sludge with nitrite. This strategy was characterised by long start-up periods (30 days after 32 days of nitrite exposure). They treated municipal wastewater, but the operational period in the nitritation unit lasted only for 50 more days. Another common strategy is the achievement of the nitritation process by subjecting it to FNA inhibitory concentrations (Wang et al. 2016, Wang et al. 2014). For successful NOB suppression, the combination of DO control and FNA treatment is required (Wang et al. 2016). Indeed, nitrate production was observed when the DO concentration increased from 0.5 to 3 mg O2/L and the stable nitritation process completely disappeared at DO concentrations of 5 mg O2/L (Jiang et al. 2018). Moreover, NOB can ultimately adapt to the relatively high FNA concentration (Duan et al. 2019, Ma et al. 2017). Duan et al. (2019) observed that this adaptation phenomenon could be overcome by alternating FNA and FA treatment of the biomass. The use of the FNA-strategy to treat the sludge requires another nitritation reactor in operation at sidestream, that has associated potential high N2O emissions due to the vast accumulated nitrite concentrations (Jiang et al. 2019, Kampschreur et al. 2009). Optimization of the sludge treatment in terms of concentration of toxic (FNA or FA), ratio of sludge treated and frequency of exposure is required (Duan et al. 2018, Jiang et al. 2018). Other authors controlled the DO/TAN ratio for establishing the nitritation process (Isanta et al. 2015, Reino et al. 2016). These authors operated granular nitritation systems, treating synthetic media, and obtained ammonium oxidation rates of 630 - 700 mg TN/(L·d) at temperatures as low as 10.0 - 12.5 °C. Nevertheless, this strategy seems to be only feasible for granular biomass. Liu et al. (2017b) did not Introduction 71 manage to suppress the NOB in flocculent sludge based on DO/TAN ratio control although higher temperature was applied (22 °C). Zhang et al. (2018) obtained promising results treating municipal wastewater with 91 - 232 mg COD/L and 39 - 79 mg NH4+-N/L at 12 - 17 °C. These authors obtained a nitrite accumulation ratio (NAR) of 97.3 % and COD removals ranging from 60 to 85 % by alternating anaerobic-aerobic conditions for 150 days. The control of the aeration time based on the ammonium valley also produced excellent results of nitrite accumulation when municipal wastewater (containing also organic matter) is treated at low temperatures (Yang et al. 2007). Indeed, (Gu et al. 2012) obtained, at 11 - 16 °C, a stable NAR of 90 % in a pilot-scale reactor (7 m3) treating municipal wastewater by controlling the frequency of the blower supplying air to the system and the pH of operation. 1.4.3.2. Anammox process Few studies addressed the study of single anammox reactors at mainstream conditions (De Cocker et al. 2018, Hendrickx et al. 2014, Laureni et al. 2015, Lotti et al. 2014b, Ma et al. 2013, Reino et al. 2018, Sánchez Guillén et al. 2016). Their main feature is the specific anammox activity (SAAMX) temperature dependence (Dosta et al. 2008, Lotti et al. 2015c, Tomaszewski et al. 2017). It was reported an SAAMX decrease of approximately 10-fold when the temperature diminished from 30 °C to 10 °C (De Cocker et al. 2018). This SAAMX loss might be overcome if sufficient biomass is accumulated in the system. Long SRT are required since doubling time increased from 35 to 77 days with the temperature decrease from 20 to 15 °C (Lotti et al. 2014b). Reino et al. (2018) successfully operated an upflow anammox sludge bed reactor at decreasing temperatures. Despite these authors reported an adverse effect on the anammox process performance after shifting from artificial feeding to municipal wastewater, they achieved a NRR of 1,200 ± 500 mg TN/(L·d) at 11 °C. Less information is available about the performance of the coupled nitritation and anammox processes to treat municipal wastewater in two-stage configuration systems. Liu et al. (2018a) operated a two-stage system treating anaerobically pretreated municipal wastewater at 7 - 15 °C achieving NAR of 98 % in the aerobic reactor by applying the DO/TAN ratio control strategy. In the anammox reactor, a NRE of 80 % and a NRR of 70 - 280 mg TN/(L·d) were reached. Nevertheless, the operational period was limited to less than 60 days. Jin et al. (2019) obtained high Chapter 1 72 NRE ranging from 78 to 90 % by controlling the nitritation process by the ammonium valley based strategy. Nevertheless, the operational temperature was higher than the ones commonly used at mainstream conditions with values of 23 - 25 ° C in the nitritation reactor and 29 - 30 °C in the anammox reactor. 1.4.4. Main challenges of autotrophic nitrogen removal in mainstream Apart from the previously explained conditions required to operate the PN/AMX systems in stable conditions, several challenges still need to be addressed before their practical application at mainstream. Besides NOB suppression, other remaining issues are the control of the fast-growing heterotrophic bacteria overgrowth, the maintenance of good AOB and anammox activities balance, the achievement of good biomass retention and to fulfil the N discharge limits while treating the nitrogen loads (Cao et al. 2017, Hoekstra et al. 2019, Ma et al. 2016). The mainstream municipal wastewater characteristics that slow down the PN/AMX processes implementation are discussed below. 1.4.4.1. Low temperature Although mainstream is characterised by low temperatures that vary throughout the year ranging from 10 to 25 °C (Lackner et al. 2015), most of the studies both at laboratory scale and pilot-scale were performed with temperature control. The study of PN/AMX processes at low temperature and/or without temperature control needs to be further explored. The low temperature causes a decrease in the bacterial activities and growth rates of all microorganisms involved decreasing the potential treated loads (Agrawal et al. 2018). Therefore, to ensure enough biomass retention to cope with the present loads (above 50 mg TN/(L·d) (Burton et al. 2014) is a requisite, in particular for anammox bacteria. This aspect is more challenging in the one-stage configuration than in the two-stage one since anammox bacteria are more sensitive to temperature changes (especially at low temperature) than the nitritation process hindering the maintenance of AOB and anammox bacteria balance (Lotti et al. 2015c, Ma et al. 2016). Moreover, the low temperatures limit the efficiency of the strategy based on controlling the SRT to suppress NOB activity (Cao et al. 2017). Introduction 73 1.4.4.2. Low concentrations and high flows The treatment of diluted wastewater is a challenge due to the low net biomass production associated to the low N concentrations (20 - 60 mg TN/L) treated and to the biomass washout provoked by the high flowrates, which lead to relatively high applied hydraulic loads. To optimise the biomass retention and define the optimum operational HRT maintaining the PN/AMX processes reliability is of great interest. Moreover, the low N concentrations difficult the accumulation of FA and FNA inhibitory concentrations needed to suppress NOB activity compared to the operation with N-rich wastewater streams (Agrawal et al. 2018, Duan et al. 2019). 1.4.4.3. Presence of organic matter The required COD/N ratio (usually below 2 - 3 g COD/g N, Agrawal et al. (2018)) to achieve stable PN/AMX processes performance is difficult to be found at mainstream conditions and hinders its application as it triggers the heterotrophic bacteria proliferation (Leal et al. 2016, Liu et al. 2018b, Xu et al. 2015). The fact that heterotrophic bacteria easily overgrow anammox bacteria is widely reported for one-stage systems (Hoekstra et al. 2019, Lotti et al. 2015b, Seuntjens et al. 2016). Moreover, if COD is present in the influent to the PN/AMX stage its oxidation will increase the aeration requirements and worsen the energy balance of the WWTPs. In the case of two-stage systems, most of the available studies are performed with synthetic wastewater without organic matter. Thus, this aspect should be further investigated to determine whether the residual COD coming from the A-stage could deteriorate the anammox process stability (by its inhibitory effect or heterotrophic bacteria overgrowth) or promote the NRE by polishing the anammox effluent (Giustinianovich et al. 2016, Jin et al. 2019). Moreover, the optimisation of a robust A-stage for organic matter removal and the definition of the COD/N limits that the PN/AMX system can tolerate at mainstream conditions is crucial. 1.4.4.4. Wastewater alkalinity limitations Ammonium oxidation consumes alkalinity. However, when combined with the anammox process, wastewater cannot compensate for this loss as in the case of the denitrification. Therefore, in these conditions, the wastewater buffering capacity could limit the feasibility of the PN/AMX processes due to both the pH decrease or the insufficient IC for the growth of autotrophic bacteria. This effect was investigated Chapter 1 74 during the treatment of industrial wastewater at moderate temperature (Sun et al. 2016) or in batch tests (Kimura et al. 2011, Torà et al. 2010). Although no information is available for the treatment of mainstream in municipal WWTPs. 1.4.4.5. Fluctuation of wastewater characteristics The impact of the fluctuations of the inlet nitrogen concentrations and temperature need to be considered in the future as a possible factor that can compromise the fulfilment of the discharge limits (Pedrouso et al. 2018). Wastewater characteristics vary seasonally (mainly due to the rainfalls and temperature changes) and daily due to the population habits (Burton et al. 2014). The frequent variation in feeding composition can cause over aeration, resulting in a failed suppression of NOB activity, or under aeration, provoking a limitation of AOB activity restricting the nitrite supplied for anammox bacteria. Severe deterioration of NRE would subsequently occur due to the imbalance between control parameters in time (Jin et al. 2019). More straightforward and dynamic control systems are required to cope with these wastewater fluctuations. However, most of the studies are performed under a controlled environment (mainly with artificial feeding and temperature control) without needing to cope with the wastewater characteristics fluctuation that usually WWTPs have to face (Agrawal et al. 2018, Pedrouso et al. 2018). The system resilience to high wastewater variability in terms of temperature and composition need to be further researched to improve the PN/AMX process control and to select the controlled and modifiable parameters under realistic mainstream conditions. 1.4.4.6. Source separation wastewater management Finally, despite source-separation on-site wastewater treatment systems arose as an efficient wastewater management approach, information about the nitrogen removal in these systems is limited in the literature (de Graaff et al. 2011, Vlaeminck et al. 2009). The digested blackwater is characterised by higher and more stable N concentrations and temperatures than municipal wastewater (Eshetu Moges et al. 2018, Gao et al. 2019). Therefore, the implementation of the PN/AMX processes should be more natural than with more diluted streams such as the mainstream of municipal wastewater (Gonzalez-Martinez et al. 2016, Morales et al. 2016). Nevertheless, in small WWTPs the daily pattern of wastewater fluctuations is more acute and wastewater availability might be an issue especially in those systems Introduction 75 treating waster from office buildings, shops or recreational areas where the lack of water in no-working periods might stop the treatment train. 1.5. Aims and scope The overall objective of the present doctoral thesis is to research the feasibility of implementing the autotrophic nitrogen removal PN/AMX processes in conditions of low temperature and low nitrogen concentrations (i.e., mainstream conditions). Special attention is paid to assessing the performance of two-stage PN/AMX systems, while one-stage system feasibility is also evaluated for decentralised wastewater treatment. To accomplish the main aim of this thesis, the following specific goals were defined: • To demonstrate the feasibility of the application of the PN/AMX processes to treat, at ambient temperature, anaerobically digested blackwater originated in decentralised wastewater treatment systems where wastewater availability is discontinuous (Chapter 3). • To stimulate the AOB growth over NOB to operate the nitritation process stable at mainstream conditions (Chapters 4 to 6). More specifically: o To promote the NOB washout while retaining the AOB inside a continuous reactor to accumulate nitrite through a progressive increase of the applied hydraulic load by progressively decreasing the HRT (Chapter 4). o To develop and asses a strategy based on the in situ FNA production to achieve the nitritation process and to maintain its long-term stability. The effect of the influent N/IC ratio over the NOB suppression based on this strategy will be studied as well (Chapter 5). o To prove the feasibility of achieving and maintaining the stable nitritation process treating primary settled municipal wastewater containing organic matter. The operational strategy will be optimised to perform in the same unit the long-term nitritation and organic matter oxidation processes in stable conditions (Chapter 6). Chapter 1 76 • To evaluate the feasibility of applying the anammox process at mainstream conditions using anammox granular sludge without being previously acclimated to low nitrogen, neither low temperature. Firstly, to study the long-term process stability, the reactor will treat mimicked wastewater to have a better understanding of the process performance under different N/IC ratios. Then, the effect of treating partially nitritified municipal wastewater will be evaluated mainly focusing on the effect of pH, low temperature and the presence of organic matter that might lead to the heterotrophic bacteria competition (Chapter 7). • To test the implementation of the combination of the partial nitritation and anammox processes fed with municipal wastewater at pilot-scale (Chapters 8 and 9). Two different configurations will be evaluated treating different pretreated wastewaters: o To analyse the performance and stability of an innovative hybrid onestage PN/AMX reactor (IFAS system) to remove the nitrogen from the effluent of an UASB reactor. The effect of the low temperatures typically found at mainstream conditions will be evaluated by step-wise decreasing the temperature from 21 to 15 °C. The microbial population segregation between suspended and biofilm fractions will be evaluated to understand the role of the reactor configuration for the NOB suppression (Chapter 8). o To explore the feasibility of implementing the two-stage PN/AMX processes treating the effluent of an HRAS full-scale plant. The twostage system will be operated without temperature control and coping with the same wastewater characteristic fluctuations that faced the municipal WWTP where it was located (Chapter 9). • To identify the main outcomes of the thesis and the main gaps that deserve further exploration before the process implementation (Chapter 10). Introduction 77 1.6. References Agrawal, S., Seuntjens, D., Cocker, P.D., Lackner, S. and Vlaeminck, S.E. (2018) Success of mainstream partial nitritation/anammox demands integration of engineering, microbiome and modeling insights. Current Opinion in Biotechnology 50, 214-221. doi: 10.1016/j.copbio.2018.01.013. 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Science of The Total Environment 643, 225-237. doi: 10.1016/j.scitotenv.2018.06.100. Zhou, Y., Oehmen, A., Lim, M., Vadivelu, V. and Ng, W.J. (2011) The role of nitrite and free nitrous acid (FNA) in wastewater treatment plants. Water Research 45(15), 4672-4682. doi: 10.1016/j.watres.2011.06.025. Materials and methods 101 Interferences The most common interference is the chloride anion that reacts with silver cation precipitating silver chloride and inhibiting the catalytic activity of silver. The used Hach Lange kits allow measuring COD in the presence of maximum chloride concentrations of 1,500 mg/L (kits for salty samples are also available), whereas Soto et al. (1989) proposed an adaptation of the method to be applied to high salinity water samples (up to 30 g Cl-/g COD). Bromide, iodide and any other reagent that inactivates the silver ion can interfere similarly. These interferences can be mainly overcome, though not wholly, by complexing these ions with mercuric sulfate. Ammonia and its derivates are not oxidised. However, elemental chlorine reacts with these compounds. Hence, corrections for chloride interferences are difficult. Nitrite exerts a COD of 1.1 mg COD/mg NO2--N but this interference can be eliminated by the addition of 10 mg of sulfamic acid per mg of nitrite present in the sample. Note that sulfamic acid should also be added to the blank sample. Reduced inorganic species (e.g., ferrous iron or sulphide) are oxidised quantitatively under the test condition. For samples containing significant concentrations of these species in a known concentration, stochiometric oxidation can be assumed and COD correction can be made to the obtained COD value. 2.1.1.2. Dissolved total, organic and inorganic carbon The organic carbon in water and wastewater samples include a variety of organic compounds in different oxidation states. Some of these carbonaceous compounds can be further oxidised by biological or chemical processes and the biochemical oxygen demand (BOD5) and COD may be used to characterise these fractions. Unlike COD, total organic carbon (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-WEF, 2012). The measure of TOC is performed according to the method 5310 of the Standard Methods for the Examination of Water and Wastewater (APHA-AWWA-WEF, 2012). 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 is determined by a Shimadzu analyser (TOC-L CSN) as the difference between the total carbon (TC) and the inorganic carbon (IC) Chapter 2 102 concentrations. The instrument is connected to an automated sampler (Shimadzu, ASI-L). Dissolved carbon forms are measured as the equipment only accepts filtered samples. The TC concentrations are determined from the amount of CO2 produced during the combustion of the sample at 720 °C by using platinum immobilised over alumina spheres as a catalyst. High purity air is used as carrier gas supplied at a flow rate of 150 mL/min. The combustion products are carried to an electronic dehumidifier where the gas is cooled and dehydrated. Then, the gas carries the sample combustion products through a halogen scrubber to remove chlorine and other halogens. Finally, the produced CO2 is detected with a non-dispersive infrared (NDIR) analyser, which generates a peak with an area related to the concentration of the compounds. The measured IC consists of carbon derived from carbonates, hydrogen carbonates and dissolved carbon oxide. The IC concentration is obtained from the amount of CO2 produced by acidifying the sample with hydrochloric acid (HCl) 1 N at room temperature to obtain a pH below 3. Then, the produced CO2 is detected with the NDIR analyser. A calibration curve in the range of 0.5 to 200 mg C/L is used for the quantification of the carbon compounds. Commercial standard solutions of total inorganic carbon and total organic carbon of 1,000 ± 5 mg C/L are used as standards to obtain an equation that mathematically expresses the relationship between peak area and TC, TOC and IC concentrations. The detection limit of the equipment is 50 µg/L and 4 µg/L for TC and IC, respectively. 2.1.2. Nitrogen compounds In water and wastewater samples, 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 dinitrogen gas (N2), are biochemically interconvertible following the processes of the nitrogen cycle. 2.1.2.1. Total nitrogen Total Nitrogen (TN) is the sum of organic and inorganic nitrogen. The TN is determined in the same Shimadzu analyser (TOC-L CSN) as the TOC concentration Materials and methods 103 coupled to a TNM-L Unit (Shimadzu). The sample is injected in the combustion tube where the oxidative pyrolysis occurs at 720 ˚C, to convert all nitrogenous compounds into nitric oxide gas (NO). Dinitrogen gas does not become NO under these conditions. The carrier gas together with the produced NO are cooled by an electronic dehumidifier to eliminate any possible condensation. The obtained NO is forced to react with in-situ produced ozone (O3) obtaining nitrogen dioxide in an unstable excited state (NO2*). The NO2* reached its fundamental state by releasing a proton while emitting a photon in the range of 590 - 299 nm, which is detected by a chemiluminescence detector (CLD). The calibration curve is carried out using a standard commercial solution of 1,000 mg NH4+/L (Merck) in the range of 0.5 - 777 mg TN/L. The analyser detection limit is 20 µg TN/L. 2.1.2.2. Ammonium Ammonium concentration is determined spectrophotometrically using two different methods based on the formation of the indophenol blue: 1) the Hach Lange Kits and Dr. Lange Spectrophometer (Hach Lange, Germany) and 2) the method proposed by Bower and Holm-Hansen (1980). In the second method, the indophenol blue is produced by the reaction of the ammonium ions, at pH 12.6, with salicylate and hypochlorite ions, in the presence of sodium nitroprusside as catalyst (Bower and Holm-Hansen 1980). The characteristic blue colour produced by increasing concentrations of ammonia makes the assay useful for the direct, visual estimation of ammonia in culture systems. In Chapters 8 and 9, the Dr. Lange cuvettes instructions are followed, whereas for the Bower and Holm-Hansen (1980) reagents preparation and procedure are detailed below. Reagents ▪ Reagent A: solution of 0.28 g/L of sodium nitroprusside (C5FeN6Na2O) and 440 g/L of sodium salicylate (C7H5NaO3). ▪ Reagent B: solution of 18.5 g/L of sodium hydroxide (NaOH) and 120 g/L of trisodium citrate (Na3C6H5O7). ▪ Reagent C: standard commercial solution of sodium hypochlorite (NaOCl) with 4.00 to 4.99 % active chlorine. ▪ Reagent D: solution prepared by mixing 7 parts of reagent B and 1 part of reagent C. Reagent D is stable for 1 hour after preparation. Chapter 2 104 ▪ Standard commercial solution of 1,000 ± 4 mg NH4+/L (Sigma-Aldrich). Determination procedure The maximum detection limit of the method is 0.9 mg NH4+-N/L; thus, the filtered sample will be diluted if necessary. Then, reagent A (600 μL) and reagent D (1 mL) are added to 5 mL of sample. Samples with reagents are gently shaken and stored, protected from light, to react for more than 2 hours but less than 3 hours. The measurements of the coloured samples are done with a spectrophotometer (Shimadzu, UV-1800) at a wavelength of 640 nm. The quantification is done with a calibration curve in the range of 0 – 0.9 mg NH4+-N/L using a commercial solution of NH4Cl as standard solution and obtaining a regression equation that correlates the measured absorbances with the corresponding ammonium concentration. Interferences Residual chlorine reacts with ammonia and should be removed by sample pretreatment. The determination should be promptly made on fresh samples in order to avoid bacterial conversion. At least filtration of the samples should be done immediately after collection. 2.1.2.3. Nitrite Nitrite concentration is determined following the method 4500-NO2--B (Colourimetric Method) described in the Standard Methods for the Examination of Water and Wastewater (APHA-AWWA-WEF, 2012). The standard procedure is used in all studies except in Chapters 8 and 9 where Hach Lange cuvette kits are employed. Nitrite reacts with primary aromatic amines in acidic solutions to form diazonium salts. These salts combine with aromatic compounds that contain an amino group or a hydroxyl group to form intensively coloured azo dyes. According to the standard method 4500-NO2--B, nitrite is determined through the formation of a reddishpurple azo dye produced at pH 2.0 - 2.5 by coupling diazotised sulphanilamide (C6H8N2O2S) with N-(1-napththyl)-ethylenediamine dihydrochloride (NED dihydrochloride). The reagents preparation and the determination procedure for this method are described below. Materials and methods 105 Reagents ▪ Sulphanilamide: 10 g of sulphanilamide (C6H8N2O2S) are dissolved in 100 mL of concentrated HCl and 600 mL of distilled water. After cooling, the volume is filled up to 1 L with distilled water. ▪ NED solution: 0.5 g of NED are dissolved in 500 mL of distilled water. ▪ Standard commercial solution of 1,000 ± 4 mg/L of NO2- (Sigma-Aldrich). Determination procedure The applicable range of the method is from 0 to 0.3 mg NO2--N/L, therefore samples might be diluted to fit within this range if it is necessary. Then, add 100 µL of each reagent (sulphanilamide and NED) to a volume of 5 mL of sample. After waiting 20 minutes (in the dark) for colour stabilisation, the coloured samples (pink) are measured in a spectrophotometer (Shimadzu, UV-1800) at a wavelength of 543 nm. The quantification is done with a calibration curve in the range of 0 - 0.3 mg NO2- -N/L, using NaNO2 as standard. Interferences Nitrite is highly unstable; therefore, the determination should be promptly made on fresh samples in order to avoid bacterial conversion of nitrite. At least filtration of the samples should be conducted immediately after collection. Chemical incompatibility makes it unlikely that nitrite, free chlorine and nitrogen trichloride (NCl3) coexist in the same sample. NCl3 imparts a false red colour when the colour reagent is added. The following ions interfere because of precipitation under test conditions and should be absent: Sb3+, Au3+, Bi3+, Fe3+, Pb2+, Hg2+, Ag+, chloroplatine and metavanadate. Moreover, cupric ion may cause low results by catalysing the decomposition of the diazonium salt. 2.1.2.4. Nitrate Nitrate concentration is determined according to the method 4500-NO3--B (Ultraviolet (UV) Spectrophotometric Screening Method) described in the Standard Methods for the Examination of Water and Wastewater (APHA-AWWA-WEF, 2012) or by the Dr. Lange cuvette kits (Hach Lange) in Chapters 8 and 9. The presence of nitrite in the sample causes interferences on the nitrate concentration measured value (determined by both methods). For this reason, a spatula-tipfull of sulfamic Chapter 2 106 acid is added to the sample to remove the associated nitrite interferences. The colourimetric method using Hach Lange kits is based on the reaction of nitrate ions in solutions containing sulphuric and phosphoric acids with 2,6-dimethylphenol forming 4-nitro-2,6-dimethylphenol. Then, following the Dr. Lange Hach instructions nitrate concentration is directly obtained in a spectrophotometer. The method 4500-NO3--B used is based on the direct measurement of the nitrate in acid media. Measurement of UV absorption at 220 nm enables the rapid determination of NO3ions. A second measurement at 275 nm is used to correct the NO3value since the dissolved organic matter can absorb at 220 nm and NO3does not absorb at 275 nm. This method is described below. Reagents ▪ Hydrochloric acid: a solution of HCl 1 N. ▪ Sulfamic acid (H3NSO3) powder. ▪ Standard commercial solution of 1,000 ± 4 mg/L of NO3- (Sigma-Aldrich). Determination procedure In a volume of 5 mL of sample (diluted if necessary, to fit the concentration range of the method, up to 3 mg NO3--N/L), 100 µL of HCl (1 N) and a spoon of sulfamic acid are added (approximately 1 - 2 mg). After being thoroughly mixed, the absorbance at 220 and 275 nm is measured in a spectrophotometer (Shimadzu, UV1800) using rectangular quartz cells. The absorbance related to nitrate is obtained by subtracting two times the absorbance read at 275 nm from that read at 220 nm. If the correction value is more than 10 % of the reading at 220 nm, this method should be substituted. The quantification is done with a calibration curve in the range of 0 - 3.0 mg NO3--N/L, using a KNO3 solution as standard. Interferences Dissolved organic matter, surfactants, nitrite and Cr6+ interfere with nitrate determination. Moreover, various inorganic ions such as chlorite and chlorate may interfere. The determination should be promptly made on fresh samples in order to avoid bacterial conversion of nitrite. At least the filtration of the samples should be performed immediately after the samples collection. For more extended storage periods of unchlorinated samples (more than two days), preserve with 2 mL of Materials and methods 107 concentrated H2SO4 (98 %) and store at 4 °C. It is worth to note that when the sample is preserved with acid, nitrate and nitrite cannot be determined as single species. 2.1.3. Inorganic ions The anions nitrite (NO2-), nitrate (NO3-), chloride (Cl-), bromide (Br-), phosphate (PO43-), sulphate (SO42-), thiosulphate (S2O32-) and the cations sodium (Na+), ammonium (NH4+), potassium (K+), magnesium (Mg2+) and calcium (Ca2+) are determined by ion chromatography using a 861 Advanced Compact ion chromatography system equipped with a CO2 suppressor (MCS 853, Metrohm) and a 838 Advanced Sample Processor (Metrohm, Switzerland). Table 2.1 shows the calibration ranges for the different inorganic ion concentrations. Therefore, if it is necessary samples with a higher concentration of one compound should be diluted with distilled water to fit these ranges. Table 2.1. Calibration range of concentrations for the different inorganic ions (mg/L). Anion Lower limit Upper limit Cation Lower limit Upper limit Cl1.0 100 Li+ 0.05 5 NO20.05 5 Na+ 1.5 150 NO30.5 50 NH4+ 0.1 10 Br0.2 20 K+ 0.5 50 PO430.5 50 Mg2+ 0.5 50 SO421.5 150 Ca2+ 0.5 50 S2O321.5 150 The analysed sample goes through the column and the ions (cations or anions depending on the column used) are separated by retention along the resin column (Metrohm, Switzerland). Anions are determined with a Metrosep A column (250 x 4.0 mm) and a carbonate-bicarbonate mobile phase (3.2 mM Na2CO3 and 1.0 mM NaHCO3) at a flow rate of 0.7 mL/min. Cations were determined with a Metrosep C3 resin column (250 x 4.0 mm) and nitric acid 3.5 mM is used as mobile phase. Then, the sample passes through a conductimetric detector after the separation, where the obtained signal corresponding to each retention time is registered. The resulting chromatograms identify each measured ion by its retention time position together with the amount of ion, which is related to the area. Standard commercial solutions Chapter 2 108 (Sigma-Aldrich) of each anion and cation are used to correlate the peak area with the ion concentration. The injection volume of the sample is 20 µL and data collection is done by using the Processor software IC Net 2.3. 2.1.4. Alkalinity The alkalinity is an indicator of the acid-neutralising capacity of a wastewater sample. The alkalinity is related to the presence of buffering agents like salts and weak acids. Since the alkalinity of many surface water is primarily a function of carbonate, bicarbonate and hydroxide content, it is taken as an indication of the concentration of these components (APHA-AWWA-WEF, 2012). However, it may also include contributions from borates, phosphates, silicates and/or other bases. Hydroxyl ions present in a sample as a result of dissociation or hydrolysis of solutes react with additions of standard acid. Alkalinity thus depends on the end-point pH used. The pH values are suggested as the equivalence points for the corresponding alkalinity concentration as mg CaCO3/L. The total alkalinity (TA) is determined at pH 4.3. It can be considered as the sum of the alkalinity due to bicarbonate together with that corresponding to the volatile fatty acids and its end-point corresponds to a pH of 4.3. The partial alkalinity (PA), measured by the titration until pH 5.75, corresponds to bicarbonate while the intermediate alkalinity (IA) is defined as the difference between TA and PA, and corresponds approximately to the alkalinity related to the volatile fatty acid content. In urban wastewater, TA and PA alkalinity are approximately equal. The alkalinity concentration is determined using colourimetric Hach Lange kits (Chapter 8), indirectly utilising the IC concentration (Chapter 3 - 7) or following the method 2320 described in the Standard Methods for the examination of Water and Wastewater (APHA-AWWA-WEF, 2012) in Chapter 9. The colourimetric method using Hach Lange cuvette kits is based on the reaction of calcium and magnesium ions with metal phthalein generating a violet dye. The method 2320 consists in a titration of a sample volume (usually 25 mL) at room temperature with a standard acid (H2SO4 standardised against Na2CO3) to reach the desired pH (pH 5.75 for PA and pH 4.3 for TA). The alkalinity value (mg CaCO3/L) is calculated from Equation 2.3: Alkalinity = A · N · 50,000 V Eq. 2.3 Materials and methods 109 where: A: volume of H2SO4 used to decrease the pH to 5.75 (PA) or to 4.3 (TA) (mL). N: normality of the H2SO4 in equivalents/L. This concentration depends on the expected alkalinity of the water and usually varies from 0.05 to 0.10 N. V: sample volume (mL). 2.1.5. Other control parameters 2.1.5.1. pH The pH is one of the key parameters measured in wastewater biological treatment systems since its control is essential to maintain the activity of the microorganisms involved in the different treatment processes and indicate which process is taking place (e.g., nitrification or denitrification). The pH value, of the unfiltered samples, is measured using different instruments. The specific device is indicated in each chapter. The electrodes are calibrated at room temperature with two standard buffer solutions of 7.0 and 4.01. 2.1.5.2. Dissolved oxygen The dissolved oxygen (DO) concentration in aerobic reactors is measured using different instruments in the laboratory and pilot-scale reactors. The specific device used is specified in the corresponding chapter. DO probes are calibrated monthly. 2.1.5.3. Temperature In the aerobic reactors, the used DO probes mentioned in the previous section are equipped with a thermopar that measures the temperature. In the anaerobic reactors, provided with a chamber in the walls, the temperature is controlled by a thermostatic bath which circulated water at the desired temperature. 2.2. Biomass characterisation In this section, the analytical methods to characterise the solid phase (sludge) are described. Both conventional sludge physical properties and more specific parameters are included, as the performance of different types of biomass is studied in this thesis: suspended (Chapters 4, 5, 6 and 9), granular (Chapters 3, 7 and 9), and attached to carriers (Chapters 8). Chapter 2 110 2.2.1. Solid concentration Solids present in water and wastewater treatment units can be organic or inorganic. The total suspended solids (TSS) and the volatile suspended solids (VSS) concentrations are 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) described in the Standard Methods for the Examination of Water and Wastewater (APHA-AWWA-WEF, 2012). The determination of VSS concentration is of high interest in the control of wastewater treatment process as it offers a rough approximation of the amount of organic matter present in the solid fraction of wastewater or activated sludge samples. Indeed, the VSS inside a reactor is assumed as the biomass concentration. Determination procedure First, a fibreglass filter (Merck Millipore Ltd., APFC04700, 47 mm of diameter, 1.2 μm of pore size) is placed in a muffle furnace (J.P Selecta, Select-Horn-TFT) at 550 °C for 30 minutes to remove the humidity and possible organics and get a constant weight. Then, it is located inside a desiccator to achieve room temperature and after that, it is weighed to obtain the value of F0 (g). For the determination of the TSS, a known well-mixed sample volume (V0 in L) is collected to obtain a residue between 2.5 and 200 mg and filtered through the dried fibreglass filter. The residue retained on the filter is dried for at least 2 hours (and generally for 24 h) at 105 °C in an oven (J.P Selecta 2000210), until reaching a constant weight. After that, it is located inside a desiccator until it achieves room temperature. Finally, it is weighed to obtain the value of F1 (g). The increase in the weight of the filter represents the TSS. The concentration of TSS, in g/L, is determined according to Equation 2.4: TSS (g/L)= F1 - F0 V0 Eq. 2.4 Finally, for the determination of the VSS, the filter previously dried to obtain the TSS concentration is burnt inside a muffle furnace at 550 °C for half an hour. Then, it is located inside the desiccator to reach room temperature and weighed to obtain the value of F2. The weight loss during ignition corresponds to the VSS content and its concentration (in g/L) is determined according to Equation 2.5: Materials and methods 117 calibration is performed with a standard mixture of gases (29.0 % N2, 49.7 % CH4, 17.5 % CO2, 2.4 % N2O and 1.4 % H2S in molar fractions), by using a response factor method. 2.2.6.2. Specific heterotrophic denitrifying activity assays The specific activity of heterotrophic denitrifying microorganisms (SAHDN) is also determined in a manometric batch test by a modified version of the SAAMX test described by Dapena-Mora et al. (2007) but adding 50 mg NO3--N/L or 50 mg NO2-- N/L and 200 mg COD/L as substrates. 2.2.6.3. Specific aerobic activities by respirometric assays Respirometric assays were conducted to determine specific aerobic activities. The aerobic heterotrophic activity and the nitrifying capability of the biomass are quantified by measuring the oxygen uptake rate (OUR). The methodology is adapted from that previously described by Lopez-Fiuza et al. (2002). Respirometric batch experiments are performed using a biological oxygen monitor (BOM, YSI Inc. model 5300) equipped with oxygen selective probes (YSI 5331) connected to a data acquisition system (Labtech) and provided with a thermostatic control chamber for the vials. The batch experiments are performed in hermetically closed vials of 15 mL with a useful volume of 10 mL. First, biomass is washed three times with a phosphate buffer solution (0.143 g KH2PO4/L and 0.747 g K2HPO4/L) to remove residual substrate or inhibitory compounds present in the biomass liquid matrix. Then, the re-suspended biomass is temperature acclimated by means of a thermostatic bath until the established assay temperature is achieved and during at least 30 min. The liquid and biomass mixture are added in the vials, which are placed in the thermostability control chamber, provided with a magnetic stirring system. Later, they are gently bubbled with air for 15 minutes to reach the oxygen saturation. At this moment, 100 µL of the concentrated solution of the specific inhibitory compound needed to ensure that only the microbial population of interest is active is added (Table 2.2). Allylthiourea (ATU) is added when the specific activity of aerobic heterotrophs (SAaerHET) is determined as this compound inhibits the nitrifying bacteria activity. Whereas sodium azide (NaN3) is added when the specific activity of ammonia oxidising bacteria (SAAOB) is determined as NaN3 is a nitrite oxidising bacteria (NOB) selective inhibitory compound. No inhibitory Chapter 2 118 compound is added for the determination of the specific activity of NOB (SANOB). Meanwhile, the two electrodes for oxygen measurement are calibrated to 100 % oxygen saturation. To begin the experiment, aeration is removed, and the oxygen probes are carefully introduced in the vessels avoiding the presence of bubbles and the data acquisition software is initialised. Table 2.2. Substrate and specific inhibitory compounds added during the respirometric activity tests according to Lopez-Fiuza et al (2002) and Mosquera-Corral et al. (2005). Activity Substrate Inhibitory compound added Stoichiometric coefficient (g O2/g N) SAaerHET CH3COONa (100 mg COD/L) ATU (5 mg/L) - SAAOB NH4Cl (35 mg N/L) Sodium azide (5 mg/L) 3.42 SANOB NaNO2 (35 mg N/L) None 1.14 The oxygen depletion is monitored during the time through the connection of the oxygen electrode to the data acquisition system (Figure 2.4). The endogenous respiration is measured at the beginning of the assay during enough time to obtain the slope (α1) of the consumed oxygen (g O2/(L·d)). Then, the substrate was injected into the vial (10 µL) and the new slope of the oxygen consumption was determined (α2) in g O2/(L·d). The oxygen consumption due to the biomass activity is determined by subtracting the endogenous consumption rate to the total oxygen consumption rate (α2 - α1). Figure 2.4. Profile of dissolved oxygen (DO) concentration decrease in the respirometric assay: oxygen depletion during endogenous respiration (blue) and endogenous respiration plus substrate consumption (orange). 0 1 2 3 4 5 6 7 8 9 0100 200 300 400 500 600 700 800 900 DO (mg O2/L) Time (s) α2 α1 Substrate addition Materials and methods 119 After the experiment, the solid concentration in each of the vials was determined according to Standard Methods (APHA-AWWA-WEF, 2012). Finally, the specific activity (SA) of the biomass is determined by dividing the oxygen consumption rate by the solids content in the vial, which can be referred to the specific substrate using the stoichiometric coefficient (Table 2.2). In Chapter 8, the aerobic activities are determined by an adaptation of the method described by Surmacz-Gorska et al. (1996). This method is also based on the OUR but the SAaerHET, SAAOB and SANOB are consecutively determined in a single test. This methodology separates the DO consumption of the different microorganisms by adding specific metabolic inhibitors of NOB and AOB during the time of the assay. The biomass is washed, and temperature acclimated as it is described for the previous method. The biomass in phosphate buffer is introduced in a 1.2 L glass bottle and placed on a magnetic stirrer in a water bath where the liquid media is aerated. Once the temperature acclimation and the DO liquid saturation are reached, the test is started by adding the substrates (COD, ammonium and nitrite simultaneously according to Table 2.3) and introducing the DO meter (Hach Lange LDO meter, HQ30D) to the bottle turning off the aeration. Table 2.3. Substrate and specific inhibitory compounds added and calculations to obtain the specific bacterial activities according to Surmacz-Gorska et al. (1996). Activity Substrate Inhibitory compound Calculation SAaerHET CH3COONa (100 mg COD/L) none α3 SAAOB NH4Cl (50 mg N/L) ATU (5 mg/L) α2 - α3 SANOB NaNO2 (15 mg N/L) NaClO3 (1.8 g/L) α1 - α2 Then, the DO concentration during the time is continuously measured and recorded obtaining a curve like the one shown in Figure 2.5. First, the total OUR is measured obtaining the slope (α1) of the consumed oxygen (g O2/(L·d)). Once the DO concentration decreased approximately 0.7 mg O2/L (or after 10 min), NaClO3 is added inhibiting the NOB activity. The OUR is determined to obtain a slope α2. The difference between the total OUR and the OUR measured in the presence of NaClO3 is ascribed to the oxygen uptake due to the nitrite oxidation (Figure 2.5). Finally, after 8 minutes or a DO concentration decrease of approximately 0.5 mg O2/L, ATU solution is added to inhibit all the nitrifying bacteria activity. The slope of the oxygen Chapter 2 120 depletion is determined (α3) corresponding to the aerobic heterotrophic bacteria activity. The AOB activity is determined by the difference between α2 and α3 in g O2/(L·d) (Figure 2.5). With this method, endogenous respiration is not determined but total activities (the sum of the DO consumption by the substrate uptake and the endogenous respiration). Figure 2.5. Oxygen depletion during the respirometric test: total oxygen consumption (green), total-NOB oxygen consumption (blue) and totalNOB-AOB oxygen consumption (orange). 2.3. Microbial population identification Molecular techniques based on ribosomal ribonucleic acid (rRNA) are presented in the next section. The fluorescence in situ hybridisation (FISH) technique makes the identification of active microorganisms at any desired taxonomical level possible, depending on the specificity of the used probe. The technique based on 16S rRNA gene amplicon libraries (sequenced by Illumina®) allows to quantitatively identify all the present microorganisms. 2.3.1. Fluorescence in situ hybridisation The main active bacterial populations present in the sludge samples are identified by applying FISH molecular technique. Specific regions of the 23S or 16S rRNA are detected by the hybridisation of fluorescent-labelled probes. These probes hybridise with the targeted sequence of any microorganism and they can be later identified microscopically. According to Amann et al. (1990), the FISH protocol includes four steps (Figure 2.6): sample fixation and permeabilisation; hybridisation 7.0 7.5 8.0 8.5 9.0 9.5 0 100 200 300 400 500 600 700 800 900 DO (mg O2/L) Time (s) NaClO3addition ATU addition α3 α2 α1 Materials and methods 121 of the targeted sequence with the specific probe; washing step to remove the unbound probe; and finally, the detection of labelled cells by microscopy. Figure 2.6. Schematic description of the FISH technique methodology. This protocol must be applied to disrupted biomass; thus, the biofilm biomass disintegration is needed before starting the procedure. To achieve the biomass breakage, it is sonicated for 1 minute at 65 % of amplitude and frequency of 0.5 s-1 using an ultrasonic probe device (UP200s, Dr. Hielscher). The time of sonication is selected to achieve the breakage of the granules but not of the cells. Biomass attached to the surface of the K1 carriers (Chapter 8) is manually detached to avoid bacterial deactivation. The hybridisation of the samples is carried out simultaneously for the general bacteria probe and for the specific probes that target microorganisms belonging to the main groups of interest AOB, NOB and anammox bacteria (Table 2.4). The used probes are 5’-labelled with the fluorochromes FITC (fluorescein-5-isocyanate) or Cy3 (Carbocyanine 3). Details about the oligonucleotide probes are available at probeBase (Greuter et al. 2016). DAPI (4,6-diamindino-2-phenylindole) is used as a universal dye for all deoxyribonucleic acid (DNA). Reagents ▪ Phosphate Buffer Solution (PBS) (3x): 0.49 g of KH2PO4 are dissolved in 80 mL of Milli-Q water, 2.3 g of NaCl are added and pH is adjusted to 7.2. Finally, the volume is adjusted to 100 mL. Fresh biomass Fixation Immobilization Hybridization Cy3 FITC Cy3 DAPI Abundances Cy3 FITC DAPI Visualization Image analysis Chapter 2 122 º Table 2.4. List of 16S rRNA-targeted oligonucleotide probes used for biomass fluorescent in situ analysis. Probe Target organism Probe sequence (5’→ 3’) % F ** Bacteria domain EUB338I* Most Bacteria GCT GCC TCC CGT AGG AGT 0-50 EUB338II* Planctomycetales GCA GCC ACC CGT AGG TGT 0-50 EUB338III* Verrucomicrobiales GCT GCC ACC CGT AGG TGT 0-50 AOB group Nso190 Betaproteobacterial ammonia-oxidizing bacteria CGA TCC CCT GCT TTT CTC C 55 Nso1225 Betaproteobacterial ammonia-oxidizing bacteria CGC CAT TGR ATT ACG TGT GA NOB group NIT3 Nitrobacter spp. CCT GTG CTC CAT GCT CCG 40 Competitor for NIT3 CCT GTG CTC CAG GCT CCG Ntspa712 Most members of the phylum Nitrospirae CGC CTT CGC CAC CGG CCT TCC 35 Competitor for Ntspa 712 CGC CTT CGC CAC CGG TGT TCC NTG840 Nitrotoga arctica sp. CTA AGG AGG TCT CCT CCC AMX group Amx368 All anammox bacteria CCT TTC GGG CAT TGC GAA Amx820 Candidatus Brocadia anammoxidans and Kuenenia stuttgartiens AAA ACC CCT CTA CTT AGT GCC C 35 BAN162 Candidatus Brocadia anammoxidans CGG TAG CCC CAA TTG CTT 40 * EUB338I, EUB338II and EUB338III are equimolarly mixed and applied as EUB338mix. ** Percentage of formamide (F) used in the hybridization. Materials and methods 123 ▪ PBS (1x): 1:3 dilution of PBS (3x) in Milli-Q water. ▪ Fixative solution: 6.5 mL of Milli-Q water are heated to 60 °C and 0.4 g of paraformaldehyde are added. One drop of 1 M NaOH is added and the solution is vigorously shaken until it becomes nearly clarified (1 - 2 min). Then, 3.3 mL of PBS (3x) are added and the pH is adjusted to 7.2 with HCl 1 M (one drop). Finally, the solution is filtered through a 0.2 μm membrane filter. ▪ Hybridisation buffer: it is prepared in a 2 mL Eppendorf tube by mixing: 360 µL of NaCl (5 M), 40 µL of Tris/HCl (1 M, pH 8.0) and 4 µL of sodiumdodecylsulfate (SDS, 10 % wt/vol). Formamide (F) and Milli-Q water are added to the mixture, according to the fluorescence probe used (Table 2.5). The hybridisation buffer is kept at room temperature. Table 2.5. Volumes of formamide and water added to the hybridisation buffer. % Formamide (v/v) Formamide (μL) Milli-Q water (μL) 0 0 1,600 5 100 1,500 10 200 1,400 15 300 1,300 20 400 1,200 25 500 1,100 30 600 1,000 35 700 900 40 800 800 45 900 700 50 1,000 600 ▪ Washing buffer: the buffer is prepared in a 50 mL Falcon tube by mixing: 1 mL of Tris/HCl 1 M (pH 8.0) and the required volumes of NaCl (5 M) and ethylenediamine-tetraacetic acid disodium salt (Na2EDTA, 0.5 M at pH 8.0), according to the used probe (Table 2.6). Then, the Falcon tube is filled up to 50 mL with Milli-Q water. The washing buffer is preheated at 48 °C before the washing step. Chapter 2 124 Table 2.6. Volumes of NaCl (5 M) and EDTA (0.5 M) added to the washing buffer. % Formamide (v/v) NaCl 5 M (μL) EDTA 0.5 M (μL) 0 9,000 - 5 6,300 - 10 4,500 - 15 3,180 - 20 2,250 500 25 1,590 500 30 1,120 500 35 800 500 40 560 500 45 400 500 50 280 500 Determination procedure 1. Cell fixation: biomass samples are collected from the reactors and washed in PBS (1x) three times. Then, three volumes of the fixative solution are added to one volume of biomass suspension. The mixture is kept at 4 °C for 2 - 3 h. Afterwards, the biomass sample is washed again, and the cells are re-suspended in PBS (1x). Finally, 1.25 volumes of ethanol 98 % (at -20 °C) are added to one volume of biomass suspension and samples are stored at -20 °C. 2. Immobilisation: a volume of 10 µL of a fixed biomass sample are spread on a well of a coated Teflon/glass microscope slide. The sample is dried at 46 °C for 10 minutes. Afterwards, cells are dehydrated by successive passage through 50, 80 and 98 % (v/v) ethanol (3 minutes each), and dried. 3. Hybridisation: a volume of 10 µL of hybridisation buffer is pipetted into each well of the microscope slide with the immobilised biomass. Then, FISH probes are added into these wells (1 µL of stock solution with a final concentration of 30 ng/µL and 50 ng/µL for Cy3 and FITC labelled-probes, respectively). The solution is mixed without scratching the slide and cell layer. A 50 mL Falcon hybridisation tube is prepared by folding a tissue inside and pouring the rest of the hybridisation buffer onto the tissue. The slide is immediately transferred into the hybridisation tube and it is incubated for 1.5 - 2.0 hour in an oven (Memmert) at 46 °C. In the meantime, the washing buffer is prepared and preheated in a water bath at 48 °C. Materials and methods 125 4. Washing: This step must be performed rapidly. The slides are quickly transferred into the Falcon tube containing the washing buffer by immersing the whole slide and incubating it for 15 minutes at 48 °C. Then, the slides are removed from the washing buffer and dipped into cold Milli-Q water for few seconds. Finally, the slide is dried. 5. Microscopy and slide image acquisition: Slide wells are embedded with Vectashield H-1200 (which amplifies the fluorescence, avoids fading and contains DAPI dye), and a coverslip is set on them. Then, fluorescence signals are recorded with an acquisition system (Coolsnap, Roper Scientific Photometrics) coupled with an epifluorescence microscope (Axioskop 2 plus, Zeiss). Images are acquired with RS Image software (v 1.7.3, Roper Scientific, Inc.). The semi-quantitative counting of the bacterial populations bases on the ratio of their specific biovolume to the total bacterial (or the more general probe applied) bacterial biovolume and it is determined using the DAIME software (Daims et al. 2006). Paired images (FITC, Cy3, or DAPI) of each field of view are stored and they are merged with the Serif Photo plus software (Figure 2.6). 2.3.2. 16S rRNA gene-based amplicon analysis (Illumina®) The 16S rRNA gene-based amplicon analysis (Base Space, Illumina®) is performed, according to the procedure described by Caporaso et al. (2011), to identify the microorganisms present in the biomass samples from Chapter 5. The homogenous biomass samples are directly collected from the reactor and immediately frozen at - 20 °C. The sequence of the V3 - V4 region of the 16S rRNA gene is used as the taxonomic basis to initially identify the bacterial populations present in the samples by the Illumina technique (Caporaso et al. 2011). First, total genomic DNA is extracted according to the phenol-chloroform protocol (Alonso-Gutierrez et al. 2009). The 16S rRNA gene region is amplified through two consecutive rounds of polymerase chain reaction (PCR) technique analysis. The first round of PCR amplifies the targeted region, while the second round of PCR attaches the sample barcode and sequencing adapters. Samples of extracted DNA are amplified using specific primers with the following sequences: 806R and 515F. Total DNA concentrations are quantified in a Qubit fluorometer (Thermo Fisher Scientific, Waltham, USA) and the size integrity is tested by gel electrophoresis. The Chapter 2 126 second PCR of 15 cycles is applied to add the individual barcode to each sample, as well as to incorporate specific sequences in the amplicon libraries. The adapter primers include the specific sequences, the unique barcodes and the universal fusion sequences CS1 (ACACTGACGACATGGTTCTACA) and CS2 (TACGGTAGCAGAGACTTGG TCT). Individual libraries are analysed using a Bioanalyser 2100 (Agilent) to estimate the concentration of the specific PCR products. After library preparation, samples are pooled at equimolar ratios, cleaned and quantified by real-time PCR using specific primers (Kapa Biosystems). Finally, samples are denatured and prepared at 12 pM to be seeded into a Miseq flowcell (Illumina) and run under a 2x250 pairedend sequencing procedure (Parque Científico de Madrid, Spain). A total amount of > 100,000 reads are obtained for each of the analysed samples. After quality filtering and demultiplexing, data are analysed using the 16S rRNA gene-based amplicon sequencing (Base Space, Illumina®). MSR software is used for the analysis. Bioinformatic assays are performed with predominant operational taxonomic units (OTUs). OTUS were considered abundant if the relative abundance was larger than 0.5 % in at least one sample. Each OTU is assigned taxonomic information using the lowest common taxonomic level. After resolving the number of sequences per OTU, the percentage of each organism is individually calculated for each sample. Relative abundances of reads are calculated by taxonomic level for each library. Values represent the percentage of reads of sequences obtained at each taxonomic identity (according to a degree of similarity) within the total set of reading from the library. OTUs that receive no matches against the sequences are identified as “Unclassified.” DNA extraction Reagents: ▪ Sodium phosphate buffer (0.2 M, pH 8.0): 5.3 mL of a stock solution of 2.76 g NaH2PO4/L are mixed with 84.7 mL of a 28.4 g Na2HPO4/L solution. Both solutions prepared with Milli-Q water. ▪ Cetyltrimetylammonium bromide (CTAB, 10 % wt/vol): dissolve 4.1 g of NaCl in 80 mL of Milli-Q water. Slowly add 10 g of CTAB and heat up to 65 °C to dissolve. Finally, the volume is adjusted to 100 mL with Milli-Q water. The solution must be heated at 65 °C before use. Materials and methods 133 2.4.6. Nitrogen removal rates in one-stage PN/AMX system The global NRR and NRE are calculated using Equations 2.19 and 2.20, respectively. Besides these parameters, ammonium nitrogen removal efficiency (ANRE), in %, and the ammonium and nitrite oxidation rates (rAOB and r NOB, respectively) are estimated based on nitrogen mass balances explained in the previous Section as shown in Equations 2.39, 2.40 and 2.41. ANRE (%)= ∆N−(∆NH4 +-N)growth (NH4 +−N+NO2 −−N+NO3 −−N)inf Eq. 2.39 rAOB (mg N L·d)= (∆NH4 +-N)AOB HRT Eq. 2.40 rNOB (mg N L·d)= NO3 −-Ninf −(NO3 −-Neff −(∆NO3 −-N)AMX) HRT Eq. 2.41 2.5. References Alonso-Gutierrez, J., Lekunberri, I., Teira, E., Gasol, J.M., Figueras, A. and Novoa, B. (2009) Bacterioplankton composition of the coastal upwelling system of 'Ria de Vigo', NW Spain. FEMS Microbiology Ecology 70(3), 493-505. doi: 10.1111/j.1574-6941.2009.00766.x. Amann, R.I., Krumholz, L. and Stahl, D.A. (1990) Fluorescent-oligonucleotide probing of whole cells for determinative, phylogenetic, and environmental studies in microbiology. Journal of Bacteriology 172(2), 762-770. doi: 10.1128/jb.172.2.762-770.1990. APHA-AWWA-WEF (2012) Standard methods for the examination of water and wastewater. WashingtonDC, USA. American Public Health Association, American Water Works Association and Water Environment Federation. ISBN 087553287X. Anthonisen, A.C., Loehr, R.C., Prakasam, T.B.S. and Srinath, E.G. (1976) Inhibition of nitrification by ammonia and nitrous acid. Journal Water Pollution Control Federation 48(5), 835-852. doi: 10.2307/25038971. 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. doi: 10.1016/S0043-1354(01)002500. 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. doi: 10.1139/f80-106. Caporaso, J.G., Lauber, C.L., Walters, W.A., Berg-Lyons, D., Lozupone, C.A., Turnbaugh, P.J., Fierer, N. and Knight, R. (2011) Global patterns of 16S rRNA diversity at a depth of millions of sequences per sample. Proceedings of the National Academy of Sciences 108(Supplement 1), 4516-4522. 10.1073/pnas.1000080107. Chapter 2 134 Daims, H., Lucker, S. and Wagner, M. (2006) Daime, a novel image analysis program for microbial ecology and biofilm research. Environmental Microbiology 8(2), 200-213. doi: 10.1111/j.1462-2920.2005.00880.x. Dapena-Mora, A., Fernández, I., Campos, J.L., Mosquera-Corral, A., Méndez, 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), 859865. doi: 10.1016/j.enzmictec.2006.06.018. Greuter, D., Loy, A., Horn, M. and Rattei, T. (2016) probeBase—an online resource for rRNAtargeted oligonucleotide probes and primers: new features 2016. Nucleic Acids Research 44(D1), D586-D589. doi: 10.1093/nar/gkv1232. Lopez-Fiuza, J., Buys, B., Mosquera-Corral, A., Omil, F. and Mendez, 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. doi: 10.1016/S0141-0229(02)00210-7. Mosquera-Corral, A., González, F., Campos, J. L. and Mendez, R. (2005) Partial nitrification in a SHARON reactor in the presence of salts and organic carbon compounds. Process Biochemistry 40(9), 3109-3118. doi: 10.1016/j.procbio.2005.03.042. Soto, M., Veiga, M.C., Méndez, R. and Lema, J.M. (1989) Semi-micro COD. determination method for high-salinity wastewater. Environmental Technology Letters 10(5), 541-548. 10.1080/09593338909384770. Strous, M., Kuenen, J.G. and Jetten, M.S. (1999) Key physiology of anaerobic ammonium oxidation. Applied and Environmental Microbiology 65(7), 3248-3250. Surmacz-Gorska, J., Gernaey, K., Demuynck, C., Vanrolleghem, P. and Verstraete, W. (1996) Nitrification monitoring in activated sludge by oxygen uptake rate (OUR) measurements. Water Research 30(5), 1228-1236. doi: 10.1016/0043-1354(95)00280-4. Chapter 3 PN/AMX system robustness under repeated starvation and reactivation periods for blackwater treatment SUMMARY Wastewater source-separation and on-site treatment systems frequently need to cope with periods of wastewater unavailability. After its separation in source, the black fraction can be treated by anaerobic digestion followed by a partial nitritationanammox (PN/AMX) systems. In this chapter, the effect of repeated short-term starvation and reactivation periods on a PN/AMX processes was assessed at room temperature (ranging from 14 to 21 °C) for the treatment of digested blackwater. The PN/AMX sequencing batch reactor, of 4 L, was fed during working hours with anaerobically digested blackwater from an office building. Repeated stops were applied during night-time (12 h) and weekends (2 days) simulating the discontinuous wastewater production. Despite the low temperature, moderate total nitrogen (TN) concentrations (120 mg TN/L) and regular stops, nitrogen removal efficiencies up to 95 % with nitrogen removal rates of approximately 66 mg TN/(L·d) were obtained. Moreover, the PN/AMX processes performance was immediately recovered after a long starvation period of 15 days (simulating holidays). During the whole operation period, nitrite oxidising bacteria (NOB) activity was suppressed and no significant differences were measured in the specific activities of the involved bacterial populations. The produced effluent presented a good quality in terms of TN concentration that was below 10 mg TN/L, the discharge limit in sensitive areas in the European Union. Results proved for the first time the feasibility of applying the PN/AMX processes for the treatment (and potential reuse) of blackwater produced in a decentralised system where wastewater is not continuously available. Chapter 3 136 OUTLINE 3.1. Introduction 137 3.2. Objectives 139 3.3. Materials and Methods 140 3.3.1. Reactor setup and operation 140 3.3.2. Ex-situ specific activity tests in batch mode 142 3.3.3. Analytical methods 143 3.3.4. Calculations 143 3.4. Results and discussion 144 3.4.1. Performance of the PN/AMX processes 144 3.4.2. Biomass and involved activities 147 3.4.3. Treatment of blackwater via PN/AMX processes 150 3.4.4. PN/AMX system robustness under repeated starvation/reactivation periods 151 3.5. Conclusions 153 3.6. References 153 PN/AMX system robustness under repeated starvation and reactivation periods for blackwater treatment 137 3.1. Introduction The increasing water scarcity and resources depletion have triggered efforts on the implementation of sustainable water management approaches (European Commission, 2016, WWAP, 2017). Decentralised wastewater treatment systems become an attractive alternative to be applied in small agglomerations enabling the energy and nutrients recovery, ensuring the local water availability by reusing the treated water and decreasing both investment and operational costs (WWAP, 2017). Source-separation systems allow segregating the different streams for a more intensive treatment, depending on their characteristics and final use, promoting the water reuse (WWAP, 2017). Blackwater (i.e., toilet water) is an organic matter and nutrients concentrated stream contributing to approximately 92 % of total nitrogen (TN), 75 % of phosphorus and 52 % of the organic matter contained in mixed domestic sewage (Gottardo Morandi et al. 2018). Moreover, blackwater composition considerably varies according to its origin, infrastructure, toilet flushing systems and user habits (Gao et al. 2019, Ren et al. 2018), as it is summarised in Table 3.1 blackwater is more concentrated in residential areas whereas the one deriving from workplaces or touristic installations is generally more diluted. The anaerobic digestion of blackwater allows recovering its energy content as biogas (Gao et al. 2019, Moges et al. 2018). When anaerobic membrane reactors (AnMBR) are used, the high quality and disinfected nitrogen-rich permeate, after the ammonium oxidation to nitrate, may be used as fertiliser while irrigating. However, the irrigation water requirements (volume and nutrients concentration) vary throughout the year and the type of crop growing (European Commission, 2016). Thus, a nitrogen removal system needs to be also considered to obtain a clean effluent suitable for irrigation only, other reuse purposes (e.g., street cleaning, recharge aquifers or industrials uses), or ultimately, for discharge, to reduce the environmental impact of the anaerobic digestion process. Nitrogen removal can be accomplished autotrophically by the combination of the partial nitritation and anammox (PN/AMX) processes which allows the recovery of the wastewater energy contained in the organic matter. Moreover, the anaerobic biodegradability of the blackwater ranges from 40 to 80 %, meaning that residual organic matter is present in the effluent of the anaerobic digester (De Graaff et al. 2010, Gao et al. 2019). Chapter 3 138 Table 3.1. Literature review about blackwater composition depending on the origin and toilets flushing system. 77BOrigin 78B tCOD (g/L) 79B sCOD 80B (g/L) 81BpH 82B TN 8 (mg N/L) 84B NH4+-N 85B (mg N/L) 86B Toilet flushing (L/flush) Reference 88B Office building 89B 2.0 ± 0.2 90B 0.67 ± 0.16 91B 8.0 92B 145 ± 24 a 93B 102 ±8 94B5 toilet + 1.5 urinal [1] 96B Office building 97B 0.4 ± 0.1 98B 0.31 ± 0.03 99B - 100B 74 ± 6 101B 66 ± 7 102B - [2] 104B 32 houses 105B 19 ± 3.4 106B 3.2 ± 0.6 107B 8.6 108B - 109B1400 ± 300 110B 1 [3] 112B 32 houses 113B 7.7 ± 2.5 114B 2.3 ± 0.8 115B 8.6 ± 0.5 116B 1200 ± 180 117B 850 ± 150 118B 1 (7.8 L/(p·d) [4] 120B 44 houses 121B 0.8 - 3.1 122B - 123B 8.9 - 9.1 124B 130 - 180 a 125B - 126B - [5] 128B 15 inhabitants residential 129B 2.9 ± 0.8 130B - 131B 9.0 ± 0.1 132B 273 ± 39 a 133B 202 ± 32 134B 9 [6] 136B 2 houses 137B 0.7 ± 0.1 138B 0.40 ± 0.06 139B - 140B 149 ± 19 141B 139 ± 20 142B - [2] 144B University 145B 9.5 ± 6.5 146B 1.4 ± 0.5 147B 8.8 ± 0.2 148B 1000 ± 130 a 149B 710 ± 10 150B 1 [3] 152B Student dormitory (48 hab.) 153B 5.5 ± 1.3 154B 1.2 ± 0.3 155B 9.0 ± 0.3 156B - 157B 900 ± 200 158B 1.2 [7] 160B Campus lodges 161B 1.1 ± 0.6 162B 0.4 ±1.2 163B 8.0 ± 0.3 164B 180 ± 28 a 165B 147 ± 18 166B 9 [8] 168B Student dormitory (48 hab.) 169B 8.9 - 11.4 170B - 171B - 172B 1400 - 1700a 173B 1.2 [9] 175B Tourist park 176B - 177B 2.2 ± 1.0 178B 8.5 ± 0.6 179B - 180B 810 ± 240 181B 0.8 [10] 183B Hotel 184B - 185B 1.1 ± 0.3 186B 7.2 ± 0.1 187B 194 ± 24 a 188B 164 ±28 189B 4 [11] 191B Clinic 192B 6.1 ± 0.8 193B 7.0 ± 0.2 194B 703 ± 267 a 195B 161 ± 20 196B 4 [11] 198B Residential school 199B 1.7 ± 0.2 200B 0.9 ± 0.2 201B 8.1 ± 0.2 202B 117 ± 28 203B 88 ± 19 204B 5 [12] 206B Fire Station 207B 0.2 ± 0.1 208B 0.13 ± 0.01 209B - 210B 62 ± 7 211B 54 ± 9 212B - [2] 214B 2 Hotels 215B 0.7 ± 0.2 216B 0.45 ± 0.12 217B - 218B 69 ± 16 219B 61 ± 14 220B - [2] tCOD: Total chemical oxygen demand; sCOD: soluble chemical oxygen demand; TN: Total nitrogen; aTKN: Total Kjeldahl Nitrogen. References: [1] 87BGallagher and Sharvelle (2011); [2] Ren et al. (2018); [3] Zeeman et al. (2008); [4] De Graaff et al. (2010); [5] Palmquist and Hanæus (2005); [6] Knerr et al. (2011); [7] Moges et al. (2018); [8] Murat Hocaoglu et al. (2010); [9] Todt et al. (2015); [10] OargaMulec et al. (2017); [11] Lansing et al. (2017); [12] Sharma et al. (2016) PN/AMX system robustness under repeated starvation and reactivation periods for blackwater treatment 139 Nevertheless, previous studies indicate that the PN/AMX processes operate stably at moderate nitrogen concentrations and temperature (i.e. the conditions of the anaerobically digested blackwater) when moderate organic matter concentrations are present in the wastewater (Hoekstra et al. 2019, Pedrouso et al. 2018). Despite the number of research studies about blackwater treatment exponentially rose, scarce information is available about the performance of the anammox based processes to treat this type of stream. When dealing with decentralised systems, the large variations of wastewater flows and concentrations must be considered when defining the operation of the treatment systems. Decentralised systems would have to deal with even more significant fluctuations in both the flow and composition of wastewaters than municipal wastewater treatment plants (European Commission, 2016). In the case of a single office building wastewater, wastewater is not produced during night-time, weekends and holidays. Thus, the biological systems are frequently exposed to famine conditions affecting the process robustness (Wang et al. 2018). Furthermore, in the periods when treated wastewater is used for irrigation nitrogen removal is not required and the corresponding treatment stopped. As anammox bacteria are traditionally considered sensitive to environmental changes, the study of the stability of the anammox process to short-term starvation conditions, mainly focussed on the biomass storage, is available (Wang et al. 2018, Ye et al. 2018). However, limited information exists about the response of the simultaneous PN/AMX processes under oxygen and nitrogen absence (Reeve et al. 2016). In addition, information is missing on the influence of repeated short-term starvation periods on the biomass from a reactor operating in transient conditions, as it can occur in a decentralised system treating blackwater. 3.2. Objectives The main goal of this chapter is to evaluate the performance of a PN/AMX based process, treating anaerobically digested blackwater at room temperature, and to assess the impact of regular supply and absence of fed wastewater. Additionally, the effect of the starvation/reactivation periods on the specific activity of the microbial populations present in the sludge (anammox, ammonium and nitrite oxidisers and heterotrophs) is evaluated. Chapter 3 140 3.3. Materials and Methods 3.3.1. Reactor setup and operation A sequencing batch reactor (SBR) reactor, with a working volume of 4 L and a volume exchange ratio (VER) of 20 % was operated to perform the PN/AMX processes in one-stage (Figure 3.1). The reactor was inoculated with biomass from a full-scale ELAN® reactor (from the Spanish, ELiminación Autótrofa de Nitrógeno, onestage PN/AMX technology with granular sludge) treating the supernatant from an anaerobic sludge digester of a municipal wastewater treatment plant located in Guillarei (Tui, NW Spain) (Morales et al. 2018). Neither temperature nor pH were controlled while the airflow rate (ranging from 1.0 - 1.5 L/min) was manually adjusted by means of a gas flow meter valve (P model, Aalborg). Mechanical stirring (with a rotational speed of 40 - 50 rpm) was provided to guarantee the reactor mixture. A) B) Figure 3.1. A) PN/AMX SBR image and B) Scheme of the PN/AMX reactor set-up. The SBR was fed with anaerobically digested blackwater (see composition in Table 3.2) collected in an office building (nearly 200 employees) located in Porto do Molle Business Center (Nigrán, NW Spain). The used blackwater was less concentrated than the one used in other studies (Table 3.1), as it was mainly composed by urine and diluted using regular flushing toilets (3.0 - 4.5 L/flush). Raw blackwater (Table 3.2) was digested in an AnMBR comprising an anaerobic stirred reactor (2.8 m3) coupled to a membrane tank (1 m3) equipped with an ultrafiltration flat-sheet membrane module (6.25 m2). The AnMBR was operated at room Effluent PLC Digested blackwater Air PN/AMX system robustness under repeated starvation and reactivation periods for blackwater treatment 141 temperature (18 - 26 °C) achieving a 90 % of chemical oxygen demand (COD) and producing an effluent with a COD to nitrogen ratio (COD/N) of 0.9 g COD/g NH4+-N. Table 3.2. Characterisation of the raw and anaerobically digested blackwater from Porto do Molle Business Center. Parameter Raw Anaerobically digested tCOD (mg/L) 2,325 ± 58 98 ± 3 sCOD (mg/L) 553 ± 3 98 ± 3 TOC (mg/L) 117 ± 12 32 ± 6 IC (mg/L) 159 ± 10 124 ± 13 TN (mg/L) 115 ± 6 121 ± 5 NH4+-N (mg/L) 95 ± 5 120 ± 12 NO2--N (mg/L) 0.02 ± 0.01 0.02 ± 0.01 NO3--N (mg/L) 0.3 ± 0.1 0.5 ± 0.1 pH 7.50 ± 0.05 7.25 ± 0.15 Conductivity (mS/cm) - 1.5 ± 0.2 IC: inorganic carbon; TN: total nitrogen; TOC: Total organic carbon; tCOD: total chemical oxygen demand; sCOD: soluble chemical oxygen demand. The SBR operated in 3-hour cycles distributed as depicted in Table 3.3. It was stopped during the night-time (12 hours) and weekends (2 days) simulating the lack of wastewater produced in a decentralised treatment system from an office building. Therefore, it was only fed 4 cycles/day and 5 days/week resulting in hydraulic retention time (HRT) of 1.75 days. Table 3.3. Configuration of the Sequencing Batch Reactor operational cycles. Cycle 1 Anoxic mixed feeding Aerated reaction Settling Withdrawal Time (min) 5 160 10 5 Anoxic mixed feeding Anoxic reaction Aerated reaction Settling Withdrawal Time (min) 5 20 140 10 5 Cycle 2 Chapter 3 142 The SBR was operated for 100 days in three operational stages depending on the cycle configuration and the regime of stops (Table 3.4). It was started-up and operated for 40 days (Stage I) with the cycle configuration named Cycle 1 (Table 3.3). Then an anoxic reaction phase (20 min) was implemented after the feeding (Stage IICycle 2, days 41 - 58). Thus, the aerobic reaction phase was shortened to 140 min (Table 3.3). The reactor was stopped from day 59 to 74 when no monitoring was done. Finally, it was re-started and operated from day 75 to 100 (Stage III) (Table 3.4). Table 3.4. Summary of the SBR operational conditions during the different stages. Stages (days) Temperature (°C)a Cycleb Stops Stage I (0 - 40) 14.0 - 21.3 (19 ± 2) Cycle 1 Nights and weekends Stage II (41 - 58) 14.0 - 20.0 (17 ± 2) Cycle 2 Nights and weekends Starvation (59 - 74) 12.8 - 15.6 (15 ± 1)c No 15 days Stage III (75 - 100) 14.2 - 20.3 (19 ± 2) Cycle 2 Nights and weekends a As temperature was not controlled, the range indicates the minimum and maximum values measured in the period, while the average value with the standard deviation for the corresponding operational period is reported in brackets. b See in Table 3.3 the definition of each cycle. c Low average values due to winter holidays, with no central heating in the laboratory building. 3.3.2. Ex-situ specific activity tests in batch mode Ex-situ specific activity (SA) tests were performed collecting the biomass from the reactor in different operational days and following the corresponding protocol in batch mode. The maximum specific anammox activity (SAAMX) was determined according to the manometric method described by Dapena-Mora et al. (2007) and employing 70 mg N/L of both nitrite and ammonium as substrates. The specific heterotrophic denitrification activity (SAHDN) was assessed by the same procedure but using 200 mg COD/L (as acetate) and 25 mg NO3--N/L as substrates. Respirometric assays were conducted to determine the specific aerobic heterotrophic activity (SAaerHET), as well as specific ammonium and nitrite oxidising activities (SAAOB and SANOB, respectively) (Lopez-Fiuza et al. 2002) using a biological oxygen monitor (BOM, Ysi Inc. model 5300) equipped with oxygen selective probes (YSI 5331). All these activity tests were performed in triplicate at 20 °C, and SAAMX Assessment of the anammox process performance 245 7.3.2. Microbial activity batch tests Maximum specific anammox activity (SAAMX) was determined according to Dapena-Mora et al. (2007) as detailed in Chapter 2. Activity tests were performed at 30 °C (as reference temperature) and 15 °C once per month with biomass samples collected from the SBR. Tests at 20 and 25 °C were performed as well, but with less frequency, to assess the effect of the temperature changes over the SAAMX. Based on the SAAMX method a modification was applied, by adding nitrate or nitrite (50 mg N/L) and acetate (100 mg COD/L) as substrate, to determine the activity of heterotrophic denitrifying (SAHDN) bacteria at 30 °C, in the periods when the reactor was fed with municipal wastewater. All the activity tests were performed in triplicate. Detailed procedure is described in Chapter 2, section 2.2.6. 7.3.3. Response surface methodology Additionally, batch activity tests were performed with the biomass from the reactor taken from days 410 - 420 to assess the effect of temperature, pH and organic matter content over the SAAMX using a response surface methodology (RSM). Statistical methods are powerful tools to find the optimal values of process parameters along with their interactive effects on the response. A three-level-threefactor Box-Behnken design (BBD) (Ferreira et al. 2007) was used for the assessment of the influence of temperature, pH and total organic carbon (TOC) concentrations over the SAAMX. Temperature was evaluated in the range of the optimal value for the test (30 °C) and the operational temperature of the reactor (15 °C), organic matter concentration between 0 and 75 mg TOC/L as they are typical values found in mainstream effluents and pH in the range of 6 to 8. The natural and coded values of the selected independent variables are shown in Table 7.3. Table 7.3. Variables and levels of each parameter assayed according to the Box-Behnken experimental design. Parameter Units Code Low (-1) Medium (0) High (1) Temperature °C x1 15.0 22.5 30.0 pH - x2 6.0 7.0 8.0 TOC mg TOC/L x3 0 37.5 75.0 Chapter 7 246 A total of 15 experiments (in triplicate), including the three replicates in the central point (E7, E8 and E9), were conducted (Ferreira et al. 2007). The full experimental design is shown in (Table 7.4). To assess the reproducibility of the results, the experiments were repeated and consequently, the final number of experiments was 30. The substrate concentrations (ammonium and nitrite) were maintained at 70 mg N/L each, according to Dapena-Mora et al. (2007). The organic matter (as sodium acetate) was added with the substrates and pH value was adjusted to the target value by adding NaOH or HCl in the washing step. Biomass concentration in the vials was similar with average values of 2.6 ± 0.1 g VSS/L. Table 7.4. Defined Box-Behnken factorial experimental design. Test Temperature (°C) TOC (mg/L) pH E1 15.0 0 7 E2 15.0 37.5 6 E3 15.0 37.5 8 E4 15.0 75.0 7 E5 22.5 0 6 E6 22.5 0 8 E7 22.5 37.5 7 E8 22.5 37.5 7 E9 22.5 37.5 7 E10 22.5 75.0 6 E11 22.5 75.0 8 E12 30.0 0 7 E13 30.0 37.5 6 E14 30.0 37.5 8 E15 30.0 75.0 7 The relationships between the response (SAAMX) and the independent variables tested were analysed by linear regression and fitted to a second-order polynomial model using Equation 7.1. 𝑌 = 𝑏0𝑗 + ∑𝑏𝑖𝑗𝑥𝑖+∑∑𝑏𝑖𝑘𝑗𝑥𝑖𝑥𝑘 3 𝑘=1 3 𝑖=1 3 𝑖=1 Eq. 7.1 Where Y represents the predicted response (SAAMX), b0j, bij, and bikj are the regression coefficients calculated from the experimental results by the least-squares method, and xi and xk (k ≥ i) are the independent variables in coded values, with variation Assessment of the anammox process performance 247 ranges from -1 to 1. The statistical analysis was performed using the analysis of variance (ANOVA), including the F-test value, which established the global model significance, the lack of fit and the determination coefficients (R2) and the adjusted R2 (R2adjusted). The significant factor affecting each dependent variable was selected according to the Student t-test establishing a 95 % confidence level. The statistical software IBM SPSS 24 was used to generate the regression analysis and analysis of factor contribution. Excel tool was used to plot the response surface and contour plots. 7.3.4. Analytical methods Liquid samples from influent and effluent of the SBR were periodically taken to follow the process performance. All samples were filtered using 0.45 µm pore-size filters before analysis. Spectrophotometric methods were applied to determine the ammonium (Bower and Holm-Hansen 1980), nitrite and nitrate (APHA-AWWA-WEF, 2012) concentrations. Dissolved total organic and inorganic carbon concentrations (TOC and IC, respectively) were measured with a Shimadzu analyser (TOC-L-CSN). Total nitrogen (TN) concentration was determined in the same Shimadzu analyser with a TNM-L Unit. The pH was measured using an electrode connected to Crison 506 measurer. Concentration of the total suspended solids (TSS), volatile suspended solids (VSS) and sludge volume index at 30 min (SVI30) were determined according to Standard Methods (APHA-AWWA-WEF, 2012). The average diameter of the granules and size distribution were periodically measured by means of a stereomicroscope (Stemi 2000-C, Zeiss) incorporating a digital camera (Coolsnap, Roper Scientific Photometrics) for image acquisition and then these images were processed using the Image ProPlus® software. Details of all the analytical methods are provided in Chapter 2. 7.3.5. Calculations Nitrogen loading rate (NLR) (without considering the nitrate supplied in the influent) is calculated according to Equation 7.2 and NRR and nitrogen removal efficiency (NRE) were calculated from the concentration of the nitrogen forms in the SBR influent and effluent according to the section 2.4.4 in Chapter 2. The specific nitrogen loading and removal rates (sNLR and sNRR) were estimated as the NLR and NRR, respectively, divided by the average VSS concentration inside the reactor. Chapter 7 248 𝑁𝐿𝑅= (𝑁𝐻4 +−𝑁+𝑁𝑂2 −−𝑁)𝑖𝑛𝑓 𝐻𝑅𝑇 Eq. 7.2 7.4. Results and discussion 7.4.1. Anammox process establishment and maintenance 7.4.1.1. Anammox reactor start-up The stable performance of the anammox process was quickly achieved fed with the synthetic media containing a nitrite to ammonium ratio of approximately 1 g NO2--N/NH4+-N and with a NLR of 50 mg TN/(L·d) (Figure 7.3.A). As the inoculum came from a one-stage PN/AMX system (Morales et al. 2015a), it was already enriched in anammox bacteria but also AOB. At the beginning of the operation, nitrate was consumed together with the decrease of the biomass concentration from 1.6 to 1.2 g VSS/L, which took place in 20 days. This might be explained by the lysis of the aerobic bacteria (like AOB) happening during the first operational days as no oxygen was available for them. Then, the organic matter coming from the biomass death was used to denitrify the fed nitrate. Once no organic matter was available, heterotrophic denitrifying activity decayed and significant nitrate consumption was no longer observed from day 10 onwards. The VSS concentration in the reactor remained almost constant from day 50 onwards. As nitrite was fixed as the limiting substrate, even if it was fully depleted, ammonium was always left at concentrations of approximately 5 mg NH4+-N/L. In these conditions, the total nitrogen (TN) concentration in the effluent was 11 ± 2 mg N/L (Figure 7.3.B). During this stage (except for the first 10 days), the obtained average values of nitrite to ammonium consumed ratio and nitrate produced to ammonium consumed ratio were 1.26 ± 0.12 g NO2--N/g NH4+-N and 0.34 ± 0.09 g NO3--N/g NH4+-N, respectively (Figure 7.4). In particular, the first ratio fits well with the anammox stoichiometry (Lotti et al. 2014a, Strous et al. 1999). The produced nitrate to consumed ammonium ratio (Figure 7.4) is too high indicating the presence of NOB, which might profit from small concentrations of dissolved oxygen entering the non-hermetically closed reactor. Assessment of the anammox process performance 249 A) B) Figure 7.3. Evolution throughout the operational time of A) applied nitrogen loading rate (NLR, ○) and achieved nitrogen removal rate (NRR, ●), in mg TN/(L·d) and nitrogen removal efficiency (NRE, ●) in percentage; B) total nitrogen (TN) concentration in the influent (○) and effluent (●) in mg TN/L. The NRE ranged from 74 to 79 % leading to a NRR of 38 ± 4 mg TN/(L·d) (Figure 7.3.A). The maximum achievable NRE according to the anammox stoichiometry would be between 89 % (Strous et al. 1999) and 92 % (Lotti et al. 2014a). It is limited by the nitrate produced simultaneously to the dinitrogen gas. In the present study, obtained maximum NRE was limited to 78 % (Strous et al. 1999) and 86 % (Lotti et al. 2014a), due to the ammonium added in excess in the feeding. During Stages II and III, the nitrate addition to the feeding was first reduced (from 25 to 10 mg N/L) and then stopped (Table 7.2) since the anammox process was properly established and the produced nitrate by the anammox process was enough to maintain the anoxic environment. The average NRE slightly decreased from Stage 0 20 40 60 80 100 0 20 40 60 80 100 050 100 150 200 250 300 350 400 450 500 NRE (%) NLR, NRR (mg N/(L·d)) Time (days) S-I S-II S-III S-VI S-V S-VI S-VII S-VIII 0 10 20 30 40 50 60 050 100 150 200 250 300 350 400 450 500 TN (mg N/L) Time (days) S-I S-II S-III S-VI S-V S-VI S-VII S-VIII Chapter 7 250 I (77.7 ± 1.5 %) to Stage II (76.1 ± 0.8), with a p-value of 0.04, but it recovered in Stage III (77.2 ± 0.7 %). As expected, no significant statistical differences regarding the NRE were detected when Stage III is compared with previous stages (Figure 7.3 and Figure 7.4) with p-values of 0.7 for Stage II and 0.3 in the case of the comparison with Stage III. Thus, it can be stated that the reduction of nitrate fed to the systems has no effect over the system performance. Figure 7.4. Evolution of the nitrite to ammonium consumed ratio (■) and nitrate produced to ammonium consumed ratio (■) in the different operational stages. Horizontal lines represent the stoichiometric values: solid lines according to Strous et al. (1999) and dashed lines according to Lotti et al. (2014a). Moreover, the applied sNLR was always below the SAAMX proving that the system was never overloaded (Figure 7.5). No remarkable differences on the SAAMX were observed from Stage I to Stage III with values of 53 ± 11 and 65 ± 9 mg N/(g VSS·d) on day 0 and day 183, respectively, with p values for the SAAMX higher than 0.10. Therefore, the NRR was limited by the applied NLR as it was also confirmed by the negligible nitrite concentration in the effluent (< 0.01 mg NO2-/L during the whole operational period). Dosta et al. (2008) operated an anammox reactor at decreasing temperatures (from 30 to 15 °C) fed with a NLR of 270 mg N/(L·d). These authors observed that nitrite accumulated in the reactor at 15 °C causing the total loss of the system efficiency. The complete efficiency of the system was only restored when the NLR was diminished to 50 mg N/(L·d). Similar NLR was applied in the present study (Figure 7.3.A) but the SAAMX indicated that higher NRR might be achieved (Figure 7.5). The 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 S-I S-II S-III S-IV S-V S-VI S-VII S-VIII Ratios Assessment of the anammox process performance 251 average SAAMX values obtained in the present study at 15 °C doubled the ones obtained by Dosta et al. (2008) of 20 mg/(g VSS·d) at the same operational temperature, but after acclimation, which indicates that in fact it is not needed. Figure 7.5. Evolution of the applied specific nitrogen loading rate (sNLR, ○) to the reactor and the maximum specific anammox activity (SAAMX, ●) obtained in batch tests at 15 °C. 7.4.1.2. Anammox process performance at decreasing alkalinity concentrations Most part of the operational time with the synthetic media it contained high alkalinity concentrations (Table 7.2). However, due to the nitritation process, this alkalinity would have been consumed. Indeed, the strategy to achieve successful nitritation that has been proven in Chapters 5 and 6 is based on the alkalinity limitation. For this reason, to evaluate the effect of low alkalinity on the anammox process performance is of interest. Different research works studied the short-term N/IC optimal ratio finding that the anammox activity decreased when the ratio is higher than 10 g NH4+-N/g IC (Kimura et al. 2011) or lower than 0.28 g NH4+-N/g IC (Liao et al. 2008). Nevertheless, Strous et al. (1999) and Lotti et al. (2014a) reported stoichiometric NH4+-N/IC consumption ratios of 17.7 and 16.4 g NH4+-N/g IC, respectively, during the anammox process. At the present study, no significant effect was observed over the NRR (p = 0.38) (Figure 7.3) due to the different IC concentrations (Table 7.2) in the feeding, which ranged between 0.19 and 2.50 g NH4+-N/g IC. Kimura et al. (2011) reported a decrease on the NRR in an anammox reactor fed with 160 mg NH4+-N/L and 200 mg NO2--N/L when the IC concentration decreased to 8 mg IC/L (10 mg N/mg IC) and 0 10 20 30 40 50 60 70 80 90 100 0100 200 300 400 500 sNLR, SAAMX (mg N/(g VSS·d)) Time (days) S-I S-II S-III S-VI S-V S-VI S-VII S-VIII Chapter 7 252 found an apparent inorganic carbon affinity constant of 1.2 mg IC/L. In the present work, the highest N/IC ratio tested was 2.50 g N/g IC and the IC concentration was higher than 5 mg IC/L, therefore, no limiting effect was expected and found. The decrease of the influent IC concentrations, in Stages IV to VIII, caused a decrease on the buffering capacity and a slight decrease of the influent pH (from 7.8 ± 0.1 to 7.3 ± 0.2). However, the effect of variable IC concentrations on the pH in the reactor bulk liquid was hard to assess, since the Argon gas bubbling (used to prevent the entry of oxygen) might cause alkalinity depletion (due to CO2 stripping) increasing the pH value (up to 9), or the entrance of oxygen to the system would contribute to the pH descent (close to 6.5) because of the AOB activity. Nonetheless, the pH in the effluent did not seem to be directly affected by the alkalinity depletion in the feeding as it remained at average values of 7.3 ± 0.3. The NRE was 82 ± 3% (Figure 7.3.A) and a ratio of the nitrite to ammonium consumed of 1.28 ± 0.05 g NO2--N/g NH4+-N and a nitrate produced to ammonium consumed ratio of 0.28 ± 0.03 g NO3--N/g NH4+-N (Figure 7.4) were obtained in Stages IV-VI, close to the stoichiometric values (Lotti et al. 2014a, Strous et al. 1999). As nitrite accumulation was not observed, again the NRR was limited by the applied NLR (Figure 7.3.A) and the NRE was limited by the equimolar ammonium and nitrite concentrations in the feeding causing an excess of ammonium compared with the needed according to the anammox stoichiometry. The TN concentration in the effluent was not affected by the alkalinity depletion and it remained at values of 11 ± 3 mg TN/L (Figure 7.3.B, p=0.81). Compared with the inoculum, the SAAMX augmented (p= 3 10 -7) up to 78 ± 8 mg N/(g VSS·d) at the end of Stage VI (Figure 7.5). Whether this SAAMX improvement was due to the alkalinity depletion in the feeding or due to the increase of the anammox enrichment level is uncertain. However, it is worth to point out that, contrary to other studies, the SAAMX did not decrease despite the long-term operation of the reactor at low temperature (15 °C) and low nitrogen concentrations (Agrawal et al. 2018, Cao et al. 2017). A comparison between the sNLR and the SAAMX values revealed that, at the end of the Stage VI, the system was able to treat almost double of the applied sNLR (40 mg TN/(L·d)) (Figure 7.5). Assessment of the anammox process performance 253 7.4.1.3. Dependence of the anammox activity with temperature The SAAMX is known to be highly affected by temperature (Agrawal et al. 2018), experiencing a decrease when the temperature of operation diminishes. In the present study, this behaviour was proven when the SAAMX of biomass samples collected from the reactor was measured at different temperatures (Figure 7.6). Throughout the operational period, all SAAMX measured at 15 °C resulted in the lowest values but they significantly increased from the start-up to Stage VI, with pvalue of 3 10-7, whereas the SAAMX values at 30 °C decreased from 270 ± 11 to 200 ± 10 mg N/(g VSS·d) (p = 2 10 -15). Nevertheless, no significant differences were found for the measured SAAMX values at 20 °C (p = 0.07) and 25 °C (p = 0.17). Figure 7.6. Evolution of the maximum specific anammox activities (SAAMX) throughout the reactor operation measured at different temperatures: 15 °C (●), 20 C (■), 25 °C (▲) and 30 °C (◊). To better understand the effect of the temperature, as an example, Figure 7.7.A shows the SAAMX values of the inoculum (considered acclimated to 30 °C) obtained at different temperatures and those determined for a biomass sample collected on day 370 from the SBR (considered acclimated to 15 °C). The values of the SAAMX for biomass adapted to 15 °C are higher at low temperature (15 and 20 °C), and their diminishing tendency is smoother than in the case of the inoculum. Furthermore, in the case of the inoculum the SAAMX at the reference temperature (30 °C) is more than 5 times higher than at 15 °C, whereas this ratio decreased to 2.6 for sample on day 370 (Figure 7.7.A). In the present study, the highest SAAMX was always obtained at 30 0 50 100 150 200 250 300 0100 200 300 400 500 SAAMX (mg N/(g VSS·d)) Time (days) S-I S-II S-III S-VI S-V S-VI S-VII S-VIII Chapter 7 254 °C, whilst Hu et al. (2013) reported that the optimal temperature of the anammox biomass after long-term operation (300 days) at 12 °C changed from 35 °C to 25 °C. A) B) Figure 7.7. A) Maximum specific anammox activity (SAAMX) temperature dependency from samples collected on day 0 (○) and day 370 (●); B) Arrhenius linearization of the SAAMX at different temperatures determined for the samples collected on day 0 (○) and day 370 (●). The fact that SAAMX values measured at different temperatures fit with the Arrhenius equation was widely reported (Ma et al. 2019, Reino et al. 2018). In this way, the SAAMX expected at low temperature might be predicted from the SAAMX determined at optimal temperatures. However, Lotti et al. (2015) found that models, like the Arrhenius equation, cannot be applied for the complete range of temperature values since the effect of this parameter is not satisfactorily described if the activation energy is considered constant. These authors reported a descent of the activation energy throughout the time of exposition to low temperature. However, in the range of the temperatures tested in the present study (15 - 30 °C) a good correlation was found after applying the Arrhenius model to the data of SAAMX obtaining R2 values higher than 0.98 (Figure 7.7.B). In this range, apparent activation energy (Ea) of 77 KJ/mol was calculated for the inoculum, close to the 70 KJ/mol reported by Strous et al. (1999). However, its value was of 44 KJ/mol on day 370 which reflected the variations of SAAMX due to temperature changes (Figure 7.8) which justified the obtained SAamx values. This might be correlated with an increase in the enrichment level or due to adaptation to the low temperature applied in the SBR. 0 50 100 150 200 250 300 10 20 30 40 SAAMX (mg N/(g VSS·d) Temperature ( C) y = -9313.8x + 33.681 R² = 0.9965 y = -5289.2x + 20.073 R² = 0.9859 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 0.0032 0.0033 0.0034 0.0035 ln (SAAMX ) Temperature-1 (K-1) Assessment of the anammox process performance 261 the low pH, being barely detected in the experiments carried out at pH 6 even at high temperature (30 °C) (Figure 7.11), in good agreement with the results found by Tomaszewski et al. (2017). Figure 7.11. Maximum specific anammox activity (SAAMX) results obtained from the BoxBehnken design experiments. Colours indicate different pH values: 6.0 (■), 7.0 (□) and 8.0 (■). Refer to Table 7.4 to see the used total organic carbon (TOC) in each experimental run. By applying the multiple regression analysis to the experimental data, a secondorder polynomial equation was defined, which allow predicting the SAAMX value at a particular temperature, TOC concentration and pH (in the tested interval). Those terms that were not significant at a 95 % of confidence (i.e., significance higher than 0.05) were excluded from the model one by one starting from the lowest significant one (backward method) (Table 7.5). The obtained polynomial equation, using the coded variables, is shown in Equation 7.3. SAAMX = 134.4 + 56.7 T + 60.8 pH + 41.0 T·pH - 50.2 pH2 Eq. 7.3 The value of the coefficient of determination (R2) of 0.975 indicated that the quadratic polynomial model proposed for SAAMX represents accurately the actual relationship between the responses. The ANOVA analysis from the model showed an adjusted determination coefficient (R2adj) of 0.966 and a high Fisher’s F value (equal to 99) suggesting a high degree of correlation between the experimental and predicted values (Figure 7.12). The points cluster close to the diagonal line (45°) indicating a good fit of the model. 0 50 100 150 200 250 300 E1 E2 E3 E4 E5 E6 E7 E8 E9 E10 E11 E12 E13 E14 E15 SAAMX (mg N/(gVSS·d) Experimental Run T= 15 C T= 22.5 C T= 30 C Chapter 7 262 Table 7.5. Model estimation by multiple regression and ANOVA analysis. Model number Excluded variable Significance of excluded variable Fisher’s Fvalue R2 R2adjusted Estimation error 1 none - 32.743 0.983 0.953 16.06 2 TOC 0.967 44.186 0.983 0.961 14.67 3 TOC_pH 0.948 58.868 0.983 0.967 13.58 4 TOC2 0.620 75.536 0.983 0.970 12.95 5 T_TOC 0.462 94.774 0.981 0.971 12.65 6 T2 0.125 99.238 0.975 0.966 13.78 The results of the Box-Behnken were further subjected to Student’s test (t) (Table 7.68). Low values of t and P indicate a high significance of the corresponding model term. The organic matter concentration (TOC, x2) and its interactions with the other variables were excluded from the model as they were not significant (p < 0.05). The same occurred in the case of the quadratic interaction for temperature while pH interaction with temperature and its square term were found to be significant (Figure 7.12 and Table 7.6) indicating that SAAMX was very sensitive to these factors. Figure 7.12. Experimental versus predicted values for the maximum specific anammox activity (SAAMX) obtained according to Box-Behnken experimental design. Daverey et al. (2015) studied the interaction of pH (5.38 - 9.62) and temperature (21.9 - 43.1 °C) over the SAAMX. These authors also found that the pH value was the most influential factor. Although the temperature was not significant in their model its effect cannot be neglected. Indeed, when they validated the obtained model at low temperature, the SAAMX at 15 °C and pH value of 6.5 was zero y = 0.9754x + 2.6458 R² = 0.9754 0 50 100 150 200 250 050 100 150 200 250 SAAMX predicted (mg N/(g VSS·d)) SAAMX experimental (mg N/(g VSS·d)) Assessment of the anammox process performance 263 whereas at 25 °C was 9 mg N/(g VSS·d). Tomaszewski et al. (2017) stated that despite a statistical correlation between temperature and pH over the SAAMX was not found, the optimal pH range was narrower at low temperatures. Contrary to those findings, the interaction between temperature and pH in the model obtained in the present study was significant (Table 7.6). Table 7.6. ANOVA results for the quadratic model and estimated regression coefficients for SAAMX obtained in the Box-Behnken experimental design Coefficient S.D t constant 134.429 5.210 25.801 Temperature 56.750 4.874 11.644 TOC NS NS -0.049 pH 64.750 4.874 13.286 Temperature_TOC NS NS 0.497 Temperature_pH 41.000 6.892 5.948 Temperature2 NS NS 0.491 TOC_pH NS NS 0.229 TOC2 NS NS -0.199 pH2 -50.179 7.134 -7.033 NS: non-significant coefficient for a 95 % confidence level; S.D: standard deviation; t: TStudent. The analysis of the response surface (Figure 7.13) obtained for the dependent variables helps to visualise the relationships between the SAAMX and the experimental levels of tested variables. The surface plots are graphical diagrams of the regression equations showing the effects of two factors, while one factor is maintained at a fixed level. In this case, as organic matter concentration did not significatively affect the SAAMX, only pH and temperature were used in response surface methodology. SAAMX increased with temperature and the maximum SAAMX was measured at 30 °C and pH 8. Thus, pH highly influenced the SAAMX which rises with the increase of the pH value. The contour plot further revealed that the interaction between temperature and pH is also significant. Chapter 7 264 Figure 7.13. Surface and contour plots showing the interactive effects of temperature and pH on the specific anammox activity (SAAMX). As the SAAMX at pH values of 6.0 was almost negligible but the anammox reactor was able to treat large NRE at pH values of 6.2 ± 0.1, experiments were carried out at these conditions (and at pH 6.5), in triplicate, to validate the adequacy of the model. Data from the SAAMX at different temperatures in the standard buffer solution (pH 7.8) were also used to validate the model. The predicted and the experimental values are compared in Table 7.7. Organic matter was not added as it was found to be irrelevant in the SAAMX values. The experimental values obtained were close to the predicted ones, confirming the validity of the model except for the low pH values for which the error was 26 % and 62 % at 30 °C and 15 °C, respectively. This fact might be explained by the low SAAMX observed that challenge to measure the SAAMX with precision. 8.00 7.50 7.00 6.50 6.00 15 18 21 24 27 30 pH Temperature (°C) 200-250 150-200 100-150 50-100 0-50 SAAMX (mg N/(g VSS·d)) Assessment of the anammox process performance 265 Table 7.7. Experimental and predicted values of the experiments used to validate the model. pH T (°C) SAAMX (mg N/(g VSS·d)) Error (%) experimental predicted 6.0 15 10 3.75 62.5 6.0 30 28 35.25 26.0 6.2 15 30 26.56 11.5 6.2 30 73 74.46 2.4 6.5 15 56 53.26 4.9 6.5 30 130 125.76 3.2 7.8 15 60 64.56 7.6 7.8 20 128 124.26 2.9 7.8 25 190 183.96 3.2 7.8 30 253 243.66 3.7 7.5. Conclusions Results demonstrate that for the establishment of the anammox process at low temperature and nitrogen concentration an acclimation period might not be needed shortening the start-up periods. The possibility of stable operation of anammox reactors at relatively low temperatures (15 °C) has been proved treating both synthetic wastewater and nitritified primary settled wastewater reaching a total nitrogen effluent concentration lower than 10 mg TN/L. The presence of low organic matter concentrations contributed to an increase of the nitrogen removal efficiency polishing the effluent. The composition of the produced effluent will meet discharge requirements only after a precise adjustment of the influent composition in terms of nitrite and ammonium to avoid any of them limit the removal of the other. Although the temperature of operation affects the maximum specific anammox activity once the biomass is adapted to low temperatures its effect becomes less relevant. This means that anammox biomass adapted to operate at low temperature exhibits higher SAAMX than non-acclimated biomass. Anammox biomass was successfully retained in the system and biomass washout was compensated by the anammox biomass growth. Anammox activity Chapter 7 266 measured to the biomass fed was higher when the reactor was with synthetic media than with the municipal wastewater. When municipal wastewater was fed to the system, the SAAMX slightly decreased since heterotrophic denitrifying bacteria were developed due to the presence of residual organic matter. 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Performance of one-stage partial nitritation-anammox processes in an IFAS system 277 monitored by online measurements of temperature, DO concentration (O2X, Cerlic, Sweden), pH (pHX, Cerlic, Sweden), suspended solids (SS) concentration (ITX, Cerlic, Sweden), oxidation-reduction potential (ORP) (ReX, Cerlic, Sweden), ammonium (AmmonLyt®plus 700 IQ, WTW, Germany) and nitrate (NitraLyt®plus 700 IQ, WTW, Germany) concentrations inside the reactor. Additionally, the reactor was also equipped with conductivity sensors of both influent and mixed liquor. All data from the on-line sensors were recorded and collected in a data acquisition system. 8.3.2. IFAS operational conditions The IFAS reactor was operated for years, before beginning the present research study. During those periods, it treated low strength nitrogen streams at temperatures of 21 - 25 °C (Malovanyy et al. 2015a, Malovanyy et al. 2015b). In the present study, the reactor was operated for 100 days, treating similar streams but at lower temperature. The operation was divided into three stages as temperature was step-wise decreased from 21 °C (Stage I, days 0 - 33) to 18 °C (Stage II, days 34 - 79) and finally to 15 °C (Stage III, days 80 - 100). The temperature was continuously monitored and regulated by both immerse cooler and heater connected to a thermostat (Julabo AB, Sweden). Moreover, the intermittent aeration pattern was changed from 15 min aerobic/45 min anoxic periods to 20 min ON/40 min OFF from day 38 onwards. The IFAS reactor was located at the Hammarby Sjöstadsverk research facility, which is managed by the Royal Institute of Technology (KTH) in Stockholm (Sweden). This facility is located directly on top of the Henriksdal WWTP and provided with a direct discharge point for municipal wastewater, which was used as feeding. The municipal wastewater was pre-treated in a primary settler and a subsequent up-flow anaerobic sludge blanket (UASB) reactor (6.3 m3) operated at 20 °C to remove the organic matter (Malovanyy et al. 2015a, Malovanyy et al. 2015b). Then, a fraction of the UASB effluent was filtrated through a 20 µm pore size filter and stored in an equalisation tank (2 m3) to reduce the incoming solids concentration and to mitigate the wastewater composition fluctuations. Afterwards, it was fed to the IFAS unit with the following composition: 38 - 48 mg NH4+-N/L, 110 - 138 mg sCOD/L, 235 – 283 mg CaCO3/L, 5 – 22 mg VSS/L and average pH values of 7.5 ± 0.1. Chapter 8 278 The hydraulic retention time (HRT) in the IFAS reactor was maintained constant at 19 h, and therefore 260 L of water were treated daily resulting in an average NLR of 56 ± 3 mg TN/(L·d), varying only due to the fluctuations of the inlet nitrogen concentration. The SRT of the system was not controlled as biomass was only removed during sampling collection. However, the sludge age could be adjusted by periodically wasting the sludge. 8.3.3. Analytical methods To monitor the pilot plant performance and to check and calibrate the online sensors, both influent and effluent were periodically sampled. Prior to the analysis, samples were filtered using 0.45 µm pore size filters. Then, the concentrations of chemical oxygen demand (COD), total nitrogen (TN), ammonium, nitrite, nitrate and alkalinity were measured spectrophotometrically with Dr Lange test kits (Hach Lange, Germany) in a Photolab® 6600 UV-Vis. Potassium and chloride ion concentrations were also spectrophotometrically determined (with WTW Spectroquant in a Dr Lange ION 500) to calibrate the ammonium and nitrate online sensors, respectively. Concentrations of total suspended solids (TSS) and volatile suspended solids (VSS) were determined according to the Standards Methods (APHA-AWWA-WEF, 2012) in the influent, effluent and the mixed liquor from the bioreactor. The biomass concentration, TSS and VSS, corresponding to the biofilm was determined by the mechanical detachment of the biomass (Figure 8.2) from a randomly selected known number of carriers. A detailed description of all the analytical methods is provided in Chapter 2. Figure 8.2. Image of the used AnoxKaldnes K1 biofilm carriers colonised with biomass (left) and without biomass (right). Performance of one-stage partial nitritation-anammox processes in an IFAS system 279 8.3.4. Microbial activity batch tests To follow up the microbial processes performance inside the reactor, activity tests were carried out on both biofilm and suspended biomass samples taken out from the IFAS reactor. The specific anammox activity (SAAMX) was determined according to the methodology described by Dapena-Mora et al. (2007) using an online pressure transducer (GHM 5155 Greisinger, Germany). A modification of the SAAMX method was also used to determine the activity of heterotrophic denitrifying (SAHDN) bacteria by adding nitrate (50 mg N/L) and acetate (100 mg COD/L) as substrate. Activities of aerobic heterotrophs (SAaerHET), AOB (SAAOB) and NOB (SANOB) were determined by an oxygen uptake rate (OUR) based method according to Surmacz-Gorska et al. (1996). All these batch assays were performed at the corresponding IFAS reactor operational temperature at the moment of the biomass collection. Moreover, SAAMX was also determined at 30 °C, considered as reference temperature for anammox bacteria activity measurement. Detailed protocols are provided in Chapter 2. 8.3.5. Identification of microbial populations The main active bacterial populations present in the suspended sludge and biofilm biomass were identified in samples collected from the IFAS reactor on day 75 (Stage II). For this purpose, the fluorescence in situ hybridisation (FISH) molecular technique was applied following the protocol described by Amann et al. (1990). Biomass was manually detached from the surface of K1 carriers to avoid bacterial deactivation. The specific oligonucleotide probes used (Table 8.2) were 5’-labelled with the fluorochromes FITC (fluorescei-5-isocyanate) or Cy3 (Carbocyanine 3). Details of the oligonucleotide probes are available at probeBase (Greuter et al. 2016). DAPI (4,6-diamindino-2-phenylindole) was used as universal dye for all DNA. Fluorescence signals were recorded with an acquisition system (Coolsnap, Roper Scientific Photometrics) coupled to an Axioskop 2 Plus epifluorescence microscope (Zeiss, Germany). The relative abundances of AOB, NOB and anammox bacteria were estimated by semi-quantitative counting of the ratio between their specific microbial populations biovolume and the total bacterial biovolume using the DAIME software (Daims et al. 2006). Chapter 8 280 Table 8.2. List of 16S rRNA-targeted oligonucleotide probes used for biomass FISH analysis. Probe name Target organism EUB338 I Most Bacteria EUB338 II Planctomycetales EUB338 III Verrucomicrobiales Nso190 Betaproteobacterial ammonia-oxidizing bacteria Nso1225 Betaproteobacterial ammonia-oxidizing bacteria NIT3* Nitrobacter spp. Ntspa712* Most members of the phylum Nitrospirae NTG840 Nitrotoga arctica Amx368 All anammox bacteria Amx820 Candidatus Brocadia anammoxidans and Candidatus. Kuenenia stuttgartiens BAN162 Candidatus Brocadia anammoxidans * For these probes, the specific competitor was applied together with the probe. 8.3.6. Calculations Statistical differences between the results obtained in the three operational stages at different temperatures were tested by one-factor analysis of variance (ANOVA) using the statistical software R version 3.5.2 (The R Foundation Statistical Computing). Mass balances were performed using the Equations described in Chapter 2 section 2.4.6. Full description of the calculations is provided in Chapter 2. 8.4. Results 8.4.1. Nitrogen removal in the IFAS reactor at decreasing temperatures The operational strategy imposed to the IFAS pilot-plant, continuous feeding and intermittent aeration cycles, was defined to promote the performance of anoxic processes and exploit the lag phase of NOB bacteria activity after oxygen starvation. The NOB suppression was crucial for the correct performance of both processes at the low temperatures to be tested of 21, 18 and 15 °C. Initially (Stage I), the daily average DO concentration was of 0.45 mg O2/L (Table 8.3) and the PN/AMX processes performed stable from day 18 onwards. The concentration of TN in the effluent at the end of this stage was lower than 10 mg TN/L corresponding to 4.0 ± 1.5 mg NH4+-N/L, 3.0 ± 2.2 mg NO3--N/L and less than 0.2 mg NO2--N/L (Figure 8.3.A). Ammonium nitrogen removal efficiency (ANRE) and Performance of one-stage partial nitritation-anammox processes in an IFAS system 281 NRE reached values of 80.0 ± 13.4 % and 72.5 ± 12.7 %, respectively (Figure 8.3.B and Table 8.3). The observed nitrate production to ammonium consumption ratio was only 0.094 ± 0.057 g NO3--N/g NH4+-N, which indicated a negligible NOB activity inside the reactor (Table 8.3). Furthermore, the NRE increased reaching average values of 85 % in the last 10 days of this period. Table 8.3. Average values for the different parameters monitored in the IFAS reactor during the three operational stages performed at decreasing temperatures. S-I (21 °C) S-II (18 °C) S-III (15 °C) Phase length (days) 0 - 33 34 - 79 80 -100 Daily DO (mg O2/L) 0.45 ± 0.05 0.58 ± 0.19 0.64 ± 0.07 pH 7.05 ± 0.04 6.83 ± 0.13 7.09 ± 0.23 NRR (mg TN/(L·d)) 40.04 ± 5.05 42.47 ± 7.80 37.35 ± 3.32 NRE (%) 72.5 ± 12.7 74.1 ± 11.4 65.8 ± 4.9 ANRE (%) 80.0 ± 13.4 84.4 ± 11.4 77.5 ± 10.5 (NO3-)produced/(NH4+)consumed 0.094 ± 0.057 0.113 ± 0.091 0.125 ± 0.091 NO2--Neff (mg N/L) 0.10 ± 0.07 0.29 ± 0.13 0.54 ± 0.23 TNeff (mg N/L) 10.9 ± 5.6 10.8 ± 5.3 15.0 ± 2.3 CODeff (mg/L) 32.8 ± 0.7 44.5 ± 3.4 46.0 ± 4.9 VSSr (g/L)* 1.03 ± 0.12 1.05 ± 0.15 1.03 ± 0.09 ANRE: ammonium nitrogen removal efficiency; COD: chemical oxygen demand; NRE: nitrogen removal efficiency; NRR: nitrogen removal rate; TN: total nitrogen; VSS: volatile suspended solids. Sub-indexes “eff” and “r” refer to concentrations measured in the effluent and inside the reactor, respectively. * Biomass concentration corresponding only to the suspended fraction. On day 34 the temperature was decreased to 18 °C (Stage II) resulting in an immediate decrease of the NRE with the consequent increase of ammonium concentration up to 13 mg NH4+-N/L, whereas nitrate concentration remained similar to the values measured at the end of Stage I (< 1 mg NO3--N/L) (Figure 8.3.A). The relatively short oxygen supply periods and the low DO concentration applied (1.5 mg O2/L), resulted in daily average DO concentrations of 0.49 mg O2/L, which might be limiting the ammonium oxidation. Thus, on day 39 the intermittent aeration length was increased from 15 min to 20 min and the anoxic phase reduced from 45 to 40 min, to diminish AOB activity limitations. Then, the average daily DO Chapter 8 282 concentration increased to 0.58 mg O2/L (Table 8.3). Consequently, the ANRE increased and the NRE stabilised at average values of 74 %. In addition, the differences between the NRE and ANRE values corresponding to Stages I (21 °C) and II (18 °C) were statistically insignificant according to the results from ANOVA (p = 0.925) and Kruskal-Wallis (p = 0.485) analysis performed at a level of confidence of 95 % (p < 0.05 for significance). Thus, no effect can be attributed to the temperature reduction, from 21 to 18 °C, over the performance of the PN/AMX processes A) B) Figure 8.3. IFAS reactor performance: A) Evolution of the concentrations of ammonium in the influent (●) and ammonium (○), nitrite (■) and nitrate (▲) in the effluent, in mg N/L. B) Evolution of ammonium removal efficiency (ANRE, ●) and nitrogen removal efficiency (NRE, ○) in %. Finally, the temperature was further diminished to 15 °C (Stage III). NRR of 43 mg TN/(L·d) was reached, with the maximum NRE of 75.6 % and average values of 65.8 ± 4.9 % (Figure 8.3.B and Table 8.3). At 15 °C, both ANRE and NRE decreased 0 5 10 15 20 25 30 35 40 45 50 010 20 30 40 50 60 70 80 90 100 Nitrogen (mg N/L) Time (days) S-I S-II S-III 0 10 20 30 40 50 60 70 80 90 100 010 20 30 40 50 60 70 80 90 100 Removal efficiency (%) Time (days) S-I S-II S-III Performance of one-stage partial nitritation-anammox processes in an IFAS system 283 compared to previous Stages (Kruskal-Wallis p=0.044 and p = 0.0006, compared to Stages I and II respectively). NRE obtained in this last Stage III were significantly different from those measured in the previous ones with 99.92 % of confidence. Whereas the post-hoc Wilcoxon test pointed out that the ANRE from Stages II and III were significantly different (p = 0.040), while no significant differences were found when Stage I and III were compared (p = 0.074). The nitrate production to ammonium consumption ratio slightly increased from 0.094 g NO3--N/g NH4+-N in Stage I to 0.125 g NO3--N/g NH4+-N in Stage III (KruskalWallis, p = 0.036). Furthermore, significant differences in this ratio corresponding to the operation at 18 °C and 15 °C were found (Wilcoxon, p = 0.027). Nitrate concentrations in the effluent fluctuated, reaching sporadic nitrate production to ammonium consumption ratios as high as 0.35 g NO3--N/g NH4+-N. These results indicated the existence of NOB activity in the biomass from the IFAS reactor, as well as the fact that its activity was suppressed inside the reactor due to the imposed operational strategy, which leads to minimum values of 0.05 g NO3--N/g NH4+-N. During the whole operational period, effluent nitrite concentration was almost negligible with a maximum value of 0.89 mg NO2--N/L (Figure 8.3.A and Table 8.3), whereas ammonium was still present, indicating that the nitritation process was limited. Although the nitrite accumulation was scarce, a particular upward trend of the nitrite concentration was observed with the temperature decrease. Thus, at low temperature, AOB activity was presumably favoured over the nitrite consuming bacteria activities (either anammox bacteria or NOB). Although the nitrate production to ammonium consumption ratio was, in Stages I and II, close to the stoichiometric value of 0.11 g NO3--N/g NH4+-N, according to the PN/AMX reactions, in Stage III, at 15 °C, it was frequently higher, diminishing the NRE (Table 8.3). Besides nitrogen, an average COD removal rate of 82 ± 22 mg sCOD/(L·d) was achieved via both aerobic and anoxic routes. Thus, the COD concentration in the effluent remained on average as 43 ± 6 mg sCOD/L, for the whole operational period. This value is lower than the discharge limit in the European Union of 125 mg COD/L. The influent sCOD/N ratio in the influent to the IFAS reactor was relatively high (2.5 ± 0.3 g sCOD/N) and close to the limit values recommended to guarantee the correct performance of the autotrophic processes (Agrawal et al. 2018, Cao et al. 2017), and avoid the development of the heterotrophic denitrifying bacteria which might Chapter 8 284 outcompete the anammox bacteria. However, the ratio corresponding to the COD and N removed obtained was of 2.1 ± 0.5 g sCOD/g N. considerably lower than that generally needed for heterotrophic denitrification of 5 g sCOD/N (Burton et al. 2014). Consequently, the PN/AMX processes were the main route for the obtained nitrogen removal. This is another advantage of the hybrid reactors that usually tolerate higher COD concentration than other one-stage PN/AMX systems. In spite of this, the further optimisation of the operation of the UASB reactor, with current limited COD removal efficiency of 50%, would help to improve the PN/AMX system performance. 8.4.2. Intermittent aerobic/anoxic cycle characterisation Although the reactor was fed in continuous mode, the alternating imposed aerobic/anoxic conditions produced the evolution of the different operational parameters monitored throughout each operational cycle. To evaluate the performance of the IFAS reactor in the three periods operated at 21, 18 and 15 °C, three consecutive operational cycles were monitored on days 20, 74 and 94, respectively (Figure 8.4). Ammonium and nitrate concentrations varied depending on the DO concentration inside the reactor. The biofilm biomass present inside the reactor was expected to be active in both, aerobic and anoxic, periods since the gradient of DO concentration created anoxic environments in the inner zones of the biofilm. The maximum penetration depth of the DO depended on the DO concentration in the media and of course, was zero during the anoxic periods. In these conditions, ammonium was consumed in the aerobic (via AOB and anammox activity) and anoxic (via anammox activity) periods. However, due to the continuous feeding, the ammonium concentration decreased during the aeration periods and increased in the anoxic ones. Ammonium consumption by AOB and anammox bacteria was counterbalanced by the ammonium continuously fed to the system softening the concentration variations. Nitrate concentration increased in the aerobic periods and decreased in the anoxic ones. This concentration decrease was due to both the presence of heterotrophic denitrification activity and a dilution effect caused by the continuous feeding (as influent did not contain nitrate), which counterbalanced its production by the anammox process. The pH values also varied with the aeration pattern decreasing during the aerated phases due to the occurrence of the nitritation process and increasing in the anoxic phases due to the denitrification activity and the effect of new influent coming to the system. Performance of one-stage partial nitritation-anammox processes in an IFAS system 285 A) B) C) Figure 8.4. Evolution of pH (⸺), and concentrations of DO (⸺), ammonium (⸺) and nitrate (⸺) monitored in 3 representative aeration/anoxic cycles with 15 min ON and 45 min OFF in A) Stage I (Day 20); and with 20 min ON and 40 min OFF in B) Stage II (Day 74) and C) Stage III (Day 94). 6.93 6.95 6.97 6.99 7.01 7.03 7.05 7.07 7.09 0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0 025 50 75 100 125 150 175 200 225 250 pH Nitrogen; DO (mg /L) Time (min) 6.70 6.72 6.74 6.76 6.78 6.80 6.82 6.84 0 1 2 3 4 5 6 7 025 50 75 100 125 150 175 200 225 250 pH Nitrogen; DO (mg/L) Time (min) 6.94 6.95 6.96 6.97 6.98 6.99 7.00 7.01 7.02 7.03 0 1 2 3 4 5 6 7 8 9 025 50 75 100 125 150 175 200 225 250 pH Nitrogen; DO (mg/L) Time (min) Chapter 8 286 The concentration profiles obtained during the measured cycles in the three stages (Figure 8.4) together with the known amount of nitrogen fed, allowed estimating the consumption and production rates for ammonium and nitrate in the aerated and anoxic phases (Table 8.4). Table 8.4. Maximum observed conversion rates during the operational cycles performed at different temperatures. 21 °C day 20 18 °C day 74 15 °C day 94 NH4+-N decrease in aerated phase (mg N/L·d)) 59 60 70 NH4+-N decrease in anoxic phase (mg N/(L·d)) 33 32 44 NO3--N increase in aerated phase (mg N/(L·d)) 26 25 30 NO3--N decrease in anoxic phase (mg N/(L·d)) 13 15 8 NLR (mg TN/(L·d)) 51 53 58 NRR (mg TN/(L·d)) 44 39 40 NLR: nitrogen loading rate; NRR: nitrogen removal rate. Despite the higher DO solubility at 18 °C compared to that at 21 °C, the nitrate production rate in the aerated phase (associated to the NOB activity) remained unaltered and the temperature did not negatively affect the rest of the observed nitrogen conversion rates (Table 8.4). The achieved NOB activity repression correlated with the appearance of a lag phase in the nitrate production after the increase of the DO concentration (Figure 8.4). Moreover, it is worth to note that the aerated phase lasted only the 25 % (Stage I) and 33 % (Stage II and III) of the total intermittent aeration cycle of 1 hour (Table 8.3). Therefore, the actual transformation in the aerobic phase accounted for less than the one in the anoxic phase. For example in Stage I, the nitrate concentration increase during the aerobic phase would be 6.6 instead of 26 mg TN/(L·d) in the anoxic one, while the ratio of ammonium transformation in the anoxic compared with the aerated phase would be 1.7 g NH4+anoxic/g NH4+aerobic instead the 0.6 g NH4+anoxic/g NH4+aerobic considering the ratio between rates. These results are in good agreement with the scarce nitrite accumulation observed during the reactor operation (Figure 8.3). At the lowest temperature tested of 15 °C (Table 8.4), ammonium consumption and nitrate production rates significantly increased, probably due to the higher DO availability. At 15 °C, the oxygen solubility increased and complicated the control of Performance of one-stage partial nitritation-anammox processes in an IFAS system 293 A) B) C) D) E) F) G) H) Figure 8.9. FISH images of the biomass sample collected from the IFAS reactor illustrating the distribution of AOB (Nso190: Cy3 dye, orange) in A) biofilm and B) suspended sludge; anammox (AMX820: Cy3 dye, orange) in C) biofilm and D) suspended sludge; Nistrospira sp. (Ntspa712: Cy3 dye, orange) in E) biofilm and F) suspended sludge and Nitrobacter spp. (Nit3: Cy3 dye, orange) in G) biofilm and H) suspended sludge. In all the images Bacteria domain (EUB338 mix) was labelled with FITC dye (green). Chapter 8 294 Anammox bacteria represented approximately 33 % of the active bacterial community in the biofilm whereas they only amounted to 12 % of the suspended sludge fraction. This result confirms the fact that the SAAMX measured in the biofilm was twice the activity in flocculent biomass (Figure 8.6). Candidatus “Brocadia Anammoxidans” was the predominant anammox specie detected (representing 90 % of the total active anammox). Table 8.5. Relative abundances of the anammox, ammonium oxidising bacteria (AOB) and nitrite oxidising bacteria (NOB) populations identified by FISH in biofilm biomass and the suspended sludge (flocs) fractions collected in Stage II from the IFAS reactor. Group (%) Biofilm Suspended Sludge Anammox 33.3 ± 6.5 12.1 ± 3.9 AOB 16.9 ± 4.0 42.9 ± 6.1 NOB Nitrospira sp. Nitrotoga arctica sp. Nitrobacter spp. 21.3 ± 4.3 11.4 ± 3.2 0.9 ± 1.1 2.13 ± 1.9 20.9 ± 1.8 3.6 ± 1.8 In the case of the suspended sludge fraction, AOB relative abundance was significantly higher (42.9 ± 6.1 %) than in the biofilm biomass (16.9 ± 4.0 %). NOB were detected in both biofilm and suspended biomass. However, differences among the dominant species were observed. Nitrospira was the predominant NOB genus in the suspended sludge at a relative abundance of 11.4 % whereas in the case of the biofilm both Nitrospira sp. and Nitrobacter spp. presented similar abundances (Table 8.5). Nitrospira is the most commonly found specie in reactor systems operated at mainstream conditions as these bacteria present high affinity for substrates being able to survive in nitrite limited environments (Cao et al. 2017). Nitrotoga arctica sp. was also detected in both fractions but at low percentages. 8.5. Discussion 8.5.1. Nitrogen removal rate and produced effluent quality In the present study, stable nitrogen removal (72 ± 11 % in average) took place while average NRR of 39 ± 6 mg TN/(L·d) were achieved in the PN/AMX IFAS pilotplant treating anaerobically pre-treated municipal wastewater at low temperature Performance of one-stage partial nitritation-anammox processes in an IFAS system 295 (21 - 15 °C) (Figure 8.3). Reached NRR were in the range of those reported for other systems and/or operated at higher temperatures. As an example, Laureni et al. (2016) operated an SBR lab-scale IFAS system under microaerobic conditions (Table 8.1), and manage to reach NRR of 23 mg TN/(L·d). They hypothesised that the SBR cycle should be optimised by decreasing the HRT to increase achieved NRR. At pilot scale, Pedrouso et al. (2018) also obtained comparable NRR (33 mg TN/(L·d)) treating primary settled municipal wastewater in a granular-biomass SBR at uncontrolled temperature ranging from 12 to 18 °C. Han et al. (2016) obtained higher NRR of 60 mg TN/(L·d), but the operational temperature was also considerably higher (19 - 31 °C). Similarly, Malovanyy et al. (2015a) operated the same IFAS system used in the present research work but at higher temperature (25 °C) achieving a NRR of 52 mg TN/(L·d). Precisely at 15 °C, a relatively high NRE was achieved in the present study (65.8 ± 4.9 % ), but further optimization of the reactor operation is required to meet the discharge limits as the total nitrogen concentration in the effluent was 12 ± 5 mg TN/L (> 10 mg TN/L). In this stage, either ammonium or nitrate contributed to the TN concentration in the effluent. The low effluent quality was already pointed out as one of the drawbacks of the continuously fed PN/AMX systems by Hoekstra et al. (2018) who operated a 4-m3 complete mixed reactor where NRR as high as 97 mg TN/(L·d) at 13.4 ± 1.1 °C was reached. In this sense, the plug flow reactors (Lotti et al. 2015, Yang et al. 2017) or SBRs (Laureni et al. 2016, Pedrouso et al. 2018) are presumably more suitable for effluent quality optimisation than the continuous complete mixed ones. It was demonstrated that using a residual ammonium concentration (2 - 5 mg NH4+/L) as set point either to terminate the reaction (Laureni et al. 2016) or to adjust the airflow rate (Yang et al. 2017) or the HRT (Malovanyy et al. 2015a), higher NRE are achieved. Indeed, Yang et al. (2017) obtained NRE of 82 % at temperatures of 24 - 26 °C (Table 8.1), in a system where also the denitrification process took place since the nitrate concentration in the effluent was lower than that expected considering the PN/AMX processes. In this way, the effluent quality was improved. Chapter 8 296 8.5.2. Overcoming AOB activity limitation Although the obtained AOB/AMX ratios in terms of both activities and corresponding maximum conversion capacity were higher than 2 (Figure 8.6 and Figure 8.8), indicating that enough nitrite could be supplied, the ammonium oxidation was identified as one of the main limiting steps during the reactor operation. A possible explanation is the fact that the SAAOB in the biofilm might be restricted by DO transfer limitation and even if it is high in the flocculent sludge, the contribution of this fraction to the global capacity could be insufficient to oxidise enough ammonium to nitrite for the anammox process. Moreover, the length of the aeration phase of 15 or 20 min might also be too short to reach this required ammonium oxidation. In addition, increasing the length of the aerated phase might also promote the NOB activity, which is an undesired effect. Only during Stage III the observed accumulation of low nitrite concentrations (Table 8.3) indicates that AOB was not the limiting capacity anymore. This fact is explained by both the AOB growth and the NOB suppression, being more nitrite is available for the anammox bacteria. Another possible action would be the optimisation of the previous UASB reactor performance to decrease the COD concentration of the influent to promote the AOB activity inside the IFAS reactor. By lowering the COD concentration of the incoming flow, the aerobic heterotrophic activity would decrease and therefore more DO would be available for the nitritation process. Additionally, in the case of continuously fed IFAS systems, as they are hybrid systems, it has to be considered that the slowest growing organisms present in the sludge grow on the carrier material while the rest of the bacterial populations grow as suspended biomass. This means that, in the present case, AOB, NOB and heterotrophic bacteria grow mainly in the suspended sludge fraction and anammox bacteria are enriched on the carriers. Obtained results support this statement as the suspended biomass fraction presents a specific AOB activity 4 times higher than that of the biofilm (Figure 8.6). Therefore, if the suspended biomass concentration augments the ammonium oxidation capacity of the system would increase too, improving the nitrogen removal of the system. However, to increase this biomass fraction was difficult, since it presented poor settling properties and tended to float in the sedimentation tank. This fact was due to the accumulated N2 gas produced by Performance of one-stage partial nitritation-anammox processes in an IFAS system 297 the heterotrophic denitrification process, which was not appropriately released. However, the system was sufficiently enriched in AOB and anammox bacteria as proven by the estimated AOB/NOB capacity ratio of approximately 1 (Stages II and III) and the AOB/AMX capacity ratio higher than 1.7 (Figure 8.8) to justify the nitrogen was mainly removed via PN/AMX processes. Therefore, IFAS provides alternative control strategies to improve the ammonium removal and overcome one of the main challenges in mainstream conditions. 8.5.3. NOB activity suppression To improve the process stability and effluent quality to suppress the NOB activity is a requisite. In previous studies, Malovanyy et al. (2015a) operated an IFAS reactor and observed that 60 % of the aerobic activity took place in the suspended sludge, while almost the total anammox activity was found in the biofilm. With this in mind, to limit the nitritation activity would be possible, without affecting the anammox bacteria present in the biofilm, by controlling the concentration of flocculent activated sludge. However, the control of the suspended biomass concentration must be applied carefully to avoid damaging the AOB activity. In the present study, the complete suppression of NOB activity based on the selective wash out of the flocculent biomass fraction (Han et al. 2016, Veuillet et al. 2014) was difficult, since NOB were also well integrated into the biofilm (Table 8.5, Figure 8.9 and Figure 8.6). As NOB are aerobic, they should be located in the outer layers of the biofilm. Therefore, they might be washed-out by increasing the shear stress imposed on the system to improve the detachment of the loose parts of the biofilm surface. Then, NOB would be washed-out by adjusting the SRT of the flocculent biomass as it happens in a well-segregated IFAS reactor. It was already shown in Chapter 4 that AOB activity might be promoted over NOB one in a continuous stirred tank reactor operated at low temperature (15 °C) by applying high hydraulic loads. Moreover, as an effect of the temperature decrease the NOB conversion capacity diminished in approximately 50 % for both suspended sludge and biofilm from Stage I to Stage III (Figure 8.7) indicating that despite NOB are present, the operation conditions were adequate to suppress them. As an additional consideration, it was generally accepted that AOB have oxygen affinities higher than NOB and thus PN/AMX reactors should be operated at low DO Chapter 8 298 concentration to decrease the NOB growth rate and improve the NRE (Lackner et al. 2014). However, at mainstream conditions, the use of a strategy based on just the DO control was not suitable for NOB suppression since at low temperatures NOB have oxygen affinities higher than AOB (Regmi et al. 2014, Val del Rio et al. 2019). In these conditions the NOB dominant genus was Nitrospira and not Nitrobacter (Cao et al. 2017). Although intermittent aeration was widely applied to create transient anoxic conditions, which help the NOB out-selection (Malovanyy et al. 2015a, Regmi et al. 2014) the induced lag phase, occurring at the beginning of the aeration phase, failed to completely suppress the NOB activity at low temperature (Agrawal et al. 2018, Cao et al. 2017, Malovanyy et al. 2015a, Regmi et al. 2014). In the present study, with the imposed aeration strategy the NOB activity inside the reactor was limited to 20 % of the SANOB. So, NOB were not completely wash-out from the system. In this sense, Malovanyy et al. (2015a) evaluated the effect of different ratios between the lengths of aerated and non-aerated phases to suppress the NOB activity. The best PN/AMX processes performance achieved in their system (at 25 °C) was with an intermittent aeration pattern of 15 min ON and 45 min OFF, a DO setpoint of 1 mg O2/L and a sCOD/N ratio equal to 1.8 g/g, achieving a NRR of 55 mg TN/(L·d) with a NRE of 70 %. However, the effectiveness of the intermittent aeration strategy also depends on the reactor type and operational conditions. For example, in the present study it was demonstrated that at lower temperatures (18 and 15 C) the aeration pattern of 20 min ON and 40 min OFF was more appropriate. Other strategies for NOB suppression have been proposed (Agrawal et al. 2018). Among them, the operation of the system at high free ammonium (FA) concentrations lead to a good NOB suppression since AOB have no substrate limitation and NOB are inhibited at lower FA concentrations than AOB (Blackburne et al. 2007, Vadivelu et al. 2007). Other authors proposed the use of free nitrous acid (FNA) to inhibit NOB as it is substantially more harmful to NOB than for AOB (Vadivelu et al. 2007). However, with the pH value of the reactor media (ranging from 6.64 to 7.52) neither FA (0.1 mg NH3-N/L) nor FNA (0.02 mg HNO2-N/L) inhibitory concentrations were achieved. Average values were of 0.03 ± 0.02 mg NH3-N/L and 0.10 ± 0.08 µg HNO2-N/L, respectively. Inhibitory FA concentrations would be only possible if large ammonium concentrations were present inside the reactor, but in this way, the NRE would not be high enough to accomplish the stringent disposal Performance of one-stage partial nitritation-anammox processes in an IFAS system 299 requirements. To obtain inhibitory FNA concentrations in a one-stage system would be difficult as anammox bacteria consume nitrite and they are more sensitive to this compound than NOB. This problem might be avoided if the reactor was an SBR or a plug flow reactor where gradient concentrations are easily achieved. 8.5.4. Online monitoring Concerning the biological treatment of wastewater, the ultimate goal when establishing a control strategy is to reach stable operational conditions, high removal efficiency, and minimal energy consumption for aeration. The control strategy relies on either one or a combination of different control parameters that are continuously measured throughout the system operation. The most commonly used control parameters for PN/AMX processes are pH and/or DO concentration (Lackner et al. 2014). In the present study, DO control was implemented; however, to control the nitrogen concentration in the effluent is usually preferred. For this purpose, ammonium nitrogen and nitrate nitrogen sensors required constant monitoring, proper maintenance (at least once a week) and calibration. From the obtained results, a good correlation between online sensors and laboratory analysis was found (Figure 8.10). However, it would be interesting to find correlations of those parameters with others such as pH or conductivity, which are more reliable and cheaper measurements. Malovanyy et al. (2015a) and Pedrouso et al. (2018) already used a pH set point to control the HRT of one-stage PN/AMX system due to its good correlation with the ammonium concentration, as it was observed in this study throughout the monitoring of the aerobic/anoxic phases (Figure 8.4). A) B) Figure 8.10. Correlations between the concentrations determined in laboratory analyses and those measured by the online sensors A) in mg NH4+-N/L and B) in mg NO3--N/L. y = 0.997x + 0.8401 R² = 0.9314 0 5 10 15 20 25 0 5 10 15 20 25 NH4+-N lab (mg N/L) NH4+-N online (mg N/L) y = 0.8205x + 0.7027 R² = 0.908 0 2 4 6 8 10 12 0 2 4 6 8 10 12 NO3--N lab (mg N/L) NO3--N online (mg N/L) Chapter 8 300 Conductivity is also a common parameter used in the control strategy when wastewater streams with high nitrogen concentrations are treated (Lackner et al. 2014). However, in the mainstream wastewater treatment, the changes in this parameter occur in a narrow range and therefore it is not suitable for controlling the PN/AMX processes, at least in a continuous system (Figure 8.11). Figure 8.11. Evolution of the influent (⸺) and effluent (⸺) conductivity (k) values, in µS/cm, collected on day 94 (Stage III). Yang et al. (2016) used a redox probe to monitor the PN/AMX processes at sidestream conditions since it is a cheap sensor and it is more sensitive to the changes in the system than other process control parameters such as DO. However, in the present study, no good correlation between the process performance and the ORP online values was found probably due to, as it was previously reported, it is not an adequate parameter when the process is not in steady-state conditions. Yang et al. (2017) found that the nitrogen removal rate showed a stronger linear correlation with the airflow rate than the one found with the DO concentration and obtained nitrogen removal efficiencies up to 82 % (Table 8.1). These authors also found that nitrate production was significantly affected by residual ammonium observing low nitrate production (4 %) when the ammonium concentration was higher than 1 mg NH4+-N/L. 8.6. Conclusions The feasibility of operating one-stage mainstream PN/AMX processes in an IFAS reactor configuration was proved at pilot scale. Anaerobically pre-treated municipal wastewater was treated at decreasing temperatures (from 21 to 15 °C) at a 500 530 560 590 620 650 935 940 945 950 955 960 025 50 75 100 125 150 175 200 225 250 k effluent (µS/cm) K influent (µS/cm) Time (min) Performance of one-stage partial nitritation-anammox processes in an IFAS system 301 maximum NLR of 56 ± 3 mg TN/(L·d) and achieving an average NRE of 72 ± 11 %. At 15 °C, the process stability was maintained reaching average NRR of 37 ± 3 mg TN/(L·d), comparable with that observed in conventional nutrient removal systems operated in mainstream conditions. The quality of the effluent generated, to fit the appropriated quality for discharge, depending on the nitrate produced due to NOB activity and to a lesser extent on the ammonium and nitrite removal efficiencies. The AOB conversion capacity inside the reactor exceeds that of the anammox biomass indicating that if an action is taken to increase its occurrence the treated nitrogen load in the system could augment. For this purpose, the recycling ratio of the biomass from the sedimentation tank could be increased to augment the concentration of AOB in the system. Microbial population segregation was observed being the AOB and NOB more abundant in the flocculent fraction while the biofilm was mostly composed by anammox bacteria. During the IFAS operation, NOB were present, but their activity was successfully limited observing less than the 20 % and 10 % of the maximum potential NOB activity inside the reactor at 21 and 15 °C, respectively. The successful suppression of most of the NOB activity was achieved by the applied intermittent aeration strategy. Nitrite oxidising capacity decreased during the IFAS operation, but further optimisation to control the nitrate production is required to improve the NRE. 8.7. References Agrawal, S., Seuntjens, D., Cocker, P.D., Lackner, S. and Vlaeminck, S.E. (2018) Success of mainstream partial nitritation/anammox demands integration of engineering, microbiome and modeling insights. Current Opinion in Biotechnology 50, 214-221. doi: 10.1016/j.copbio.2018.01.013. Amann, R.I., Krumholz, L. and Stahl, D.A. (1990) Fluorescent-oligonucleotide probing of whole cells for determinative, phylogenetic, and environmental studies in microbiology. Journal of Bacteriology 172(2), 762-770. doi: 10.1128/jb.172.2.762-770.1990. APHA-AWWA-WEF (2012) Standard methods for the examination of water and wastewater, Washington DC, USA. American Public Health Association, America Water Works Association and Water Environment Federation. ISBN 087553287X. Chapter 8 302 Blackburne, R., Vadivelu, V.M., Yuan, Z. and Keller, J. (2007) Kinetic characterisation of an enriched Nitrospira culture with comparison to Nitrobacter. Water Research 41(14), 3033-3042. doi: 10.1016/j.watres.2007.01.043. Burton, F.L., Tchobanoglous, G., Tsuchihashi, R., Stensel, H.D. and Metcalf & Eddy, I. (2014) Wastewater engineering: treatment and resource recovery, McGraw-Hill Education. ISBN 1259010791. Cao, Y., van Loosdrecht, M.C. and Daigger, G.T. (2017) Mainstream partial nitritationanammox in municipal wastewater treatment: status, bottlenecks, and further studies. Applied Microbiology and Biotechnology 101(4), 1365-1383. doi: 10.1007/s00253-0168058-7. Daims, H., Lücker, S. and Wagner, M. (2006) Daime, a novel image analysis program for microbial ecology and biofilm research. Environmental Microbiology 8(2), 200-213. doi: 10.1111/j.1462-2920.2005.00880.x. Dapena-Mora, A., Fernández, I., Campos, J.L., Mosquera-Corral, A., Méndez, 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), 859865. doi: 10.1016/j.enzmictec.2006.06.018. Greuter, D., Loy, A., Horn, M. and Rattei, T. (2016) probeBase—an online resource for rRNAtargeted oligonucleotide probes and primers: new features 2016. Nucleic Acids Research 44(D1), D586-D589. doi: 10.1093/nar/gkv1232. Han, M., Vlaeminck, S.E., Al-Omari, A., Wett, B., Bott, C., Murthy, S. and De Clippeleir, H. (2016) Uncoupling the solids retention times of flocs and granules in mainstream deammonification: A screen as effective out-selection tool for nitrite oxidizing bacteria. Bioresource Technology 221(Supplement C), 195-204. doi: 10.1016/j.biortech.2016.08.115. Hoekstra, M., Geilvoet, S.P., Hendrickx, T.L.G., van Erp Taalman Kip, C.S., Kleerebezem, R. and van Loosdrecht, M.C.M. (2018) Towards mainstream anammox: lessons learned from pilot-scale research at WWTP Dokhaven. Environ Technol, 1-13. doi: 10.1080/09593330.2018.1470204 Lackner, S., Gilbert, E.M., Vlaeminck, S.E., Joss, A., Horn, H. and van Loosdrecht, M.C.M. (2014) Full-scale partial nitritation/anammox experiences – An application survey. Water Research 55, 292-303. doi: 10.1016/j.watres.2014.02.032. Laureni, M., Falås, P., Robin, O., Wick, A., Weissbrodt, D.G., Nielsen, J.L., Ternes, T.A., Morgenroth, E. and Joss, A. (2016) Mainstream partial nitritation and anammox: longterm process stability and effluent quality at low temperatures. Water Research 101, 628-639. doi: 10.1016/j.watres.2016.05.005. Lotti, T., Kleerebezem, R., Hu, Z., Kartal, B., de Kreuk, M.K., van Erp Taalman Kip, C., Kruit, J., Hendrickx, T.L.G. and van Loosdrecht, M.C.M. (2015) Pilot-scale evaluation of anammoxbased mainstream nitrogen removal from municipal wastewater. Environmental Technology 36(9), 1167-1177. doi: 10.1080/09593330.2014.982722. Malovanyy, A., Trela, J. and Plaza, E. (2015a) Mainstream wastewater treatment in integrated fixed film activated sludge (IFAS) reactor by partial nitritation/anammox process. Bioresource Technology 198, 478-487. doi: 10.1016/j.biortech.2015.08.123. Testing a two-stage partial nitritation and anammox pilot-scale plant implemented in a WWTP 309 units at pilot scale is available. Only, Zeng et al. (2016) treated nitrified sewage in an anammox pilot-scale (160 L) anaerobic biofilter at ambient temperature (18.0 - 21.5 °C) achieving a NRR of 99 mg TN/(L·d). Besides all these research works indicate that the long-term stability of the PN/AMX processes fed with municipal wastewater at pilot and full scale is still uncertain and several challenges need to be addressed before the PN/AMX processes are applied at mainstream conditions. 9.2. Objectives 344BThis study aims at exploring the feasibility of a pilot-scale two-stage PN/AMX system to treat municipal wastewater at not-controlled ambient temperature (11 to 28 °C). The robustness of the system to cope with the fluctuations of the characteristics of the treated wastewater was evaluated as it was implemented in situ in the WWTP. The effluent from the HRAS system, characterised by 22 to 63 mg NH4+-N/L (average 52 ± 2 mg NH4+-N/L) and 25 - 135 mg COD/L (average 75 ± 8 mg COD/L), was used to feed the PN/AMX system. The performance of each PN and AMX unit was evaluated and optimised separately. Special attention was paid to the accomplishment of the discharge limit (< 10 mg TN/L) of the effluent produced in the system and aspects related to the fullscale application (e.g., energy savings, costs) were discussed. 9.3. Materials and Methods 9.3.1. Valdebebas WWTP The Valdebebas municipal WWTP was selected to test the feasibility of the pilot-scale two-stage PN/AMX system. It is located in Madrid (Spain), managed by the public entity Canal de Isabel II and operated by the company FCC Aqualia. This WWTP has a treatment capacity of approximately 260,000 population equivalents and an average wastewater flowrate of 52,000 m3/day. This facility is adequate to evaluate the applicability of the autotrophic nitrogen removal processes since its biological process consists of an HRAS system to remove the organic matter without affecting the nitrogen concentration of the wastewater. In this HRAS system, the aeration was mechanically supplied. The hydraulic retention time (HRT) was low (7 ± 1 h) to enhance the adsorption of COD, from the wastewater, onto the biomass to maximise biogas production during the subsequent anaerobic digestion of the Chapter 9 310 separated sludge. Phosphorus is chemically removed by ferric chloride (FeCl3) dosage in these aeration tanks (approximately 3,500 kg/day), just before going to the secondary settlers also improving the sludge settleability. The effluent from the HRAS unit was characterised by annual average values of 66 ± 11 mg COD/L, 62 ± 11 mg TN/L, 43 ± 10 mg NH4+-N/L, 3 ± 1 mg NO3--N/L, 0.9 ± 0.1 mg TP/L and 18 ± 4 mg SST/L. The low COD/N ratio in the HRAS effluent (approximately 1 g COD/g TN) indicated the excellent process performance and its adequacy to be fed into a PN/AMX system. 9.3.2. Pilot plant set-up and operation Two pilot-scale rectors (600 L each), one aerobic for the partial nitritation process (PN unit) and another anoxic for the anammox process (AMX unit), were operated without temperature control in the Valdebebas WWTP (Figure 9.2). The HRAS effluent was stored in a 220 L intermediate tank from which wastewater was fed to the system. A) B) C) Figure 9.1. Images from: A) the PN/AMX pilot-plant, B) PN unit and C) AMX reactor. 9.3.2.1. Partial nitritation reactor First, the PN unit (Figure 9.1.A and B) was started-up treating the full-scale HRAS effluent. This reactor was seeded with sludge from a conventional activated sludge (CAS) system located in La Gavia WWTP (Madrid, Spain) as it was the nearest Testing a two-stage partial nitritation and anammox pilot-scale plant implemented in a WWTP 311 WWTP also managed by Canal Isabel II where biological nitrogen removal processes took place. After a week facing problems of sludge retention, the reactor was reinoculated by reintroducing the sludge washed-out from the system mixed with the flocculent fraction of ELAN® biomass (enriched in AOB but with presence of NOB too) coming from a one-stage PN/AMX reactor treating the supernatant of the sludge digester operated at mesophilic conditions (i.e. sidestream PN/AMX reactor), located in Guillarei WWTP (Tui, NW Spain) (Morales et al. 2018). A set of two peristaltic pumps were used to feed (from the top) and discharge (by adjusting the final liquid level) the reactor, respectively (Figure 2). Dissolved oxygen was supplied by introducing air through diffusers placed at the bottom of the reactor. The applied airflow rate ranged between 5 and 10 L/min. The DO concentration was monitored, but not controlled, utilising an online sensor (Hach LDO). Moreover, the pilot plant was equipped with online probes to measure pH (pHD, Hach) and ion-selective electrodes to determine ammonium and nitrate concentrations (AN-ISE, Hach). All the probes were connected to a sc100TM controller (Hach) for data acquisition. Figure 9.2. Scheme of the PN/AMX pilot-plant implemented in the Valdebebas WWTP. The pilot plant operated as a sequencing batch reactor (SBR) and the SCADA software was used to monitor the pilot-plant process and operation. The cycle distribution (Figure 9.3) was designed to promote the nitritation process and then, due to technical limitations, the anammox cycle duration was the same. The SBR Municipal wastewater PLC Air Effluent pH probe Chapter 9 312 cycles from the PN unit were divided into four phases (Figure 9.3.A) comprising: aerated feeding, aerobic reaction (with variable length), settling and discharge. A) Feeding Aeration Settling Discharge Time (min) 90 and level Variable 40* 25 or level B) Stirring Settling Discharge Feeding Time (min) Variable 20 20 25 or level Figure 9.3. SBR-cycle configurations: A) Partial nitritation reactor and B) anammox reactor. Details about the variable phase length are provided in Table 9.2. The discharge pump from the nitritation unit directly fed the anammox reactor. *The settling time in the partial nitritation reactor from day 0 to 16 was 20 min instead of 40 min. The PN unit was operated for 118 days, from March to July 2018, and it comprised three different Stages according to the aerobic phase duration (Table 9.1). Initially, in Stage I (days 0 - 16) the aerobic reaction phase only lasted 10 min resulting in a short cycle duration of 145 minutes. This cycle length was imposed to reach a high hydraulic load at a hydraulic retention time (HRT) of 4.8 hours, to promote the growth of AOB over the NOB ones (as it was shown in Chapter 4). During this Stage I, the settling time was 20 min and the volume exchange ratio (VER) was set at 50 %, in order to select the AOB while NOB are washout. Moreover, the length of the total reaction phase (including feeding) was of 100 min. Later, in Stage II (days 17 - 50) the settling time increased from 20 to 40 min to improve biomass retention. During this Stage II, the length of the aeration phase was first extended to 150 min to promote the ammonium oxidation and then stepwise reduced to 90 min on day 21 to be finally set at 60 min from day 25 onwards. The cycle length was manually adjusted in order to obtain an effluent with the nitrite to ammonium ratio close to the required value for the anammox process of 1.32 g NO2--N/g NH4+-N. The resulting HRT ranged from 10.2 to 7.2 hours (Table 9.1). Finally, during Stage III (days 51 to 118), the VER was diminished to 33 % and the SBR cycle configuration was changed to one with variable duration based on a pH set-point that terminates the aeration Testing a two-stage partial nitritation and anammox pilot-scale plant implemented in a WWTP 313 phase. A maximum aeration time of 120 min was set in order to finish the cycle in case that pH-set point was not reached due to pH probes failure or process instabilities. The objective of this cycle variable duration was to adjust the nitrite to ammonium ratio in the effluent to feed it to the posterior AMX unit. As the wastewater composition and temperature fluctuated, the pH-set point was periodically readjusted according to the performed periodic single cycle characterisation. In all stages the effluent withdrawal lasted 25 minutes or till the required remaining liquid level inside the reactor was reached. Table 9.1. Different operational stages distribution for the pilot-scale PN reactor. Stage Duration (days) Cycle length (min) Aerationa (min) Settling (min) VER (%) HRT (h) S - I 0 -16 145 10 20 50 4.8 S - II 17 -49 310 – 215b 150 - 90 (day 21)- 60 (day 25) 40 50 7.2 - 10.2 S - III 50 -118 140 – 275c 120 min or pH 40 33 7.5 - 13.9 a In all cases the previous 90 min was used for the simultaneous feeding + aeration stage. b Cycle length manually adjusted. c Cycle length variable because the aeration phase was of 120 min (maximum) or finalised according to the defined pH set point. 9.3.2.2. Anammox reactor Once the partial nitritation process was stable, on day 69 of operation (Stage III), the AMX unit (Figure 9.1.C) was started-up and operated for 48 days simultaneously with the PN unit. The anammox reactor was an open SBR provided with mechanical stirring (Figure 9.2). The VER was equal to 33 %, the same as the one from the PN unit, and the discharge pump from the partial nitritation unit directly fed the anammox reactor. Thus, the SBR cycles were coupled. The feeding phase lasted 25 minutes (coinciding with the discharge from the PN unit) (Figure 9.3). Then, the reactor was mixed during at least 90 min and the reaction time terminated by stopping the mechanical stirring when the defined set-point was reached on the previous PN unit (Figure 9.3.B). Thus, the reaction time in the anammox reactor was not optimised by itself. The biomass settled for 20 min and then the effluent was discharged by till the defined liquid level. The reactor was equipped with ammonium, nitrate (AN-ISE, Hach), conductivity and redox sensors connected a sc100TM Chapter 9 314 controller (Hach) for data acquisition and to the SCADA software for anammox process monitoring. The anammox reactor was inoculated with anammox-enriched granular biomass from the full-scale ELAN® reactor located in the Guillarei WWTP (Tui, NW Spain) (Morales et al. 2018) (Figure 9.4). Figure 9.4. Image of the inoculum used for the anammox pilot-scale reactor. 9.3.3. Analytical methods Influent and effluent streams were periodically sampled to follow the pilot plant performance and to calibrate the online sensors. Single operational cycles were monitored to evaluate the evolution of the concentrations of the different compounds in the liquid phase in order to assess the process performance and to adjust the pH set point in the PN unit. Before analysis samples were filtered through 0.45 µm pore size. The concentration of chemical oxygen demand (COD), ammonium, total phosphorous, total nitrogen, nitrite and nitrate were determined spectrophotometrically using Dr. Lange test kits (Hach Lange, Germany). Alkalinity was determined by titration according to Standard Methods for Water Examination as well as the concentration of the total (TSS) and volatile suspended solids (VSS) and sludge volume index (SVI) (APHA-AWWA-WEF, 2012). In order to obtain a homogenous and representative sample to determine the solid concentrations in the effluent, approximately 2 L of this stream were collected each 5 min, throughout the length of the discharge phase. Moreover, pH, temperature, conductivity and DO concentration probes were also used to corroborate the online measurements. The maximum specific nitrifying activities (both for AOB and NOB, SAAOB and SANOB respectively) were determined by respirometric tests according to Lopez-Fiuza et al. (2002). All the methodologies used are described in detail in Chapter 2. [Document text truncated for crawler view.]