Wastewater treatment with aerobic granular sludge : challenges and opportunities for modelling and off-gas analyses

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1

No lies were written in this book, though I cannot guarantee that it tells the truth

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Promotors

prof. dr. ir. Eveline Volcke

Department of Green Chemistry and Technology, Faculty of Bioscience Engineering Ghent University

Coupure Links 653, 9000 Gent, Belgium Eveline.Volcke@ugent.be

prof. dr. ir. Mark van Loosdrecht

Department of Biotechnology, Faculty of Applied Sciences Delft University of Technology

Van der Maasweg 9, 2629 HZ, Delft, the Netherlands M.C.M.vanLoosdrecht@tudelft.nl

Examination board

prof. dr. ir. Korneel Rabaey (Chairman)

Department of Biotechnology, Faculty of Bioscience Engineering Ghent University, Belgium

prof. dr. ir. Christophe Walgraeve (Secretary)

Department of Green Chemistry and Technology, Faculty of Bioscience Engineering Ghent University, Belgium

dr. ir. Elena Torfs

Department of Data Analysis and Mathematical Modelling, Faculty of Bioscience Engineering Ghent University, Belgium

prof. dr. ir. Jan Dries

Biochemische Afvalwater Valorisatie & Zuivering University of Antwerp, Belgium

dr. ir. Sylvie Gillot UR Reversaal

French National Research Institute for Agriculture, Food, and Environment (INRAE), France ir. Edward van Dijk

Royal HaskoningDHV, The Netherlands Dean

Prof. dr. ir. Marc Van Meirvenne Rector

Prof. dr. ir. Rik Van de Walle

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W ASTEWATER TREATMENT WITH AEROBIC GRANULAR SLUDGE :

CHALLENGES AND OPPORTUNITIES FOR MODELLING AND OFF - GAS ANALYSES

Thesis submitted in fulfilment of the requirements for the degree of Doctor (PhD) in Bioscience Engineering: Environmental Sciences and Technology

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Dutch translation of the title

Waterzuivering met aëroob korrelslib: uitdagingen en mogelijkheden voor modellering en afgasanalyses

Please refer to this work as follows

Baeten, J.E. (2020). Wastewater treatment with aerobic granular sludge: challenges and opportunities for modelling and off-gas analyses. PhD thesis, Faculty of Bioscience Engineering, Ghent University, Ghent, Belgium, pp. 254.

Financing institution

The doctoral research work of Janis Baeten has been financially supported by the Research Foundation - Flanders (FWO) through a PhD fellowship.

Cover

Photo of bubbles in water by Tsiumpa downloaded and edited from Adobe Stock.

Printing: University Press, Zelzate Copyright: 2020 Janis Baeten ISBN: 9789463573221

All rights reserved. No part of this thesis may be reproduced, stored in a retrieval system of any nature, or transmitted in any means, without permission of the author, or when appropriate, of the publishers of the publications.

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Acknowledgements

Who else but Eveline, my promotor, can I praise first? From the start of my PhD, you gave me the feeling that you trusted my work. You showed this by motivating me to publish, present at conferences, record instructional videos, give lectures, collaborate with your (international) contacts etc., even when I felt insecure myself. While you seemed to trust that my PhD would be allright in the end, you asked me a thousand questions and gave critical remarks during the research. Through this combination of personal trust and professional scepticism, you taught me to appreciate every question and remark as an opportunity to learn, instead of interpreting it as an attack. This valuable lesson has helped me to improve this thesis and it will help me wherever I go. I enjoyed the journey towards a whole lot of new knowledge and hopefully also a little bit of wisdom under your supervision. Mark, my second promotor, also possesses the wonderful virtue of trust and generosity. You let me stay in your research group for two months, connected me with interesting people and shared your considerable knowledge and insight. After all our meetings, mails and tweeting, I can only conclude that you are as open and versatile in your scientific work as you are on a personal level. You both truly ‘promote’, in both senses of the word.

“to encourage people to like, buy, use, do, or support something”

“to raise someone to a higher or more important position or rank”

Cambridge dictionary The comments of the examination board of this thesis were highly appreciated. They ranged from technical details to broad philosophical discussion topics and everything in between. Some of the questions I had never thought of before, which helped me to clarify and more critically assess my results, and which helped to define new ideas for further research.

As such, the jury was an important driving force in the last stage of writing.

My colleagues from UGent, thank you for helping me with my research and for the friendly atmosphere at work. Especially Luis, thanks for the many philosophical, inspiring, silly, depressing and open talks during our regular walks, for the board games we played and for sharing music, drinks, ideals and sorrows. I also want to thank Celia for sharing her wisdom, Thomas for our brief but kind encounter, Caroline for the comforting talks and hugs, Kris for his follow-up and always remembering my birthday, Mingsheng for his very open and tangibly kind mind, Quan for his exemplary calm presence, Tinne for her frank speech and liveliness, Annelies for her exquisite talent for listening and the pleasant co-teaching sessions, Stijn for his youthful energy and sharing his considerable historical and political knowledge, Laurence for her idealistic enthusiasm and the nice talks during our juggling and walking sessions, Kim for her ability to say nothing but things that are helpful, kind or funny, Thiago for his southern

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passion and eastern wisdom and his concern for the world, Farhang for sharing his radically liberal view on society and life and David for demonstrating how to live (and eat) truly authentically. I am also grateful to Christophe for letting me perform methane oxidation tests in his lab. I am grateful to both Christophe and Joren for sharing their problem solving skills in the analytical lab.

I also thank the people from TU Delft and Royal haskoningDHV, which I met during my research stay in the Netherlands and the meetings that followed. In particular Edward and Mario, you were an inspiring duo, especially during my stay in The Netherlands, but also after that, during meetings, conferences and e-mail communication. Your seemingly endless enthusiasm for aerobic granules was highly contagious. Whenever I felt a bit discouraged, even a quick talk with one of you could again spark my curiosity about the complex behaviour of these aggregates. I am also particularly grateful to Andreas for the trust in our collaboration.

I want to express my gratitude towards Mark for the help at the treatment plant and Suellen and Udo for the help with the off-gas measurements. Danny and Ingrid, thanks for the nice moments of sharing drinks, food, music, places to sleep and thoughts. I also thank Damien Batstone from the Advanced Water Management Centre and Oliver Schraa from inCTRL solutions for the fruitful collaboration to review existing granular sludge models.

Then of course, I would not have had the courage, inspiration nor the motivation to perform PhD research without the mental support and love of my family and friends. I can hardly thank my parents, brother, extra parents and extra sisters and their families enough for their unconditional love and support for the choices I make and the work I do. I can only hope that I make you feel as supported and loved as you make me feel. A big thank you to the Solid Spacers, our practices and gigs in the first part of my PhD and our less musical excursions during the second part were the greatest kind of escapism. Especially Pieter and Ella, thanks a million for making me godfather of your lovely son Menno, which is an honour and a great motivation to work in the field of environmental technology. Klara and Fran, you were dear friends and sports buddies in Antwerp and Ghent respectively. Without you, I would have felt much lonelier and heavier, both in the literal and figurative sense. Then my roomies, Sofie, Laurien, Lien, Sara it was great living together! We managed to convert our house into something I could call our home. Eva, thanks for being such a good friend, I always enjoyed our inspiring talks, dinners and meditation sessions very much. Leen, I loved to Fonkel with you, game with you, hike with you, swim with you and save the world with you via De Groene Locomotief. Yael, Matthijs, Andrea, Milan, Rozanne, Simon, Renée and Sam, thanks for sticking with me even though I was too stubborn to buy a smartphone with WhatsApp. Susan, thanks for your constant reminders that the scientific way of thinking is just one possibility, by taking me with you to theatre plays, recommending books (which I mostly didn’t read), cooking for each other, being lazy together and letting me believe that I help you to solve Karel’s Crypto.

Laura and Jelle, thanks for our epic dancing and cinema dates. To my friends from the UA,

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trips where I could temporarily forget all my sorrows.

After listing all these people I owe so much, I realize even more that I am just a single wave on a vast ocean. To acknowledge all the people that came before me, especially scientists and other truth seekers, I quote an early Greek natural philosopher.

“Ever-newer waters flow on those who step into the same rivers”

Heraclitus Originally, it was probably meant quite literally, to express that all things are constantly changing in nature: seeds become trees, clouds become rain and people become ash. For me, it also expresses the way that every new-born has to go through a lot of the same obstacles in life, while every life is still unique. In this same way, I am certainly not the first to perform PhD research, but no one has ever written this specific doctoral thesis before. Thirdly and lastly, this quote about water seemed appropriate since this thesis is about water. So let us plunge into it! Oops, I forget to mention, it is wastewater we will be plunging into …

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List of symbols

Symbol Definition Unit

t Time d

z Space coordinate m

C Concentration g.m^-3

ΔC Driving force for stripping g.m^-3

S Concentration of solutes g.m^-3

X Concentration of particulates g.m^-3

XC Concentration of particulates and colloidals g.m^-3

K Intrinsic half-saturation coefficient g.m^-3

Kapp Apparent half-saturation coefficient g.m^-3

x Mole fraction mole.mole^-1

RH Relative humidity %

M Molecular mass g.mole^-1

p Pressure Pa

T Temperature K

R Ideal gas constant J.mole^-1.K^-1

g Gravitational acceleration m.s^-2

V Volume m³

ρ Density kg.m^-3

ε Porosity of biofilm or gas hold-up in a reactor m³.m^-3

A Area m²

H Height m

δ Radius m

Q Volumetric flow rate m³.d^-1

u Advective velocity m.d^-1

m Mass g

ṁ Mass flow rate g.d^-1

Ṙ Reaction rate g.d^-1

R Reacted mass g

ṙ Volumetric reaction rate g.d^-1.m^-3

q Specific reaction rate g¹.g^-1.d^-1

µ Specific growth rate d^-1

b Decay rate d^-1

η Rate reduction factor depending on redox conditions -

θ Temperature correction factor -

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D Diffusion coefficient m².s^-1 Dr Relative diffusion coefficient compared to water -

h Henry solubility coefficient g.m^-3 in the

liquid per g.m^-3 in the gas

Y Growth yield g¹.g^-1

f Fraction g¹.g^-1

i Content g¹.g^-1

τ Pure time delay d

d Specific waste rate g.g^-1.d^-1

fVER Volume exchange ratio m³.m^-3

SRT Solids retention time d

HRT Hydraulic retention time d

KLa Overall volumetric liquid-gas transfer coefficient d^-1

OTE Oxygen transfer efficiency %

AE Aeration efficiency kg O2.kWh^-1

Subscripts

i Index referring to a substance, microbial group or compartment, e.g.

oxygen, COD, water, AOO or the second compartment out Leaving the reactor

in Entering the reactor reactor Inside the reactor

eq Equilibrium value

atm Atmospheric

BM Biomass

max Maximum

tot Total

lim Limiting substrate

SS Steady State

app Apparent

20° Value corrected to a temperature of 20°C Superscripts

L Liquid phase

G Gas phase

L-G From the liquid to the gas

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List of abbreviations

Abbreviation Definition

AOO Ammonia Oxidizing Organisms NOO Nitrite Oxidizing Organisms OHO Ordinary Heterotrophic Organisms PAO Polyphosphate Accumulating Organisms GAO Glycogen Accumulating Organisms

PHA Polyhydroxyalkanoates (storage compound in PAO and GAO) PP Polyphosphate (storage compound in PAO)

COD Chemical Oxygen Demand

N Nitrogen

P Phosphorus

TSS Total Suspended Solids

F Fermentable organic matter

VFA Volatile Fatty Acids (fermentation products) NHx Ammonium and ammonia (NH3+NH4+) NOx Nitrate and nitrite (NO3-+NO2-)

U Undegradable organics

B Biodegradable organics

CFD Computational Fluid Dynamics ASM Activated Sludge Model BSM Benchmark Simulation Model

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Summary

An average Belgian person produces 150 litres of wastewater per day (Aquafin, 2020). This poses a risk to human health and ecosystems if discharged directly into surface waters. To reduce these risks, wastewater is often treated before discharge. To this end, biotechnologies can be used which rely on microorganisms that consume the harmful substances in wastewater. In conventional aerated systems, these microorganisms grow together in loose flocs (Henze et al., 2008), but a recent breakthrough technology makes them grow in dense, millimetre thick granules. These settle much faster and can therefore be kept in the reactors more easily. This aerobic granular sludge technology results in space and energy savings up to 75% and 50% respectively. The distinct design and operation of these new reactors opens up new possibilities for monitoring and control. Yet, the mathematical models that are used for design and optimization of conventional reactors, and which are based on a fundamental understanding of the processes, are not directly applicable for aerobic granular sludge reactors. This thesis investigated the challenges and opportunities of mathematical modelling and off-gas analyses for aerobic granules sludge reactors via a literature review (Chapter II), mathematical modelling and simulation (Chapter III, IV, V and VII) and full-scale measurements (Chapter VI). Apart from new fundamental insights, it paves the road to further minimize energy and space requirements and greenhouse gas emissions and improve effluent quality.

To understand the rationale behind this thesis, Chapter I first explains what wastewater is and why, how and in which installations it is treated with microorganisms. The aerobic granular sludge technology is described and put into context by comparing the feeding and aeration strategy, biomass retention and biological conversions with alternative technologies. The remaining challenges for wastewater treatment are identified and it is envisioned which role mathematical models and off-gas analyses of aerobic granular sludge reactors can play in this respect. The chapter ends with an outline of the thesis, describing the topics and links between its different chapters.

To get an overview of the current challenges and opportunities for mathematical modelling, Chapter II reviews 167 models in literature for granular sludge reactors, including not only aerobic, but also anaerobic and partial nitritation-anammox systems with or without small carriers, as they comprise common phenomena and thus pose similar challenges. A systematic overview of assumptions, goals, scales, software and calibration and validation is made available to accelerate the search for a suitable starting point to tackle specific questions.

Granular sludge reactor models have been used for three types of goals: to gain insight in the relationship between small scale phenomena (e.g. intragranule transport) and large scale operation (e.g. effluent quality), to assess the effect of alternative operational strategies or design and to monitor unmeasured variables. However, one third of the papers did not clearly

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define a goal, which poses a challenge for model development. The observed wide variety of model assumptions is attributed not only to the different goals, but also to different reactor types, a lack of mechanistic understanding and habits within fields of research. Suggestions for further research are given, aiming for better guidelines regarding the required model complexity, for more mechanistic understanding and for optimization of full-scale reactors.

A specific biofilm modelling approach in the software Aquasim was found popular for aerobic and partial nitritation-anammox granular sludge reactors during the review (Chapter II), but one author mentioned that unreliable results arose under some conditions. Chapter III describes, explains and solves the numerical issues that can arise with this approach. The approach constitutes linking several hypothetical compartments with artificial advective flows to represent a single physical well-mixed reactor. It is shown that numerical errors occur when too high artificial advective flows are used, while the intended mixing is not obtained when too low flows are used. The numerical errors arise by a loss of significant digits during the addition of the large, artificial mass flow rates, to the much smaller rates that were actually of interest (the reactions, input and output). A new method based on diffusive links between compartments is proposed instead, which leads to more reliable and faster simulations and is easier to implement.

The review (Chapter II) revealed that biofilm models are popular for aerobic and partial nitritation-anammox granular sludge reactors. Biofilm models explicitly describe the simultaneous diffusion and reactions of substrates inside microbial aggregates and can thus predict microscale concentration gradients. However, apparent conversion kinetics are widely accepted for flocculent sludge and anaerobic granular sludge. As such, complex biofilm models are avoided and the effect of diffusion is simply lumped in the conversion kinetics.

Chapter IV assesses how the simultaneous diffusion and reactions affect the apparent conversion kinetics of aerobic granular sludge and investigates the applicability of such kinetics in models. Using the new biofilm modelling approach from Chapter III, it is shown that also for aerobic granular sludge, the effect of substrate concentrations on the macroscale rates can be described using apparent half-saturation coefficients in Monod expressions.

However, apparent kinetic parameters depend on the microbial population distribution, which can be affected by long-term changes in operating conditions, and on the activity of organisms that compete for the same substrates, which can vary even within one reactor cycle.

Applications of apparent kinetics are suggested based on these findings. As an example, a model for ammonium removal is applied to a full-scale reactor, taking advantage of the sequential operation of aerobic granular sludge reactors for regular recalibration. This could be used to determine the rate-limiting substrate or for model predictive control.

Most granular sludge reactor models neglect any heterogeneity in the gas phase (Chapter II) and the effects of this simplification was not yet know. Therefore, Chapter V investigates the effects of the changing composition and pressure of bubbles while they rise

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(macroscopic) liquid-gas transfer rate, using either the complete gradients of the mole fraction and pressure or a uniform value, for one or both. Through simuluations, it is shown that the effects of the molecular properties, reactor design and operation are highly interactive and often non-linear if both gradients are considered. The gradients are found to affect the total transfer rate, but the degree of these effects strongly depends on the solubility of the substance and driving force for stripping or absorption (i.e. the content in the inlet gas and liquid phase concentration). The effects of both gradients strengthen with an increasing reactor height, which makes the model choice especially relevant for aerobic granular sludge reactors, as they are typically taller than 5 m. A simple procedure is made available to select appropriate assumptions for reactor or plant-wide models. With this procedure, it is demonstrated that some common simplifications lead to significant errors of the predicted transfer rate of oxygen, carbon dioxide, nitrous oxide and nitrogen gas in tall aerobic wastewater treatment reactors. A simple improvement based on the mean composition and pressure is proposed.

Due to the sequential operation of aerobic granular sludge reactors, aerobic reactions occur separated in time from feeding and discharge. This causes a thoroughly mixed liquid phase and simpler mass balances during the reaction phases, which creates new opportunities for monitoring and control. Chapter VI illustrates the potential of off-gas analyses for monitoring and control via a measurement campaign on a full-scale aerobic granular sludge reactor. A single off-gas sampler and analyzer enables derivation of more variables and gives more accurate results for sequentially operated than for continuously operated reactors. The derivation of multiple variables is demonstrated: liquid-gas transfer rates, aeration characteristics (e.g. KLaO2), liquid phase concentrations of emitted substances and conversion rates within every cycle and also the greenhouse gas emissions, influent characteristics (TOC/COD and COD/N ratio) and sludge production over one or more cycles.

Moreover, the results have led to novel insights on the process. A gradual increase of the KLaO2 is observed within cycles and is attributed to surfactants degradation. Besides, carbon dioxide and oxygen transfer rates are correlated due to the coupling of carbon dioxide production and oxygen consumption via biological and chemical conversions. Nitrous oxide often shows two peaks per cycle: one stripping profile immediately after feeding and one coinciding with high nitrification rates. Finally, methane shows a stripping profile immediately after feeding.

Methane oxidation is known to occur in conventional activated sludge reactors, which limits emissions of this greenhouse gas. Chapter VII studies the effect of typical design and operating conditions of aerobic granular sludge reactors on the fate of methane, which was so far unknown. A model with apparent conversion kinetics (Chapter III) and an expression for the liquid-gas transfer rate of methane (Chapter V) shows that a shift from continuous to

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sequential aeration and feeding lowers the percentage of influent methane that is oxidized.

This is due to the accumulation of dissolved methane during unaerated feeding, which favours stripping more than conversion in the preceding aerobic phase. Neither a lower resistance for methane transport, nor a longer residence time of the biomass, nor a taller reactor is expected to stimulate an equally high methane oxidation efficiency as for continuous reactors.

Consequently, it might be better to minimize methane emissions from aerobic granular sludge plants with other strategies, such as polymer extraction from waste sludge, which avoids the uncontrolled methane emissions associated with anaerobic digestion.

Finally, Chapter VIII summarizes the main findings of the previous chapters and gives perspectives for further research and applications of models and off-gas analyses.

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Samenvatting (Dutch Summary)

Een gemiddelde Belg produceert 150 liter afvalwater per dag (Aquafin, 2020). Bij rechtstreekse lozing in oppervlakte water vormt dit een risico voor de gezondheid van mensen en ecosystemen. Om dit risico te verkleinen, wordt afvalwater vaak behandeld vóór de lozing.

Hiertoe kunnen biotechnologieën aangewend worden die micro-organismen aanzetten om de schadelijke stoffen in afvalwater te consumeren. In klassieke beluchte systemen groeien deze organismen samen in vlokken (Henze et al., 2008), maar een recente technologie doet ze in compacte korrels groeien. Deze bezinken sneller dan vlokken en kunnen daarom eenvoudiger in grote getalen in de reactoren gehouden worden. Dit zogenaamd aeroob korrelslibproces neemt tot 75% minder ruimte in beslag en vereist tot 50% minder energie. Het specifieke ontwerp en bedrijf van de reactoren opent nieuwe mogelijkheden voor toezicht en sturing.

Anderzijds zijn de wiskundige modellen die worden gebruikt voor het ontwerp en de optimalisatie van klassieke systemen, en die gebaseerd zijn op een fundamenteel begrip van de processen, hier niet direct toepasbaar. Dit proefschrift onderzocht de uitdagingen en mogelijkheden van wiskundige modellering en afgasanalyses voor aeroob korrelslibreactoren via een literatuurstudie (Hoofdstuk II), wiskundige modellering en simulatie (Hoofdstuk III, IV, V en VII) en metingen op volle schaal (Hoofdstuk VI). Dit levert niet alleen nieuw begrip op, maar het effent ook de weg om de energie-, ruimtevereisten en uitstoot van broeikasgassen verder te minimaliseren en de effluentkwaliteit te verbeteren.

Om de onderzoeksvragen te kaderen, legt Hoofdstuk I eerst uit wat afvalwater is en waarom, hoe en in welke installaties het wordt behandeld met micro-organismen. De aeroob korrelslibtechnologie wordt beschreven en vergeleken met alternatieven met andere voedings- en beluchtingsstrategieën, strategieën om biomassa in de reactoren te houden en biologische conversies. De resterende uitdagingen voor afvalwaterzuivering en de mogelijke rol van wiskundige modellen en afgasanalyses van aeroob korrelslibreactoren worden uiteengezet. Het hoofdstuk beschrijft ten slotte de structuur van het proefschrift en de verbanden tussen de verschillende hoofdstukken.

Om een overzicht te krijgen van de huidige uitdagingen en mogelijkheden van wiskundige modellering, bespreekt Hoofdstuk II 167 korrelslibmodellen uit de literatuur.

Korrelslibreactoren met of zonder kleine dragers worden beschouwd en voor zowel aerobe als voor anaerobe conversies en partiële nitritatie-anammox, aangezien hierin gelijkaardige fenomenen plaatsvinden en de modellering daarom vergelijkbare uitdagingen kent. Om de zoektocht naar een geschikt uitgangsmodel in de toekomst te versnellen, wordt een systematisch overzicht van de veronderstellingen, doelen, schaalgrootte, software en kalibratie en validatie van de 167 modellen beschikbaar gesteld. De modellen worden gebruikt voor drie soorten doelen: voor inzicht in de relatie tussen kleinschalige fenomenen (bijvoorbeeld het stoffentransport ín de korrel) en de grootschalige werking (bijvoorbeeld de

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effluentkwaliteit), om het effect van een alternatieve bedrijfsvoering of ontwerp in te schatten en om niet-gemeten variabelen op te volgen. Een derde van de artikels definieerde echter geen duidelijk doel. Dit vormt een uitdaging voor de verdere ontwikkeling van modellen. De grote verscheidenheid aan modelaannames wordt niet alleen toegeschreven aan de verschillende doelen, maar ook aan de verschillende reactortypes, aan een gebrek aan fundamenteel begrip en aan gewoontes binnen specifieke onderzoeksgebieden. Er worden suggesties gedaan voor verder onderzoek met als doel om betere richtlijnen te verkrijgen voor de benodigde modelcomplexiteit, voor meer fundamenteel begrip en om volle schaal reactoren te optimaliseren.

Tijdens de literatuurstudie bleek een specifieke techniek voor biofilmmodellering met de software Aquasim populair voor aerobe en partiële nitritatie-anammox korrels (Hoofdstuk II). Eén artikel vermeldde echter dat deze techniek onbetrouwbare berekeningen oplevert onder bepaalde omstandigheden. Hoofdstuk III beschrijft, verklaart en lost de numerieke fouten op die bij deze aanpak kunnen ontstaan. De techniek houdt in dat verschillende hypothetische compartimenten worden gekoppeld met kunstmatige vloeistofstromen om één enkele perfect gemengde reactor voor te stellen. Er wordt aangetoond dat numerieke fouten optreden wanneer te snelle kunstmatige stromen worden gebruikt. Aan de andere kant wordt de beoogde perfecte menging niet behaald met te trage stromen. De numerieke fouten ontstaan door een verlies van beduidende cijfers tijdens de optelling van de grote, kunstmatige transportsnelheden bij de veel kleinere snelheden die werkelijk van belang zijn (nl. via reacties, influent en effluent). Een verbeterde methode met artificiële diffusie wordt voorgesteld. Dit leidt tot meer betrouwbare en snellere simulaties en het is eenvoudiger om uit te voeren.

Uit de literatuurstudie (Hoofdstuk II) blijkt biofilmmodellering vooral populair voor aerobe en partiële nitritatie-anammox korrels. Biofilmmodellen beschrijven expliciet de gelijktijdige diffusie en reacties van substraten in microbiële aggregaten en kunnen daarom concentratiegradiënten op deze kleine schaal voorspellen. Nochtans, voor vlokken en anaeroob korrelslib wordt schijnbare conversiekinetiek algemeen aanvaard. Hierbij worden complexe biofilmmodellen vermeden door het effect van diffusie simpelweg te integreren in de vergelijkingen voor de conversiesnelheid via aangepaste parameter waarden. In Hoofdstuk IV wordt d.m.v. de nieuwe techniek uit Hoofdstuk III onderzocht hoe de schijnbare conversiekinetiek van aeroob korrelslib wordt beïnvloed door de gelijktijdige diffusie en reacties en er worden mogelijke toepassingen gedefinieerd. Er werd aangetoond dat ook voor aeroob korrelslib het effect van substraatconcentraties op de totale conversiesnelheden kan worden beschreven met behulp van schijnbare halfverzadigingscoëfficiënten in Monod- vergelijkingen. Deze schijnbare kinetische parameters hangen echter af van de microbiële populatieverdeling in de korrels en die wordt beïnvloed door langdurige veranderingen van de condities in een reactor. Ook de activiteit van organismen die concurreren voor dezelfde

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substraten beïnvloedt de schijnbare conversiekinetiek, en deze kan zelfs binnen één reactorcyclus variëren. Er worden toepassingen van schijnbare kinetiek voorgesteld op basis van deze bevindingen. Als voorbeeld wordt een model voor ammoniumverwijdering toegepast op een volle schaal reactor, door gebruik te maken van de cyclische werking van de reactoren voor regelmatige herkalibratie. Dit zou kunnen dienen om het snelheidslimiterende substraat te identificeren of voor modelgebaseerde controle.

Meestal wordt de heterogeniteit in de gasfase genegeerd en de effecten van deze vereenvoudiging waren tot nu toe onbekend (Hoofdstuk II). Daarom onderzoekt Hoofdstuk V de effecten van de veranderende samenstelling en druk van bellen terwijl ze van de bodem naar de top van een reactor stijgen. Analytische uitdrukkingen voor de totale (macroscopische) vloeistof-gas overdrachtssnelheid worden afgeleid, zowel door de volledige gradiënten van de molfractie en druk te beschouwen als door één of beide uniform te veronderstellen. Via simulaties wordt aangetoond dat de effecten van andere moleculaire eigenschappen en een ander reactorontwerp en –bedrijf interactief en vaak niet-lineair zijn wanneer beide gradiënten worden beschouwd. De gradiënten beïnvloeden weldegelijk de totale overdrachtssnelheid, maar de intensiteit van dit effect hangt sterk af van de oplosbaarheid van de stof en de drijvende kracht voor strippen of absorptie (d.w.z. het gehalte in het geïnjecteerde gas en in de vloeistof). De effecten van beide gradiënten worden versterkt voor hogere reactoren. De modelkeuze is dus extra belangrijk voor aeroob korrelslibreactoren, aangezien deze meestal hoger zijn dan 5 m. Er wordt een procedure ter beschikking gesteld om snel adequate aannames te selecteren voor toekomstige modellen. Met deze procedure wordt aangetoond dat enkele veelvoorkomende vereenvoudigingen een afwijkende overdrachtssnelheid voorspellen van zuurstof, koolstofdioxide, stikstofoxide en stikstofgas in hoge, beluchte afvalwaterzuiveringsreactoren. Een eenvoudige verbetering op basis van de gemiddelde samenstelling en druk wordt voorgesteld.

Het cyclische bedrijf van aeroob korrelslibreactoren scheidt de aerobe reacties van de toevoer en afvoer van stoffen. Dit creëert nieuwe mogelijkheden voor monitoring en sturing doordat de vloeistoffase goed gemengd is en de massabalansen van stoffen eenvoudiger zijn tijdens de reactiefasen. Hoofdstuk VI illustreert het potentieel van afgasanalyses voor monitoring en sturing via een meetcampagne op een volle schaal reactor. Eén enkele opstelling met monstername- en analyseapparatuur laat hier toe om meer variabelen op te volgen en geeft nauwkeurigere resultaten dan voor continu bedreven reactoren. De afleiding van meerdere variabelen wordt gedemonstreerd: niet enkel vloeistof-gas overdrachtssnelheden, beluchtingseigenschappen (bv. KLaO2), opgeloste concentraties van uitgestoten stoffen en reactiesnelheden binnen elke cyclus, maar ook broeikasgasemissies, influentkarakteristieken (TOC/COD en COD/N verhouding) en slibproductie over één of meer cycli. De resultaten leiden ook tot nieuwe inzichten in het proces. Een geleidelijke toename van KLaO2 binnen de cycli wordt toegeschreven aan de afbraak van oppervlakte-actieve

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stoffen. De overdrachtssnelheden van koolstofdioxide en zuurstof is gecorreleerd vanwege de koppeling van koolstofdioxideproductie en zuurstofverbruik via biologische en chemische reacties. Stikstofoxide vertoont vaak twee pieken per cyclus met de specifieke beluchtingsstrategie die werd toegepast: één stripprofiel onmiddellijk na het voeden en één piek die samenvalt met hoge nitrificatiesnelheden. Ten slotte vertoont methaan onmiddellijk na het voeden een stripprofiel.

Het is geweten dat methaanoxidatie optreedt in klassieke actief slibreactoren, wat de uitstoot van dit broeikasgas beperkt. Hoofdstuk VII bestudeert het lot van methaan uit de riolering wanneer men overschakelt naar het typische ontwerp en bedrijf van aeroob korrelslibreactoren, aangezien dit tot nu toe onbekend was. Een model met schijnbare conversiekinetiek (Hoofdstuk III) en een uitdrukking voor de vloeistof-gasoverdrachtssnelheid van methaan (Hoofdstuk V) laat zien dat een verschuiving van gelijktijdige naar gescheiden beluchting en voeding minder methaanoxidatie veroorzaakt. Dit komt door de accumulatie van opgelost methaan tijdens niet-beluchte voedingsfasen. Dit stimuleert het strippen meer dan de oxidatie tijdens de beluchtingsfase die hierop volgt. Noch een lagere weerstand voor methaantransport in de korrels, noch een langere verblijftijd van de biomassa, noch hogere reactoren kunnen de methaanoxidatie-efficiëntie stimuleren tot de hetzelfde niveau als bij continu bedrijf. Hierdoor lijkt in-situ oxidatie geen haalbare strategie om methaanemissies van aeroob korrelslibinstallaties te minimaliseren. Andere strategieën worden voorgesteld, zoals polymeerextractie uit afvalslib, aangezien dit de ongecontroleerde methaanemissies vermijdt die gepaard gaan met anaerobe vergisting.

Hoofdstuk VIII vat de belangrijkste bevindingen van de vorige hoofdstukken samen en geeft een aanzet voor verder onderzoek en toepassingen van modellen en afgasanalyses.

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Chapter I: General introduction

What is wastewater and why is it treated?

Wastewater is a compound word combining ‘waste’ and ‘water’. According to European legislation (2008/98/EC), a substance or object is called ‘waste’ if the holder discards it, intends to discard it or is required to discard it. In this definition, the action or even just the intention of the holder determines whether something is waste or not. This implies that there is no inherent chemical, physical or biological, but rather a social and psychological difference between ‘normal’ and ‘waste’ substances or objects. For example, if you would now decide that you will throw this thesis away, it would suddenly become waste. If waste ends up in water, this mixture is called wastewater. Also here, it is hard to give a rigorous scientific definition so one has to resort to legislation, which distinguishes two types of wastewater. The first type is ‘domestic wastewater’. This comes from residential settlements and services and originates predominantly from the human metabolism and from household activities (91/271/EEC). More precisely, it consists of tap water combined with faeces, urine, toilet paper, sweat, cleaning products, food residues, pharmaceuticals, health care products and other wastes, together with rainwater and groundwater that intentionally or unintentionally entered the sewers. A second type is ‘industrial wastewater’, which is wastewater discharged from premises used for carrying on any trade or industry, other than domestic wastewater and run- off rain water (91/271/EEC). The variety of wastes it can contain is even larger, ranging from blood and fruit fibres to synthetic dyes and oil, depending on the type of industry. Municipal wastewater or sewage is often a combination of domestic and industrial wastewater.

Given the immense diversity of waste objects and substances that can be present in wastewater, it is impractical to determine and list the complete composition of a specific stream. Therefore, its ecological harmfulness, toxicity, corrosiveness and/or nuisance potential is usually characterised via a limited amount of standardized measurements of concentrations of chemical substances or elements, of reactiveness and of physical properties.

The arguably most important characteristics are the Chemical Oxygen Demand (COD), Nitrogen (N) and Phosphorus (P) content. COD is determined by adding a strong chemical oxidant (e.g. dichromate) and measuring how much of it is consumed by reacting with the substances in the water, which explains the term ‘chemical’. This reactiveness is then expressed as the amount of oxygen that would have been required to perform this same amount of oxidation (g O2.m^-3) (Henze et al., 2008), hence the term ‘oxygen demand’. It is a proxy for the amount of organic pollutants, e.g. fatty acids, carbohydrates, proteins, DNA, hydrocarbons etc. The COD content gives an indication of how strongly a wastewater would lead to oxygen depletion if it would end up in a river, ocean, lake or other natural water body.

As a part of this COD cannot be biologically degraded, the biodegradable oxygen demand

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(BOD) is a more precise indicator for this harmful effect, which is often estimated by measuring the oxygen consumption by microorganisms in a wastewater sample over 5 days (BOD5) (Roeleveld and van Loosdrecht, 2002). Moreover, if a high fraction of the COD is caused by the presence of toxic organic substances like phenols, the COD value is also indicative of the direct toxicity.

The N and P content are simply the concentration of these chemical elements in the water, which can be present in different molecules and ions. For N, the most important fractions are inorganic ammonium/ammonia (NH4+/NH3), nitrate (NO3-), nitrite (NO2-) and organically bound N, e.g. in proteins or urea. Similarly, P is mostly present as inorganic phosphate, but it is also bound in organic compounds (Henze et al., 2008) such as phospholipids (important constituents of cell membranes), RNA and DNA (Wade and Simek, 2017). N and P are important nutrients for plants and algae and consequently their contents are primarily indicators of how strongly a wastewater can contribute to eutrophication of natural waters, which is an excessive growth of algae which disturbs the natural ecosystem (Smith and Schindler, 2009). However, there is also direct toxicity for many species including humans, especially by ammonium and nitrite(Philips et al., 2002).

As indicated in the previous paragraphs, wastewater contains many substances which can disturb ecosystems when discharged directly into natural waters. Using the word

‘ecosystems’ may give the impression that humans would stay unaffected, but this is not true.

For example, if eutrophication occurs in a natural water body, it can be dangerous to swim there due to toxic algal species. Besides, a decreasing fish population can make fishing less lucrative, odour nuisance can occur, the aesthetic value of the water body can decrease and drinking water production from this source becomes more complicated, all of which has economic implications as well (Smith, 2003). Human health is also directly at risk when too many pathogens and toxic substances from wastewater end up in the environment, from where they can come in contact with humans via irrigation of food crops, via drinking water production from natural waters or by swimming (Crockett, 2007, Delli Compagni et al., 2019).

To reduce these risks, wastewater can be treated, meaning that harmful objects, substances and organisms are removed from it.

How is wastewater treated biologically?

Since the late 19th century, wastewater treatment heavily relies on microorganisms, primarily bacteria. Microorganisms which can convert the harmful substances to less harmful ones are naturally present in wastewater (Henze et al., 2008). This section lists the most important types and the main reactions that they can catalyse and discusses the most popular method to remove the excess microbial biomass that is inevitably cultivated during treatment.

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2.1. Organic matter (COD) removal

Chemo-organo-heterotrophs are organisms that can degrade organic substances (COD) to obtain energy while incorporating the carbon atoms from these substances into cellular compounds. The type that is of most importance for this thesis are aerobic heterotrophic microorganisms. They can degrade organics with oxygen as electron acceptor (a typical combustion reaction), while incorporating some carbon from the organics together with other nutrients into new biomass (Eq. 1.1) (Henze et al., 2008).

COD + O2 + other nutrients → CO2 + H2O + heterotrophic biomass Eq. 1.1 It can be seen that potentially harmful COD is converted into water, which is harmless, carbon dioxide, which is a gas and thus spontaneously leaves the water over time, and some microbial biomass, which still requires removal from the water (see section 2.4). This metabolism is in essence the same as that of humans. We eat organic substances (fats, sugars and proteins) and create energy by burning them with the oxygen we breathe in, while we use a part of the organics to build new cellular material. Some aerobic heterotrophs can first store the organics intracellularly and oxidize them later on (see section 2.3) (Mino et al., 1998).

Some chemo-organo-heterotrophs can also use nitrite or nitrate as electron acceptor for COD degradation and are called denitrifiers (see section 2.2). There are also heterotrophs that can degrade organic compounds under anaerobic conditions, i.e. without oxygen or nitrate/nitrite, by using iron, sulphate or even another organic compound as electron acceptor (Henze et al., 2008). The latter, which are termed fermentative and methanogenic heterotrophs, degrade larger compounds like glucose to smaller ones like acetate and eventually to methane, while simultaneously releasing carbon dioxide. Both methane and carbon dioxide escape as gases, thus cleaner water is also obtained in this case.

2.2. Nitrogen (N) removal

Microorganisms that degrade inorganic substances to obtain energy while incorporating the carbon atoms from inorganic carbon (CO2) into cellular compounds are called chemo-litho-autotrophs. The specific type that can aerobically oxidize ammonium to nitrite via the so-called nitritation reaction are Ammonia Oxidizing Organisms (AOO) (Eq. 1.2), while the ones that oxidize nitrite further to nitrate via the nitratation reaction are called Nitrite Oxidizing Organisms (NOO) (Eq. 1.3) (Corominas et al., 2010, Henze et al., 2008).

NH4+ + O2 + CO2 + other nutrients → NO2- + H2O + AOO biomass Eq. 1.2 NO2- + O2 + CO2 + other nutrients → NO3- + H2O + NOO biomass Eq. 1.3 The complete oxidation of ammonium to nitrate is called nitrification. Recently, so-called comammox organisms have been discovered which can perform both sub-steps (van Kessel

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et al., 2015). Overall, nitrification converts ammonium into the ecologically less harmful nitrate, but some much more harmful nitrite can remain (Camargo and Alonso, 2006). Denitrifiers (section 2.1) can remove the residual toxicity by nitrate and nitrite by reducing these compounds with organics as electron donor (Eq. 1.4).

COD + NO3-/NO2- + nutrients → CO2 + H2O + N2 + heterotrophic biomass Eq. 1.4 As such, nitrate and nitrite are converted to harmless nitrogen gas, while organic pollutants are removed.

Even though nitrification and denitrification are the most commonly applied pathways for N removal, there is an alternative. Anaerobic Ammonium-Oxidizing (Anammox) organisms, which are again chemo-litho-autotrophs, can reduce nitrite with ammonium as electron acceptor (Eq. 1.5), with some nitrate as a byproduct (Jetten et al., 1998, Strous et al., 1999).

NH4+ + NO2- + CO2 + other nutrients → N2 + H2O + NO3- Anammox biomass Eq. 1.5 This reaction thus converts two highly toxic N compounds to primarily nitrogen gas directly, which decreases the required oxygen for nitrification (Eq. 1.2 and Eq. 1.3) and does not need organics, opposed to denitrification (Eq. 1.4). Moreover, autotrophs such as Anammox, form less biomass (Eq. 1.5) for the same amount of nitrogen removal, so less waste is produced compared to denitrification (Eq. 1.4).

2.3. Phosphorus (P) removal

P can be removed from wastewater by a specific type of aerobic heterotrophic microorganisms called Polyphosphate Accumulating Organisms (PAO). These have a cyclic metabolism based on the alternating degradation and storage of three different intracellular storage polymers. Under anaerobic conditions (no oxygen, nitrate or nitrite), they are able to store organics from the wastewater intracellularly as polyhydroxyalkanoates (PHA). They obtain the required energy for this storage by degrading other storage polymers, namely polyphosphate (PP) and glycogen. Degradation of PP causes phosphate release into the water (Eq. 1.6).

COD + PP + glycogen → PHA + PO42- Eq. 1.6

Under aerobic or anoxic conditions (in the presence of oxygen or nitrate/nitrite), they are able to oxidize the PHA that was stored earlier. At the same time, they take up phosphate to restore the intracellular PP and synthesize new glycogen (schematically shown in Eq. 1.7).

PHA + O2(NO3-/NO2-) + PO42- + other nutrients →

CO2 + H2O (+ N2) + PAO biomass + PP + glycogen Eq. 1.7 This equation shows that PAO can remove phosphate, organics (under the form of intracellular PHA) and nitrate/nitrite, which makes them highly versatile and useful for wastewater treatment. To obtain a true removal of phosphate, PAO biomass containing the

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stored PP should be removed from the wastewater. Otherwise, it can be released again (Mino et al., 1998, Santos et al., 2020). In the next section, it is explained how the P-rich PAO biomass can be removed from wastewater along with the other heterotrophic and autotrophic biomass.

2.4. Biomass separation

Until now, all removal mechanisms that were discussed were based on the capability of different microorganisms to remove different pollutants. However, when these microorganisms are fed, they unavoidably multiply (see Eq. 1.1-Eq. 1.7). This cultivated biomass could still pollute natural waters, e.g. via its COD, N and P content under the form of cellular compounds. Since biomass consists of cells, which are typically heavier than water, they can automatically separate from the water by gravity. This settling is speeded up thanks to the tendency of the cells to aggregate as flocs (Farnsworth and Dick, 1972, Wahlberg et al., 1994). The next section explains how this process is stimulated in wastewater treatment plants.

Where is wastewater treated?

Biological wastewater treatment plants are engineered environments which stimulate the growth of the purifying microorganisms that are naturally present in wastewater (section 2) without letting the grown microorganisms exit to the receiving water body. Several physical or chemical processes can be used to enhance, supplement or sometimes substitute this primary function.

3.1. Conventional activated sludge plants

The core of a wastewater treatment plant with biological removal of COD, N and P is a series of reactors which contain activated sludge, a mixture of purifying bacteria and other solids originating from the wastewater, followed by a settling tank to avoid discharge of this sludge to the receiving water body. The most basic configuration for biological removal of these three types of pollution consists of three reactors or compartments (Henze et al., 2008) (Figure 1.1). The first one receives the influent wastewater and combines it with return sludge from the settler. This unaerated compartment hardly contains any oxygen or nitrate/nitrite and is therefore called anaerobic. Such conditions allow intracellular storage of COD by PAO (Eq.

1.6). Next, the mixture of wastewater and sludge (called mixed liquor) is successively sent to an unaerated and aerated compartment, which are coupled with a recycle. In the aerated compartment, air bubbles are constantly injected to get oxygen into the water. Here, aerobic removal of the remaining COD by heterotrophs takes place, both by PAO (Eq. 1.7) and by Ordinary Heterotrophic Organisms (OHO) (Eq. 1.1). While PAO oxidise their intracellular COD storage, phosphate is stored in turn, removing it from the wastewater. Also nitrification by AOO

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and NOO takes place under these aerobic conditions (Eq. 1.2 and Eq. 1.3). Part of the nitrate/nitrite that is produced in this manner, is sent to the preceding anoxic reactor via a mixed liquor recycle. Here, the absence of oxygen and still higher abundance of (intracellular) COD allows N removal via denitrification by both PAO and OHO (Eq. 1.4 and Eq. 1.7) (Barnard, 1997). Finally, the mixed liquor is directed to a settler, which is a tank where enough time is provided for the sludge to settle and as such obtain a clear top layer of water, which is suitable for discharge. Since the microorganisms grow as long as wastewater enters, excess sludge has to be regularly wasted from the system, e.g. from the bottom of the settler. The PAO in this stream are the main route for phosphorus removal from the wastewater, since they are wasted soon after the anoxic and aerobic reactors, where phosphate was stored intracellularly (Henze et al., 2008).

The part of a wastewater treatment plant which contains activated sludge can be preceded, augmented and followed by a variety of other physical, chemical and biological processes (Figure 1.1). Pre-treatment primarily aims at the removal of large solid objects and substances like plastic bags and sand which can cause clogging and damage to equipment (Tchobanoglous et al., 2014). Primary treatment via a settler is often installed right before the activated sludge line (called ‘secondary treatment’), to already remove part of the solid pollutants. Tertiary and quaternary treatment can be used to more thoroughly remove pathogens, nutrients or particular toxic substances such as pharmaceuticals. Finally, also the waste sludge can be treated further, e.g. via thickening followed by anaerobic digestion, during which the biomass is partly converted to methane (section 2.1), which can be used as a renewable fuel (Henze et al., 2008).

3.2. Alternative feeding and aeration strategies

While it is conventional to use nearly continuous flows for the influent entering the plant, for the streams between the different tanks and for the aeration air in the aerobic tank (Figure 1.1), the different biological conversions can also be performed in a single reactor instead (Figure 1.2). Sequencing batch reactors operate in cycles, starting with a feeding phase, during which influent enters. For a plant with biological P removal, no aeration should be provided during and possibly also for some time after feeding, to allow PAO to store COD intracellularly (Kuba et al., 1993). Then, the reaction phase is started, which often consists of periods with high and low (or no) aeration to allow nitrification and denitrification respectively.

Then, both feeding and aeration are turned off to allow sludge settling inside the reactor and finally, the top layer of clear water is discharged. As such, a single reactor is used for all biological conversions and settling (Artan and Orhon, 2005).

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Figure 1.1. (A) A basic activated sludge configuration for biological removal of COD, N and P (Barnard, 1997) and (B) a picture of a complete wastewater treatment plant applying a similar configuration in Gdańsk to treat 92 200 m³ of sewage per day, adapted from SNG (2019).

Figure 1.2. A possible cycle configuration of a sequentially operated reactor with biological removal of COD, N and P.

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One advantage of sequentially operated reactors is the flexibility they maintain after construction. For example, the aeration time can be increased to handle changing effluent limits or influent characteristics, while the tank sizes and diffusers in a continuous system fix the maximal aerated volume during construction (Irvine and Ketchum, 1988). Moreover, sequential operation improves the settling characteristics of sludge (Caluwe et al., 2017).

Another advantage is that batch feeding leads to high substrate concentrations, which enables substrate-limited reactions to occur near their maximal rate for part of the cycle time (Irvine and Ketchum, 1988), e.g. high ammonium concentrations after the feeding phase can stimulate nitrification. A related disadvantage is that the limited dilution of the influent can cause higher concentrations of inhibiting substances (Andreottola et al., 1997). Moreover, larger pumps are required because the total amount of wastewater needs to be injected in a shorter amount of time. A final disadvantage is the requirement of several reactors in parallel and/or a buffer tank, to allow treatment of a continuous flow of wastewater (Kent et al., 2018).

3.3. Alternative biomass retention methods

Alternatives to the conventional activated sludge system (3.1) do not only comprise different feeding and aeration regimes (3.2), but also different methods of biomass retention in the system. Instead of settling and recycling the thickened sludge, a membrane can be used to extract the effluent while withholding the sludge. This led to the development of so- called membrane bioreactors, which are more compact due to the better retention of biomass and thus faster reactions (Judd, 2008, Yamamoto et al., 1989). Another alternative is to add inert materials with a high surface area to a reactor to which bacteria can attach as a biofilm.

As such, microbial biomass can be efficiently retained in the system and thus more compact reactors can be obtained. Examples of modern technologies applying biofilms are moving bed biofilm reactors (Odegaard et al., 1994), integrated fixed-film activated sludge systems (Randall and Sen, 1996) and membrane aerated biofilm reactors (Martin and Nerenberg, 2012).

3.4. Alternative biological conversions

Alternatives to the conventional activated sludge system also comprise exposure to a different sequence of conditions (concentrations of COD, N compounds, P and oxygen) by using different configurations of reactors or reactor cycles. As such, different dominant microorganisms (section 2) can be selected. For example, purely anaerobic treatment is often applied to remove COD from municipal wastewater in warm climates and from concentrated industrial streams, but is still challenging for municipal wastewater in colder climates (Stazi and Tomei, 2018). N can be removed via alternative conversions as well, but this is so far only practically applicable on warm, N rich wastewaters, such as the water remaining after

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anaerobic digestion of sludge or industrial wastewaters (Cao et al., 2017, Henze et al., 2008).

One alternative technology for N removal comprises nitritation-denitritation, where influent ammonium is only converted to nitrite (nitritation; Eq. 1.2) and subsequently denitrified (Eq.

1.4 and Eq. 1.7) in the same reactor by regularly turning the aeration off and adding external COD (Hellinga et al., 1998). Another alternative is a partial nitritation-anammox reactor, where influent ammonium is only partly converted to nitrite (partial nitritation) and subsequently removed together with the remaining ammonium via the anammox reaction (Eq. 1.5). Different reactor configurations are possible to obtain these alternative conversions (Lackner et al., 2014).

3.5. Aerobic granular sludge technology Principles of granule formation

The morphology of microbial aggregates like sludge flocs or biofilms depends on the operating conditions in a reactor. If the diffusion of substrates towards the aggregate is rate- limiting, the structure becomes heterogeneous and porous. This is the case if a substrate is consumed faster than it can diffuse from the water to the organism (van Loosdrecht et al., 2002). Diffusion limitation can thus be avoided by speeding up the diffusion process, e.g. by ensuring higher substrate concentrations in the water, or by slowing down the consumption, e.g. by selecting slower growing organisms. One method to do both is using a sequentially operated reactor with an unaerated feeding phase (Figure 1.2). This feeding method favours PAO and Glycogen Accumulating Organisms (GAO) instead of OHO because only they can store the influent COD under anaerobic conditions. Since a large part of the COD is present intracellularly during the aerobic and anoxic phase, no more diffusion is required before its oxidation. Moreover, PAO and GAO grow slower than OHO, with typical maximum specific growth rates of 1 d^-1 (Yagci et al., 2004) versus 6 d^-1 (Henze et al., 2000). The discontinuous feeding also increases the substrate concentrations in general, because the influent is diluted less with the already partly cleaned water in the reactors. Compact and large sludge aggregates, called aerobic granules (Figure 1.3), can be favoured even more if the reactor is also operated with short settling phases during which the slower settling, i.e. heterogeneous and porous, aggregates are selectively wasted from the reactor (de Kreuk and van Loosdrecht, 2004).

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Figure 1.3. Microscopic image of (A) flocculent sludge and (B) granular sludge, adapted from de Kreuk et al. (2005b).

Full-scale application

Aerobic granular sludge has been applied on full-scale in sequentially operated reactors with the selective pressures described in the previous paragraph (RHDHV, 2020).

The commercial name is Nereda® and the reactors are distinguished by their simultaneous feeding and discharge phase. By slowly feeding from the bottom of the reactor via a distribution system, a semi plug flow is obtained, which pushes the upper layer of water out via several effluent weirs with minimal short-circuiting. For this reason, the reactors are taller than most activated sludge reactors, typically between 6 and 9 m (Berkhof et al., 2010, de Bruin et al., 2013). Compared to the typical loose activated sludge flocs, the obtained granules are more dense and structurally stable, leading to faster settling, which allows higher biomass concentrations and thus more compact reactors (Winkler et al., 2018). At the same time, the aerobic and anoxic environment needed for nitrification and denitrification cannot only exist when a sequence of aerated and non-aerated periods is used, but can also exist simultaneously inside the granules when oxygen is consumed before it reaches the core of the granules (de Kreuk et al., 2005a). The removal of COD, N and P in a single tank avoids the energy intensive recycle pumps required in continuously fed plants (Figure 1.1). The aerobic granular sludge technology therefore combines alternative feeding and aeration regimes (section 3.2) with the typical compactness and anoxic niches of biofilm reactors (section 3.3).

Since the first municipal installation in 2008, more and more full-scale Nereda®

reactors have been installed world-wide for both municipal and industrial wastewater treatment. Currently, more than 70 plants with this technology are built or under construction (RHDHV, 2020). The rather slow initial market uptake could be explained by the already existing wastewater treatment plants in many countries, for which replacement was not yet needed. For example, in Flanders, most pants have been built between 1992 and 2008 (VMM, 2018). To make use of the existing installations while increasing the treatment capacity,

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retrofiting to sequentially operated reactors is an option (Pronk et al., 2017), but also the development of granular sludge in continuous flow systems is being investigated (Kent et al., 2018).

Modelling and off-gas analyses for aerobic granular sludge reactors

Even though many sewers and wastewater treatment plants are in place and more and more sustainable technologies, such as aerobic granular sludge, are developed, important challenges and trends in wastewater treatment remain and are discussed in this section. The possible role of modelling and off-gas analyses to tackle some of these challenges is highlighted, focussing on aerobic granular sludge as the research topic of this thesis.

4.1. Remaining challenges and trends and the role of aerobic granular sludge A large amount of wastewater still ends up untreated in natural waters. The fraction of municipal and industrial wastewater that remains completely untreated amounts to 30% in high-income countries, 62-72% in middle-income countries and even 92% in low-income countries (WWAP, 2017). At the same time, continual maintenance and regular replacement of existing sewer and plant infrastructure is needed to maintain the wastewater collection and treatment that is in place (EPA, 2016, Huang et al., 2018). Also the daily operation of existing infrastructure is accompanied with costs for chemicals and electricity required for pumping, aeration, mixing etc. (Piao et al., 2016). Additionally, the well-functioning infrastructure in place has to cope with changes in the wastewater quantity and composition. Short-term variations (e.g. hourly, daily fluctuations, weekly and seasonal fluctuations) can be largely tackled via process control, which is the (automatic or manual) adaptation of the operation to changing conditions, but longer-term influent variations caused by changes in the sewer system, population, lifestyle and precipitation may require construction works to increase the capacity (Henze et al., 2008, VMM, 2018). Moreover, the legal effluent standards can become more stringent with advances in the understanding or awareness of the impact of pollutants (91/271/EEC, 98 15/EEC and 2000/60/EC), which may also require adaptations in infrastructure. In summary, the wastewater sector will require continued investments to maintain and improve the quality of our ecosystems and health.

While the protection of the local population’s health and ecosystems remains the primary goal of wastewater collection and treatment infrastructure, there is a growing pressure to do this with less unintended environmental, societal and economic side-effects (GWRC, 2008). These side-effects can occur during construction, operation and maintenance. For example, construction materials (e.g. cement), energy use and chemicals are associated with