The effect of particulate organic substrate on the formation, composition and performance of aerobic granular sludge

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Doctoral Thesis

The effect of particulate organic substrate on the formation, composition and performance of aerobic granular sludge

Author(s):

Layer, Manuel Publication Date:

2021

Permanent Link:

https://doi.org/10.3929/ethz-b-000474564

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In Copyright - Non-Commercial Use Permitted

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The effect of particulate organic substrate on the formation,

composition and performance of

aerobic granular sludge

DISS. ETH NO. 27048

The effect of particulate organic substrate on the formation, composition and performance of

aerobic granular sludge

A thesis submitted to attain the degree of

DOCTOR OF SCIENCES of ETH ZURICH (Dr. sc. ETH Zürich)

Presented by Manuel Layer MSc. ETH Zurich

born 06.02.1990 citizen of

Federal Republic of Germany

accecpted by the recommendation of Prof. Dr. Eberhard Morgenroth

Dr. Nicolas Derlon Prof. Dr. Christof Holliger

Kim Sörensen Edward van Dijk

2021

Abstract

Today, aerobic granular sludge (AGS) technology is an established alternative to con- ventional activated sludge for the biological treatment of municipal wastewater (WW).

But despite countless studies and full-scale applications of AGS, the effect of partic- ulate organic substrate (X

) - the major constituent of organic substrate in munici- pal WW - on AGS start-up, performance, stability and microbial community is not well understood. Therefore this PhD thesis evaluated physical retention and microbial turnover pathways of X

during AGS sequencing batch reactor (SBR) operation, as well as the influence of X

on formation, process stability, settling performance, nutri- ent removal and microbial community composition of AGS systems.

Physical retention of X

during AGS-SBR operation was identified as a 2-step pro-

cess. Firstly, X

sedimented and accumulated in the bottom of the settled sludge

bed and was retained through surface filtration by the emerging filter-cake. Thus, at-

tachment to biomass was quite limited. Secondly, X

then was preferentially attach-

ing to the flocs during fully-mixed conditions. Results from mathematical modelling

of X

hydrolysis, conversion and turnover resembled and expanded those prior find-

ings. Flocs played a major role in aerobic X

utilisation via aerobic oxidation by or-

dinary heterotrophic organisms (OHO). Synergies between flocs and granules were

observed, where flocs diverted aerobic X

oxidation from the granules and thus sup-

ported anaerobic-feast aerobic-famine conditions for the granules, despite the minor

mass fraction of flocs. Long-term operation of AGS systems fed with different WW

additionally reinforced the results from physical X

retention experiments and mathe-

matical modelling. Complex WW composed of low volatile fatty acids (VFA) and high

X

fractions led to the formation of small granules and 10-40 % ( % of total suspended

solids) of flocs as well as to increased start-up time, decreased nutrient removal and

settling performance. The microbial community of AGS treating WW composed of

X

was characterised by a high abundance of fermenting bacteria, like fermentative

glycogen and phosphorus accumulating organisms (fGAO, fPAO, respectively). Simul-

taneous nitrification-denitrification (SND) was quite limited when AGS was fed with

municipal WW containing X

. The main factors influencing SND in those AGS sys-

tems were identified to be the dynamic of anoxic formation and decay inside the gran-

ule and the availability of organic substrate in the anoxic granule layers. In addition,

the PhD thesis could significantly contribute towards practical understanding of AGS

for the treatment of municipal WW. Most importantly, AGS was distinguished as hy-

brid biofilm system, whereby biofilm (granules) and suspended growth (flocs) coexisted

in synergy. Optimised aeration strategies such as intermittent aeration were identi-

fied to increase SND and total nitrogen removal by AGS systems fed with municipal

WW significantly. Overall, AGS proved to be a simple, efficient and stable process for

the treatment of municipal WW. However, expectations towards settleability, start-up

duration and nutrient removal performance must be lowered if AGS is applied for the

treatment of low-strength municipal WW containing X

.

Zusammenfassung

Aerob granulierter Belebtschlamm (AGS) ist eine relativ neue Alternative zum kon- ventionellen Belebtschlammverfahren in der biologischen Abwasserreinigung. Trotz zahlreicher Studien und Anwendungen im Vollmassstab von AGS wurde der Effekt von partikulärem Substrat (X

), welcher den Hauptanteil an organischem Substrat in kommunalem Abwasser darstellt, noch nicht ausreichend untersucht. Die Zielset- zung der vorliegenden Doktorarbeit war es deshalb, den physischen Rückhalt und die mikrobielle Verwertung von X

im AGS Sequencing-Batch-Reaktor (SBR) Betrieb zu erforschen. Zusätzlich wurde der Einfluss von X

auf die Bildung, Prozessstabilität, Absetzbarkeit, Nährstoffentfernung und mikrobielle Gemeinschaft von AGS analysiert.

Der physische Rückhalt von X

im AGS SBR wurde als zweistufiger Prozess identi-

fiziert. Während der Beschickung im «Plug-Flow» wurde X

hauptsächlich am Bo-

den des Reaktors durch Oberflächenfiltration im Porenraum des abgesetzten Schlamm-

betts, welches vor Allem aus Granula besteht, zurückgehalten. Der Kontakt zwischen

X

und AGS und damit der Stoffumsatz waren dadurch stark limitiert. Sobald der

Reaktor volldurchmischt wurde, adsorbierte X

präferenziell an den Flocken. Diese

Resultate wurden durch mathematische Modellierung des AGS SBR Prozesses bestätigt

und erweitert. Die Flocken spielten eine wichtige Rolle im Rückhalt, der Hydrolyse

und der mikrobiellen Verwertung von X

. Flocken leiteten die mikrobielle Verwertung

von X

weg von den Granula, obwohl Flocken nur einen kleinen Teil an der Gesamt-

biomasse ausmachten. Langzeitversuche zeigten weiter auf, dass niedrige Anteile an

volatilen Fettsäuren zusammen mit X

im Zulauf zu einem erhöhten Anteil an Flocken

führten (10-40% an suspendierten Feststoffen), die Dauer zur Bildung von Granula

deutlich erhöhten, sowie zu einer Verschlechterung der Nährstoffentfernung und Abset-

zverhalten führten. Fermentierende Bakterien wie fermentierende Glykogen- und Phos-

phor akkumulierende Bakterien (fGAO, fPAO) dominierten die mikrobielle Gemein-

schaft in Gegenwart von X

im Zulauf. Die simultane Nitrifikation-Denitrifikation

(SND) war stark beeinträchtigt, sobald AGS zur Reinigung von kommunalem Ab-

wasser eingesetzt wurde. Die limitierenden Faktoren der SND waren die Dynamik der

Bildung und des Verfalls der anoxischen Zonen im Inneren der Granula, sowie die Ver-

fügbarkeit von organischem Substrat in den anoxischen Zonen. Die vorliegende Dok- torarbeit konnte des Weiteren wichtige Erkenntnisse für den praktischen Betrieb von AGS Systemen zur Reinigung von kommunalem Abwasser bereitstellen. AGS wird als

«Hybrid-Biofilm» Verfahren anerkannt, in welchem Flocken und Granula in Synergie

koexistieren. Des Weiteren konnte aufgezeigt werden, dass die Stickstoffentfernung

von AGS Prozessen mithilfe optimierter Belüftungsstrategien (z.B. intermittierender

Belüftung) signifikant erhöht werden kann. Abschliessend erwies sich AGS als stabiles,

effizientes und simples Verfahren zur biologischen Reinigung von kommunalem Ab-

wasser. Die Erwartungen an die Sedimentationsleistung, Dauer zur Bildung von Gran-

ula und Nährstoffentfernung müssen jedoch gedämpft werden, falls AGS zur Reinigung

von stark verdünntem, X

-reichem kommunalem Abwasser eingesetzt wird.

Acknowledgement

First and foremost I want to thank my supervisor Nicolas Derlon and my advisor Eber- hard Morgenroth. Your patience, feedback, sharp minds and positive attitude towards work and life had significant influence on the success of my PhD and you had a very positive impact on my personal development during this time, too. Eberhard, your lec- tures and enthusiasm sparked my interest in biological wastewater treatment and pro- cess engineering during my Master studies, which left me with no choice but to con- duct a PhD with you. Nico, your enthusiasm and excitement about granules, biofilms, reactor operation and process engineering in general is contagious and the best moti- vation and foundation for research I can imagine.

Furthermore, I would like to thank all my students, who conducted their Master’s Thesis under my main- or co-supervision: Eva Reynaert, Antonio Hernandez, Antoine Brison, Franziska Golz, Kristina Bock and Matthias Stähle. Also, two students con- ducting their Master’s Thesis before or during the start of my PhD also significantly contributed to the success of my PhD: Michael Cunningham and Mercedes Garcia, the latter of which later became a colleague and contributed even more towards my PhD.

Your contributions and critical feedback had significant influence on the success of my PhD and my personal development during the PhD.

Chapter 4 of my PhD was set-up, executed, and written in close and fruitful collabora- tion with EPFL. I would like to thank Aline Adler and Christof Holliger and everyone else involved in this project for their contribution, critical feedback and patience. Also, I would like to thank Arnaud Gelb, who helped with microbial community analysis and interpretation in other projects countless times.

In the process engineering department, I would specifically thank Marco Kipf, Richard

Fankhauser, Adriano Joss, Karin Rottermann, Sylvia Richter, Jacqueline Traber, Brian

Sinnet, Canan Aglamaz and Ariane Eberhardt who were directly or indirectly involved

with my experimental, lab, or office work.

The PhD groups of the process engineering (ENG) and urban water management (SWW) departments of Eawag always were a great source of inspiration, very criti- cal feedback and overall great atmosphere and motivation. Thanks to Michele Lau- reni, Ann-Kathrin McCall, Peter Desmond, Lena Mutzner, Christian Thürlimann, Jonas Wielinski, Wenzel Gruber, Bruno Hadengue, Mariane Schneider, Isabell Köp- ping, Natalia Duque, Angelika Hess, Damian Hausherr, Antoine Brison, Matthew Moy de Vitry, Stanley Sam, Barbara Jeanne Ward, Liliane Manny, Valentin Faust, Cassio Schambeck, Omar Wani, Abishek Narayan and Aurea Heusser.

I also want to thank my office mates, who were always a great source inspiration and motivation in hard times, endured my sometimes bad but mostly enthusiastic mood and overall made me a better human being after four joint years: Michel Riechmann, Wenzel Gruber, Carina Doll, Andy Disch, Hans-Peter Zöllig, Antoine Brison, Valentin Faust, Xiaobin Tang, Livia Britschgi and Simon Obrecht.

I am grateful to all my family and friends who supported me during my PhD, which was to be the most challenging time of my life so far. Thank you so much Frieda. You were born right after I started my PhD and now you are almost four years old! You are the most welcome distraction I could ever imagine. Thank you Jule for being a great mother and support during my PhD. Lena, I am so thankful for support, pa- tience and source of inspiration especially during the last few months of my PhD and for life in general.

I cannot thank you all enough for this.

Manu

Contents

Abstract

v

Zusammenfassung

vii

Acknowledgement

x

Contents

xii

1 Introduction

2

1.1 Municipal wastewater treatment . . . . 2

1.2 From activated sludge to advanced biofilm reactors . . . . 3

1.3 The composition of municipal wastewater . . . . 4

1.4 Aerobic granular sludge (AGS) systems . . . . 5

1.4.1 Discovery and definition of AGS . . . . 5

1.4.2 Mechanisms involved in the formation of AGS . . . . 6

1.4.3 AGS for the treatment of municipal wastewater . . . . 7

1.5 Objectives . . . 14

1.6 Thesis outline . . . 14

1.7 References . . . 16

2 Particulate substrate retention in plug-flow and fully-mixed conditions dur- ing operation of aerobic granular sludge systems

24 2.1 Abstract . . . 24

2.2 Introduction . . . 25

2.3 Materials and Methods . . . 28

2.3.1 Experimental approach . . . 28

2.3.2 Experimental set-up . . . 28

2.3.3 Magnetic Resonance Imaging (MRI) . . . 33

2.4 Results . . . 34

2.4.1 Retention of X

during the anaerobic plug-flow feeding of AGS

systems . . . 34

2.4.2 How is X

retained in fully-mixed conditions? . . . 37

2.5 Discussion . . . 39

2.5.1 X

accumulates within the sludge bed during plug-flow feeding but does not attach to granules . . . 39

2.5.2 Large fractions of X

are retained by flocs during fully-mixed conditions . . . 42

2.5.3 Practical implications . . . 43

2.6 Conclusions . . . 44

2.7 Acknowledgements . . . 45

2.8 References . . . 46

3 Pathways of microbial particulate substrate utilisation during anaerobic plug-flow feeding and aerobic fully-mixed conditions in aerobic granular sludge operation

52 3.1 Abstract . . . 52

3.2 Introduction . . . 53

3.3 Materials and Methods . . . 54

3.3.1 Mathematical modelling . . . 54

3.3.2 Calculations . . . 61

3.3.3 Sensitivity analysis . . . 62

3.4 Results . . . 63

3.4.1 How much X

is hydrolysed in anaerobic vs. aerobic conditions and what is the contribution of flocs? . . . 63

3.4.2 What is the role of flocs in X

hydrolysis? . . . 63

3.4.3 What are the main microbial conversion pathways of X

? . . . 64

3.4.4 To what extent can operating conditions be optimised to en- hance anaerobic X

utilisation in AGS-SBR operation? . . . 69

3.4.5 Sensitivity analysis . . . 70

3.5 Discussion . . . 71

3.5.1 X

is a disadvantageous organic substrate in AGS systems . . . . 71

3.5.2 Flocs play a major role in X

conversion and utilisation in AGS systems . . . 72

3.5.3 How can anaerobic X

utilisation be enhanced in AGS-SBR op-

eration? . . . 73

3.5.4 Modelling limitations and outlook . . . 74

3.6 Conclusions . . . 75

3.7 Acknowledgements . . . 76

3.8 References . . . 77

4 Organic substrate diffusibility governs microbial community composition, nutrient removal performance and kinetics of granulation of aerobic granu- lar sludge

84 4.1 Abstract . . . 84

4.2 Introduction . . . 85

4.3 Materials and methods . . . 88

4.3.1 Experimental approach . . . 88

4.3.2 Experimental set-up . . . 88

4.3.3 Start-up approach . . . 89

4.3.4 Start-up definition . . . 89

4.3.5 Wastewater composition and sludge inoculum . . . 90

4.3.6 Physical sludge parameters . . . 92

4.3.7 Analytical methods . . . 92

4.3.8 Microbial community analysis . . . 92

4.4 Results . . . 94

4.4.1 Settling properties . . . 94

4.4.2 Sludge size fractions . . . 95

4.4.3 Evolution of the bacterial community composition from inocula- tion to stable state . . . 97

4.5 Bacterial communities of granules and flocs . . . 101

4.5.1 Nutrient removal performance . . . 103

4.5.2 Correlations between settling properties, nutrient-removal perfor- mances and microbial community composition . . . 103

4.5.3 Identification of correlations between the discriminant taxa and the sludge size distributions, the settling properties, and the nutrient-removal performances. . . 105

4.6 Discussion . . . 108

4.6.1 Diffusibility of organic substrates has significant influence on for- mation of AGS . . . 108

4.6.2 Start-up . . . 109

4.6.3 Steady-state . . . 111

4.6.4 The role of flocs in AGS systems . . . 113

4.6.5 Implications for research and practice . . . 114

4.7 Conclusions . . . 115

4.8 Acknowledgments . . . 116

4.9 References . . . 117

5 Limited simultaneous nitrification-denitrification (SND) in aerobic granular sludge systems treating municipal wastewater: Mechanisms and practical implications

126 5.1 Abstract . . . 126

5.2 Introduction . . . 127

5.3 Materials and Methods . . . 130

5.3.1 Experimental approach and reactor configuration . . . 130

5.3.2 Modelling . . . 132

5.3.3 Calculations . . . 136

5.3.4 Analytical methods . . . 137

5.4 Results . . . 137

5.4.1 How is SND influenced by WW composition during long-term operation? (experimental results) . . . 137

5.4.2 How does DO concentration affect SND performance? (experi- mental and modelling results) . . . 137

5.4.3 Dynamics of redox zone formation (modelling results) . . . 139

5.4.4 Which electron donors are actually used for NO

-N removal? (modelling results) . . . 141

5.4.5 Is there potential for electron-donor shift towards anoxic utilisa- tion pathways? (modelling results) . . . 141

5.4.6 Can SND and TN removal be improved by optimising the aera- tion strategy? (modelling results) . . . 145

5.5 Discussion . . . 145

5.5.1 SND and TN removal is limited in AGS systems treating munici- pal WW . . . 145

5.5.2 What mechanisms limit SND in AGS systems treating municipal WW? . . . 149

5.5.3 How to optimise for TN removal in AGS systems? . . . 151

5.6 Conclusions . . . 153

5.7 Acknowledgements . . . 154

5.8 References . . . 155

6 Conclusions

161

163 7.1 Is AGS suitable for the treatment of municipal wastewater? . . . 163

7.2 AGS vs. established systems for the treatment of municipal wastewater . 165 7.3 AGS in continuous flow systems . . . 166

7.4 Modelling of AGS systems . . . 167

7.5 What is the role of flocs in AGS systems? . . . 167

7.6 References . . . 169

Appendix 175

175 A.1 Sludge images . . . 175

A.2 Calculation of velocity gradient (G) for jar tests . . . 176

A.3 Image processing and analysis of 3D MRI images . . . 177

A.4 Flocs / granules number and surface area calculations . . . 177

B Supplementary Information: Chapter 3

179 B.1 Distribution of granules and flocs during AGS SBR operation . . . 179

B.2 Kinetic model . . . 179

B.3 S

storage kinetics of PAO and GAO . . . 180

B.4 Microbial community composition of the default scenario . . . 181

C Supplementary Information: Chapter 4

182

196 D.1 SND efficiency calculation from literature . . . 196

D.2 AGS model description . . . 205

D.2.1 Sumo Biofilm Model . . . 205

D.2.2 Biokinetic Model . . . 206

D.2.3 Reactor Model . . . 206

E References

210

Curriculum Vitae 213

Publications 215

1. Introduction

1.1. Municipal wastewater treatment

The treatment of municipal wastewater (WW) is often referred to as the most impor- tant development in human health of the last two centuries (Henze et al., 2008). Its introduction became necessary, whenever the “capacity of a river for self-purification”

(Jenkins and Wanner, 2014) was exceeded, which was especially the case in densely populated areas, and led to hygienic and ultimately human health problems. The ini- tial goals of municipal WW treatment were therefore to purify water in order to limit risks towards humans and ensure hygiene (Henze et al., 2008). During the 1970s the goals of WW treatment were extended in order to protect the downstream users of water and the aquatic environment (Jenkins and Wanner, 2014), and ultimately to today’s immission control, which incorporates potentially harmful impacts on human beings and their environment (Umweltbundesamt, 2020). Today, wastewater treatment plants (WWTP) have to fulfil ever increasing treatment requirements, such as reaching very low effluent concentrations of nitrogen (N) and phosphorus (P) compounds, with a simultaneous minimization of space requirements, energy usage and greenhouse gas (GHG) emissions or recovery of valuable resources ( e.g. , N, P) from WW streams. The transition from WWTP, which remove pollutants from human waste streams, towards water resource recovery facilities (WRRF), which produce valuable goods such as clean water and nutrients, is therefore required.

Today, WWTP typically consists of 3 treatment steps: primary (mechanical / physi-

cal), secondary (biological) and tertiary (filtration, disinfection, advanced oxidation)

treatment. Primary treatment consists of mechanical pre-treatment and is typically

constituted of fat and grit-removal, as well as primary sedimentation, which removes

a large part of settleable solids and organics, sand, fat and other coarse fractions from

WW. The subsequent secondary treatment is a biological treatment step. The over-

whelming majority of biological WW treatment today is accomplished by utilising

the so-called “activated sludge” process, whereby microorganisms are cultivated to re-

move or convert organics and nutrients from WW, first published by Ardern and Lock-

ett, 1914. The growth of those microorganisms occurs in the suspended phase, i.e. , in

the form of activated sludge flocs. Activated sludge flocs are then separated from the

treated WW in the secondary clarifier. In AS WWTP, secondary clarifiers often limit

the treatment capacity of the entire WWTP, since settleability and thus separation

rates of activated sludge and treated WW are low. Different reactor configurations ex-

ist, such as continuous flow systems or sequencing batch reactors (SBR), whereby the

processes of reaction and separation are either separated in time (SBR), or in space

(continuous flow systems). Activated sludge processes can be operated independently from reactor configuration. For a long time, the activated sludge process was solely designed to remove organic substrate (carbon) and pathogens by aerobic oxidation.

Advances in process understanding, design and operation allowed to integrate biologi- cal nitrification (NH

to NO

) and biological nitrogen removal (nitrification + den- itrification, NH

to NO

to N

) within the activated sludge process. Eventually, en- hanced biological phosphorus removal (EBPR) activated sludge integrated within the activated sludge process. Today, many existing WWTP are reaching their maximum capacity. Also, due to urbanisation, it is required to intensify and improve WW treat- ment. Therefore, research is required to identify and evaluate new technologies with improved treatment capacity, lower space requirements and high economic efficiency.

1.2. From activated sludge to advanced biofilm reactors

Biofilm reactors for the treatment of municipal WW have been applied in the form of trickling filters since the early 1900s (Morgenroth, 2008). However, since the 1980 - 1990s more advanced biofilm reactor systems, like submerged biofilm, integrated fixed- film activated sludge (IFAS), membrane bio-reactors (MBR), or moving bed biofilm reactors (MBBR) are widely applied in municipal WW treatment. Biofilms for WW treatment rely on a substratum - a surface material like a membrane, cloth or car- rier material - for attachment and growth, and are oftentimes referred to as “attached growth” in opposition to “suspended growth” as in activated sludge systems. Biofilm reactors promise high volumetric loadings, good effluent quality without requiring solids separation (secondary clarifier) or recirculation (both not in hybrid biofilm - suspended growth systems, like IFAS) resulting in a lower footprint as compared to ac- tivated sludge systems (Rittmann, 1982). The newest addition to aerobic biofilm sys- tems are aerobic granular sludge (AGS) systems, in which the biomass self-immobilises into so-called granules, without relying on a substratum (Beun et al., 1999; Morgen- roth et al., 1997).

In general, aerobic biofilm reactors can achieve similar treatment performance as acti-

vated sludge systems, like organic substrate removal, nitrification, denitrification and

EBPR (Morgenroth, 2008). The main difference between activated sludge and biofilm

WW treatment is that mass-transport of substrate (electron-donor and electron-acceptor)

is limited by diffusion in biofilms. Limited mass-transport has vast impacts on micro-

bial ecology, reactor operation and design (Morgenroth, 2008). Mass-transport limita-

tions - in theory - allow for simultaneous nutrient removal processes like simultaneous

nitrification - denitrification (SND) (Falkentoft et al., 1999; Hibiya et al., 2004; Nielsen et al., 1990).

Besides aerobic biofilms, also anaerobic biofilm systems for municipal or industrial WW treatment exist. Such systems include anaerobic filters (AF), upflow anaerobic sludge blankets (UASB) or integration of anammox technology in a biofilm environ- ment (either through attached-growth or as anammox granules) (Arrojo et al., 2006;

McHugh et al., 2003; Tsushima et al., 2007). Biomass in UASB or anammox granular reactors consist of granular biofilms, which self-immobilise (without requiring a sub- stratum) (Lettinga, 2001), similar to AGS technology.

However, despite many full-scale applications of advanced biofilm systems for the treat- ment of municipal WW, the majority of WWTP worldwide operates as activated sludge system. Further research and improvement of advanced biofilm systems is therefore re- quired in order to intensify and improve treatment of municipal WW.

1.3. The composition of municipal wastewater

Municipal WW is a complex mixture of organic and inorganic pollutants and differen- tiates from other WW streams ( e.g. , industrial WW, stormwater) by its heterogeneity in terms of composition and typically low concentration ranges (resulting from high di- lution). Municipal WW is composed of microorganisms (pathogens, bacteria, viruses, etc. ), biodegradable and other organic materials (organic substrate, carbonaceous com- pounds), macro- and micronutrients (nitrogen (N), phosphorus (P), etc. ), as well as metals and other inorganic materials (Cu, Zn, acids, bases, etc. ) (Henze et al., 2008).

Organic substrate, the main pollutant in municipal WW, is typically characterised by the sum parameters chemical oxygen demand (COD) or biological oxygen demand (BOD). COD can be differentiated in terms of physical properties (material size) be- tween total, soluble (sCOD, after filtration at 0.45 µ m), and particulate (pCOD, total minus soluble COD) fractions. COD can also be differentiated in terms of biological availability between readily biodegradable COD (rbCOD), slowly biodegradable COD (sbCOD) and inert (unbiodegradable) COD. Another important measure of organic substrate is BOD. BOD indicates a highly biodegradable fraction of organic substrate and is typically measured after 5 days of aerobic degradation (BOD

). Volatile fatty acids (VFA) are a fraction of organic substrate with very high biodegradability, specifi- cally, acetate (Ac) and propionate (Pr). VFA are part of BOD

and rbCOD. Availabil- ity of rbCOD is essential for EBPR (Gerber et al., 1986).

Particulate and sbCOD are both proxies for polymeric organic substrate, or sheer par-

ticles, both referred to as X

, contrary to readily biodegradable organic substrate (S

). X

typically represents > 50 % of organic substrate in the influent WW (mea- sured as COD) (Metcalf and Eddy, 2014), but the specific composition in terms of protein, carbohydrate and lipid can vary significantly between different municipal WW (Sophonsiri and Morgenroth, 2004). In opposition to monomeric organic substrate (S

), X

cannot be transported through the cell envelope of microorganisms due to its size larger than 10

amu. Therefore, organic substrate polymers have to be trans- formed into monomers by hydrolysis – enzymatic depolymerisation – before microor- ganisms can utilise this type of organic substrate (Morgenroth et al., 2002). Protozoa contribute to the degradation of particulate organic substrate through direct uptake and intracellular degradation. However, it remains unclear, whether protozoa signifi- cantly contribute to X

degradation in biological WW treatment (Morgenroth et al., 2002). In general, it is assumed that hydrolysis rates are reduced in anoxic or anaer- obic redox conditions, in comparison to aerobic redox conditions (Henze and Mladen- ovski, 1991). The reduction could be related to the activity of protozoa, whose typi- cally aerobic-only metabolism would significantly be reduced in anoxic or anaerobic re- dox conditions (Morgenroth et al., 2002). In general, it is assumed that the microbial community and protozoa cultivated within the biological stage of the WWTP carry out hydrolysis. Therefore, hydrolysis by microorganisms is initiated only once physical contact between X

and the microorganisms has been established. However, research has hypothesised that hydrolytic microorganisms could already be attached to X

be- fore it enters the WWTP, and that only a small fraction of the microorganisms of the WWTP actually contribute to hydrolysis (Benneouala et al., 2017). Due to hydrol- ysis, the rate of biological oxidation of X

is typically one order of magnitude lower compared to S

(Levine et al., 1985), and hydrolysis is often referred to as the rate- limiting step in biological WW treatment (Morgenroth et al., 2002). Understanding and behaviour of X

retention, hydrolysis and utilisation is still limited in biological WW treatment, and research is therefore required.

1.4. Aerobic granular sludge (AGS) systems

1.4.1. Discovery and definition of AGS

AGS is a technology for biological WW treatment and composed of self-immobilised

biofilm aggregates without relying on a carrier material (substratum). Granules are

defined as biofilm aggregates of “microbial origin, which do not coagulate under re-

duced hydrodynamic shear, and which settle significantly faster than conventional [ac- tivated] sludge” (de Kreuk et al., 2007). AGS was first discovered and cultivated in two laboratories in Munich and Delft, respectively (Beun et al., 1999; Morgenroth et al., 1997). The minimum size of granules was set to 0.2 mm - 0.25 mm, and high set- tleability of AGS can be expressed as sludge volume indices after 10 and 30 minutes (SVI

, SVI

, respectively) being similar: SVI

≈ SVI

. High settleability of AGS originates from high settling velocities of 25-70 m h

, which is significantly higher than the settling velocity of activated sludge flocs of 7-10 m h

(Adav et al., 2008).

1.4.2. Mechanisms involved in the formation of AGS

AGS was first cultivated in SBR by using an aggressive washout strategy on slow set- tling biomass (activated sludge flocs) and relatively high shear force (Beun et al., 1999;

Morgenroth et al., 1997). Since 2000, AGS was studied intensively, and most stud- ies focused on the cultivation of AGS in synthetic WW, mostly composed of readily biodegradable substrates (VFA) (de Kreuk et al., 2007). The fast formation of AGS in such conditions was closely linked to the application of (1) high wash-out stress on slow settling biomass (short settling), (2) application of high shear stress (aeration in- tensity or mixing) and (3) microbial selection through anaerobic-feast - aerobic-famine SBR operation (Adav et al., 2008). AGS fed by synthetic WW can achieve the follow- ing characteristics within 30 days after start-up (de Kreuk et al., 2005; Lochmatter and Holliger, 2014):

(1) Formation of large granules, complete transformation of flocs to granules: diame- ter 0.5-2 mm, granule fraction > 95 %

(2) Excellent settleability: SVI

< 40 mL g

, SVI

/SVI

ratio ≈ 1 (3) Simultaneous removal of carbon, nitrogen and phosphorus

Granules form through a gradual transformation process of floccular biomass to com- pact granules, and AGS formation is proposed to include four steps: (1) cell-to-cell attraction by physical, chemical or biochemical forces, (2) microbe-to-microbe attach- ment by hydrodynamic, diffusional and / or thermodynamic forces, (3) microbial ad- hesion enhancement through excretion of extracellular polymeric substances (EPS) and (4) hydrodynamic shear force to stabilise the granule structure (Winkler et al., 2018).

AGS have mainly been cultivated using SBR. Operation of SBR allow for different

operational phases, which were identified as being important for AGS formation and

long-term stability. Feeding of AGS-SBR systems is typically done in plug-flow con- ditions, whereby fresh WW is injected from the bottom into the settled sludge bed, while effluent is withdrawn simultaneously from the top of the reactor (referred to as

“simultaneous fill-draw” mode or “constant volume” operation) (Derlon et al., 2016).

Plug-flow conditions allow for high substrate gradients, which act as driving force for diffusion of substrate into the granules. Anaerobic redox conditions prevail dur- ing feeding and benefit the proliferation and growth of carbon-storing organisms like phosphorus and glycogen accumulating organisms (PAO, GAO, respectively). PAO and GAO demand a key role in AGS formation and stability (de Kreuk and van Loos- drecht, 2004; Winkler et al., 2018). Ideally, all organic substrate is removed through internal carbon-storage during the anaerobic phase by GAO and PAO, respectively. If organic substrate is not fully used up during anaerobic conditions, organic substrate is available in aerobic conditions resulting in aerobic growth of ordinary heterotrophic organisms (OHO), which harm AGS system stability (Sturm et al., 2004). The subse- quent aerobic SBR phase is applied to perform nitrification, EBPR and oxidation or replenishment of internally stored carbon. Afterwards, a short settling period is ap- plied. Excess sludge is removed selectively, i.e. , ideally only slow settling biomass is removed, either after settling (from the sludge fraction still in suspension) or feeding (from the top of the settled sludge bed) (Lochmatter and Holliger, 2014).

1.4.3. AGS for the treatment of municipal wastewater

AGS was first cultivated for the treatment of municipal WW by de Kreuk and van Loosdrecht, 2006. But despite very fast formation and excellent settling properties of AGS in municipal WW conditions, nutrient removal was not satisfactory (Table 1.1).

This early research indicated that the transition of AGS cultivated in the laboratory, typically fed by simple synthetic VFA-rich WW, to cultivation of AGS in municipal WW is quite challenging. Also, operating conditions to achieve granulation as well as mechanisms involved in the formation of AGS might be a function of the influent WW composition.

Nevertheless, Dutch company RoyalHaskoningDHV brought the Nereda

process to

the market, the first full-scale AGS system for treatment of municipal WW (Giesen et

al., 2013; van der Roest et al., 2011). RoyalHaskoningDHV claims significantly lower

energy and chemical consumption, savings in space, lower investment and operational

costs, as well as simultaneous removal of carbon, nitrogen and phosphorus and overall

excellent treatment performance. Today, over 70 full-scale Nereda

plants exist, one

being in operation (ARA Alpnach) and one being under construction (ARA Kloten- Opfikon) in Switzerland (Ali et al., 2019; Pronk et al., 2017).

However, further research indicated that expectations towards AGS cultivated in real municipal WW must be lowered in comparison to early lab studies which conducted research on AGS using synthetic VFA-rich WW (Figure 1.1, Table 1.1), specifically:

(1) Much increased start-up time: 20 d up to > 1 year

(2) Decreased settleability: SVI

40-100 mL g

, SVI

/SVI

ratio 0.6-1.0 (3) Increased presence of flocs: granule fraction 60-98 %

(4) Limited (simultaneous) nutrient removal performances: incomplete nitrification, incomplete denitrification, increased effluent TN

(5) Distinct morphology of AGS

A literature review on AGS cultivated in municipal WW revealed inconsistencies and wide ranges in applied operational strategies, sludge characteristics and nutrient re- moval performances, indicating that understanding of feasibility, operation and perfor- mance of AGS for the treatment of municipal WW is still limited (Table 1.1). Specifi- cally, many studies cultivated AGS using municipal WW composed of either very high total organic substrate concentrations of 500-1000 mgCOD L

, very high fractions of S

(and hence low X

), or municipal WW with the addition of VFA. In addition, some studies relied on inoculation with (crushed) AGS or excess sludge from other AGS plants, rather than starting up from activated sludge flocs, which likely signif- icantly decreased start-up time (Linlin et al., 2005; Pijuan et al., 2011; Pronk et al., 2015).

Therefore, one of the main questions is whether AGS can be cultivated without inoc- ulation in municipal WW composed of low S

and high X

concentrations, typically also referred to as low-strength municipal WW. Further research is therefore required.

Cultivation of AGS in municipal WW is quite challenging (Table 1.1). Operational

strategies to cultivate AGS in lab-scale environments fed by VFA as sole carbon source

were found to be not or only partially applicable in municipal WW conditions. Too

high wash-out stress in municipal WW fed AGS systems results in breakdown of the

nutrient removal performance of AGS systems (Derlon et al., 2016), and high shear-

stress is not required to form AGS in municipal WW (Devlin et al., 2016; Devlin and

Oleszkiewicz, 2018). Researchers therefore focused on and successfully applied the

Figure 1.1.: Morphology after 10 months of operation of 100%-VFA synthetic WW fed AGS (left) and raw WW fed AGS (right) (Layer et al., 2019). Scale bars = 1 mm.

strategy of microbial selection (via anaerobic-feast aerobic-famine SBR operation) in

combination with increased, but not very high washout stress (via short settling time)

to cultivate AGS in municipal WW (Bassin et al., 2019; Campo et al., 2020). Micro-

bial selection requires preferentially supporting the growth of slow-growing organisms

like PAO and GAO. The presence of PAO and GAO is ultimately determined by the

amount of biodegradable organic substrate (and ortho-P for PAO) in the influent WW

(Gerber et al., 1986; Wei et al., 2020). Often, both ortho-P and organic substrate con-

centrations are low, especially in Switzerland (Sollfrank and Gujer, 1991). However,

the proliferation of PAO and GAO can be additionally promoted by operational mea-

sures. Such measures include the promotion of full organic substrate uptake during the

anaerobic phase (feast), in order to (1) support the competitive advantage of carbon-

storing organisms over OHO and (2) limit the organic substrate availability in aerobic

conditions to restrict the growth of OHO. One of the key differences between lab-scale

synthetic WW and municipal WW is the presence of non-diffusible X

in municipal

WW (like polymeric carbohydrates, proteins, etc. ). The presence of X

could poten-

tially harm anaerobic-feast - aerobic-famine conditions and therefore limit microbial

selection in AGS-SBR operation (de Kreuk et al., 2010; Derlon et al., 2016; Jabari et

al., 2016; Wagner et al., 2015b). For the application of AGS for the treatment of mu-

nicipal WW it is therefore key to understand the effect of X

on AGS systems.

Table1.1.:PerformancecharacteristicsofmunicipalWW(MWW)fedAGSsystems,adoptedandextendedfromDerlonetal.,2016 InfluentSubstratecon- centrations [mgCODL-1]

Organic loading [kgCODtot m-3d-1]

SBRoperation (settlingtime [min],vww[m h-1]) Granulation time[d]Settlingprop- erties[SVImL g-1,SVI30/5 ratio]

Granule fractionSizeof granules [mm]

Effluentquality [mgL-1]InoculumReference MWW (China)CODtot:95-200 CODsol:35-1200.6-1Variablevol- ume300SVI30:40 Ratio:180%0.2-0.8

NH4-N0.5 NO3-N:4 PO4-P:0.5 TSS:15

Nietal., 2009 MWW (40%)and industrial WW (60%) (China)

CODsol250-1800Variablevol- ume400SVI30:<50 Ratio:180-90%0.35Liuetal., 2010 MWW (40%)and indus- trialWW (60%) (Singa- pore)

CODsol1000Variablevol- ume(settling 2-10min)

48

SVI30:30-120 (fluctuating,then stabilisingat <50) Ratio:1

75%2.25NH4-N:0 NO3-N:8Liuetal., 2011 MWW (South Africa)

CODtot:1265Constant volumeSVI30:<50

NH4-N:<2.0 TN:14 TP:5 TSS:<10

Giesenetal., 2013;van derRoest etal.,2011 MWW+ industrial WW(The Nether- lands)

Constant volume400SVI30:40 Ratio:0.6690%

NH4-N0.5 NO3-N:2-4 PO4-P:0.5 TSS:10-20

Giesenetal., 2013;van derRoest etal.,2011 MWW+ acetate (Australia)

CODtot:326 CODsol:179Variablevol- ume SVI30:<100 Ratio:0.9PO4-Paccumu- lationduring aerobicphase

Comaetal., 2012

InfluentSubstratecon- centrations [mgCODL-1]

Organic loading [kgCODtot m-3d-1]

SBRoperation (settlingtime [min],vww[m h-1]) Granulation time[d]Settlingprop- erties[SVImL g-1,SVI30/5 ratio]

Granule fractionSizeof granules [mm]

Effluentquality [mgL-1]InoculumReference MWW (Brazil)CODsol:4301.0-1.4Variablevol- ume(15min settling)

140SVI30:53 Ratio:0.90.3-1.3NH4-N:3 TSS:25-125ASWagnerand DaCosta, 2013 MWW+ acetate (40%of COD) (Germany)

CODtot:287-4920.5-2.0Variablevol- ume(15-30 minsettling) 12598%1.1-1.8NH4-N:<5 PO4-P:<3 TSS:40-100

EBPR ASRocktäschel etal.,2015 MWW (The Nether- lands)

CODtot:330 CODsol:280 VFA:80 1.0-1.5Variablevol- ume,Plugflow feeding(6-15 minsettling) 20SVI10:381.1NH4-N:51 PO4-P6 VSS100

Activated sludge (AS) + AGS efflu- ent solids

deKreuk andvan Loosdrecht, 2006 MWW (Switzer- land)

CODtot:304 CODsol:1271.0Constant volume,Plug flowfeeding (3-10min settling,vww 1-16) SVI30:80(at lowandhigh vww) Ratio:0.7-0.95

60-80%>0.63 (forhigh vww) 0.25 -0.63 (forlow vww)

NH4-N:<0.2 NO3-N3.9 PO4-P<0.2 atlowvww, athighervww performance worse

ASDerlonetal., 2016 MWW (Brazil)

CODtot:150-450CODsol:88-428 BOD5:90-3801.1Variablevol- ume(13-35 settling) 56SVI30:<70 Ratio:160%0.29

NH4-N:14 NO2-Naccumu- lation(16.5) PO4-P<4

ASGuimarães etal.,20177 MWW (Brazil)

CODtot:588 CODsol:304.42.1Variablevol- ume(settling 10-35) 140SVI30:76 Ratio:0.9582%0.5NH4-N:0.2-7.5 NO2-Naccumu- lation(17.3)ASWagneret al.,2015a

InfluentSubstratecon- centrations [mgCODL-1]

Organic loading [kgCODtot m-3d-1]

SBRoperation (settlingtime [min],vww[m h-1]) Granulation time[d]Settlingprop- erties[SVImL g-1,SVI30/5 ratio]

Granule fractionSizeof granules [mm]

Effluentquality [mgL-1]InoculumReference MWW (China)CODtot:200-320 CODsol:80-1500.4-1.4Variablevol- ume(8-20min settling)

45SVI30:20-35NH4-N:0 TN:9 TP:1.4

Anaerobic di- gested sludge

Suetal., 2012 MWW (China)CODtot:1790.6-1.3Constant volume(5min settlingvww 0.35-0.62)

85SVI30:<500.2-0.8NH4-N:0.3-15 TN28-37 TP4.8-5.9

ASWangetal., 2018 MWW (Italy)CODtot:133-178 CODsol:111-1280.6-0.8Variablevol- ume(13-25 minsettling)

50SVI30:24 Ratio:191%1.5NH4-N:4.3-4.8 TN:10 TP:0.09

AS+ AGSCampoet al.,2020 MWW+ acetate (Italy)

CODsol:2901.1Variablevol- ume(15-25 minsettling)

45SVI30:28 Ratio:199%0.2-1.5ASSguanciet al.,2019 MWW (Poland)CODtot:1320 BOD5:11201.3Constantvolume (20minsettling,vww1.1)SVI30:48 Ratio:0.7580%0.09- 0.35

NH4-N:0.4 TN:12 NO3-N:1.6 NO2-N:0.5 TP:0.9

ASŚwiątczak andCydzik- Kwiatkowska, 2018 MWW (Turkey)

CODtot:900-1100 CODsol:760-7902.1-2.4Variablevol- ume(3-15min settling)

25-44SVI30:25 Ratio:0.7-0.957-88%0.6-1.2NH4-N:21-24 TP:1.9-2.5ASCetinetal., 2018 MWW (The Nether- lands)

CODtot:720 BOD5:263Constant volume SVI30:<40 Ratio:0.95

NH4-N:1.4-3.0 TN7.2 TP0.9

Pronketal., 2017 MWW (Austria)CODtot:3881.0Constant volume(1-3 minsettling, vww2.4)

28SVI30:<40 Ratio:0.9-1.0>90%0.55 (d50) TN:5-11 TP:0.6-1.6ASJahnetal., 2019

InfluentSubstratecon- centrations [mgCODL-1]

Organic loading [kgCODtot m-3d-1]

SBRoperation (settlingtime [min],vww[m h-1]) Granulation time[d]Settlingprop- erties[SVImL g-1,SVI30/5 ratio]

Granule fractionSizeof granules [mm]

Effluentquality [mgL-1]InoculumReference MWW (Brazil)CODtot:462 CODsol:172 BOD5:148

1.3Variablevol- ume(15-55 minsettling) 35SVI30:<55 Ratio:0.9-1.0>85%0.9-1.3 (d50)NH4-N:4-8 NO2-N:4-14 NO3-N:2-4 PO4-P1-2

ASRollemberg etal.,2020

1.5. Objectives

The main objective of the thesis was thus to understand the effect of X

on the mech- anisms of formation, process performance, stability and microbial community composi- tion of AGS systems (Figure 1.2). The research questions were:

(1) How is X

physically retained during AGS-SBR operation (short-term)?

(2) What are the microbial X

utilisation pathways in AGS-SBR operation (short- term)?

(3) What is the effect of influent WW composition in terms of diffusible and non- diffusible organic substrate on start-up, settling performance, nutrient removal and microbial community of AGS systems (long-term)?

(4) Why is simultaneous nitrification-denitrification (SND) limited in AGS systems treating municipal WW (short and long-term)?

1.6. Thesis outline

The thesis was structured as a cumulative dissertation, and split into 7 chapters. Chap-

ter 1 contains the introduction, research questions and outline of the PhD thesis. Chap-

ter 2 focused on the short-term, i.e. , SBR cycle duration, physical interaction between

X

and AGS, i.e. , before microbial conversion of X

is initiated, during anaerobic

plug-flow and aerobic fully mixed conditions. Chapter 3 focused on the microbial con-

version of X

in AGS systems during anaerobic plug-flow feeding and aerobic fully

mixed conditions (short-term) using a mathematical model implemented in SUMO

.

Chapter 4 focused on the effect of influent WW composition in terms of diffusible and

non-diffusible (X

) organic substrate on the formation, performance, nutrient removal

and microbial community development of AGS. Therefore, four AGS systems were

inoculated with conventional activated sludge, started up and operated for 400 d in

parallel. Similar operating conditions were applied, but different influent WW charac-

terised by different levels of complexity (from X

-free 100 % -VFA synthetic to X

-rich

municipal raw WW) were fed to the reactors. Chapter 5 aimed at understanding what

mechanisms govern limited SND in AGS systems treating municipal WW, based on

long-term observations originating from Chapter 4 as well as a mathematical model

implemented in SUMO

. Chapters 6 and 7 contain an overall conclusion, as well as

an outlook of the PhD thesis.

Figure 1.2.: Overview of the main objectives of the PhD Thesis.

1.7. References

Adav, S., Lee, D.-J., Show, K.-Y., & Tay, J.-H. (2008). Aerobic granular sludge: Re- cent advances. Biotechnology advances , 26 (5), 411–423.

Ali, M., Wang, Z., Salam, K. W., Hari, A. R., Pronk, M., van Loosdrecht, M. C. M.,

& Saikaly, P. E. (2019). Importance of species sorting and immigration on the bacterial assembly of different-sized aggregates in a full-scale aerobic granular sludge plant. Environmental Science and Technology , 53 (14), 8291–8301.

Ardern, E., & Lockett, W. T. (1914). Experiments on the oxidation of sewage without the aid of filters. Journal of the society of chemical industry , 33 (10), 523–539.

Arrojo, B., Mosquera-Corral, A., Campos, J. L., & Méndez, R. (2006). Effects of me- chanical stress on anammox granules in a sequencing batch reactor (sbr). Jour- nal of Biotechnology , 123 (4), 453–463.

Bassin, J. P., Tavares, D. C., Borges, R. C., & Dezotti, M. (2019). Development of aer- obic granular sludge under tropical climate conditions: The key role of inocu- lum adaptation under reduced sludge washout for stable granulation. Journal of Environmental Management , 230 , 168–182.

Benneouala, M., Bareha, Y., Mengelle, E., Bounouba, M., Sperandio, M., Bessiere, Y.,

& Paul, E. (2017). Hydrolysis of particulate settleable solids (pss) in activated sludge is determined by the bacteria initially adsorbed in the sewage. Water Research , 125 (15), 400–409.

Beun, J. J., Hendriks, A., van Loosdrecht, M. C. M., Morgenroth, E., Wilderer, P. A.,

& Heijnen, J. J. (1999). Aerobic granulation in a sequencing batch reactor.

Water Research , 33 (10), 2283–2290.

Campo, R., Sguanci, S., Caffaz, S., Mazzoli, L., Ramazzotti, M., Lubello, C., & Lotti, T. (2020). Efficient carbon, nitrogen and phosphorus removal from low c/n real domestic wastewater with aerobic granular sludge. Bioresource Technology , 122961.

Cetin, E., Karakas, E., Dulekgurgen, E., Ovez, S., Kolukirik, M., & Yilmaz, G. (2018).

Effects of high-concentration influent suspended solids on aerobic granulation in pilot-scale sequencing batch reactors treating real domestic wastewater. Wa- ter Research , 131 , 74–89.

Coma, M., Verawaty, M., Pijuan, M., Yuan, Z., & Bond, P. L. (2012). Enhancing aer- obic granulation for biological nutrient removal from domestic wastewater.

Bioresource Technology , 103 (1), 101–108.