Microbial resource management of OLAND focused on sustainability

Promotors:

Prof. dr. ir. Willy Verstraete

Department of Biochemical and Microbial Technology, Faculty of Bioscience

Engineering, Ghent University, Gent, Belgium

Prof. dr. ir. Nico Boon

Department of Biochemical and Microbial Technology, Faculty of Bioscience


Members of the examination committee:

Prof. dr. ir. Pascal Boeckx

Department of Applied analytical and physical chemistry, Faculty of Bioscience


Prof. dr. Juan M. Lema

Department of chemical engineering, School of engineering, University of Santiago de

Compostela, Santiago de Compostela, Spain

Dr. Bernhard Wett

ARAconsult GmbH, Innsbruck, Austria

Prof. dr. ir. Ingmar Nopens (Secretary)

Department of Mathematical Modelling, Statistics and Bioinformatics, Faculty of

Bioscience Engineering, Ghent University, Gent, Belgium

Peter Bossier (Chairman)

Departement of animal production, Faculty of Bioscience Engineering, Ghent University,

Gent, Belgium

Dean:

Prof. dr. ir. Guido Van Huylenbroeck

Rector:

Prof. dr. Paul Van Cauwenberge

http://www.usc.es/

http://www.usc.es/

ir. Haydée De Clippeleir

Microbial resource management

of OLAND

focused on sustainability

Thesis submitted in fulfillment of the requirements for the

degree of Doctor (PhD) in Applied Biological Sciences

Dutch translation of title:

Microbial resource management van OLAND met de focus op duurzaamheid

This work was supported by the Institute for the Promotion of Innovation by Science and

Technology in Flanders (IWT-Vlaanderen, number SB-81068).

Cover illustration:

"Energieweelde" made by Lutgarde Van Hoey, based on a design of Kantschool Artofil,

Nadine Pauwels. Photography by Verne.

To refer to this thesis:

De Clippeleir, H. (2012) Microbial resource management of OLAND focused on

sustainability. PhD thesis, Ghent University, Belgium.

ISBN: 978-905989-551-5

The author and the promotors give the authorisation to consult and to copy parts of this work

for personal use only. Every other use is subject to the copyright laws. Permission to

reproduce any material contained in this work should be obtained from the author.

Notation index

i

Notation index

A/B process 2-stage activated sludge system

AD anaerobic digestion

ADP abiotic depletion potential

AerAOB aerobic ammonium-oxidizing bacteria

AnAOB anoxic ammonium-oxidizing bacteria

AS activated sludge

AX/B system A/B process with X% COD removal in the A-stage

bCOD biodegradable fraction of the chemical oxygen demand

BOD biological oxygen demand

CAS conventional activated sludge system

CHP combined heat and power

COD chemical oxygen demand

CSTR continuous stirred tank reactor

DEMON OLAND with pH controlled aeration

DM dry matter

DO dissolved oxygen

E-index energy index, ratio produced over consumed electricity

EUP eutrophication potential

FA free ammonia

FET freshwater ecotoxicity

FISH fluorescent in-situ hybridization

FNA free nitrous acid

GHG greenhouse gases

GWP global warming potential

HRM humane resource management

HRT hydraulic retention time

IE inhabitant equivalent

LCA life cycle assessment

MABR membrane aerated bioreactor

MBBR moving bed bioreactor

Notation index

ii

MBR moving bed reactor

MRM microbial resource management

narr nitrite accumulation rate ratio

N/DN nitrification/denitrification

NOB nitrite-oxidizing bacteria

OD ozone depletion potential

OFMSW organic fraction of municipal solid waste

OLAND oxygen-limited autotrophic nitrification/denitrification

PE person equivalent

PO photochemical oxidation potential

RBC rotating biological contactor

SBR sequencing batch reactor

SRT sludge retention time

SS suspended solids

TET terrestrial ecotoxicity

TSS total suspended solids

UASB upflow anaerobic sludge blanket

VSS volatile suspended solids

WWTP wastewater treatment plant

Table of contents

iii

Table of contents

PART I: Introduction

Chapter 1: Introduction.......................................................................................................... 3

1 Autotrophic nitrogen removal ...................................................................................................... 3

2 Cost and energy effectiveness of OLAND ................................................................................... 5

3 OLAND design parameters ............................................................................................................. 7

3.1 Choice of reactor technology ................................................................................................................... 7

3.2 Key control mechanism to obtain stable performance ............................................................. 11

3.3 OLAND performance ............................................................................................................................... 12

4 Microbial resource management (MRM) ............................................................................... 14

4.1 Maximizing nitrogen removal efficiency ......................................................................................... 14

4.2 Minimizing harmful gas emissions .................................................................................................... 17

4.3 OLAND enabling energy positive sewage treatment ................................................................. 21

5 Objectives and outlines of this research ................................................................................ 24

PART II: MRM output optimization

Chapter 2: A low volumetric exchange ratio allows high autotrophic nitrogen removal

in a sequencing batch reactor ................................................................................................ 29

1 Introduction ..................................................................................................................................... 30

2 Materials and methods ................................................................................................................. 31

2.1 OLAND SBR .................................................................................................................................................. 31

2.2 SBR cycle ....................................................................................................................................................... 32

2.3 Aerobic and anoxic batch tests ............................................................................................................ 32

2.4 Chemical analyses ..................................................................................................................................... 32

2.5 Physical aggregate characteristics ..................................................................................................... 33

3 Results ................................................................................................................................................ 33

3.1 OLAND SBR performance ...................................................................................................................... 33

3.2 Biomass morphology ............................................................................................................................... 34

3.3 Control of the microbial balance in the reactor ........................................................................... 36

4 Discussion ......................................................................................................................................... 37

4.1 OLAND SBR performance ...................................................................................................................... 37

4.2 Biomass morphology ............................................................................................................................... 38

Table of contents

iv

4.3 Control of the microbial balance in the reactor ........................................................................... 39

5 Conclusions ....................................................................................................................................... 39

6 Acknowledgements ........................................................................................................................ 39

Chapter 3: Interplay of intermediates in the formation of NO and N2O during full-scale

partial nitritation/anammox .................................................................................................. 41

1 Introduction ..................................................................................................................................... 42


2.1 Reactor operation ..................................................................................................................................... 43

2.2 Emission measurments .......................................................................................................................... 43

3 Results and discussion .................................................................................................................. 45

4 Conclusions ....................................................................................................................................... 51


PART III: Exploration of new applications

Chapter 4: OLAND maximizes net energy gain in technology schemes with anaerobic

digestion ................................................................................................................................... 55

1 Treatment of digestates by OLAND .......................................................................................... 55

1.1 Organic fraction of municipal solid waste (OFMSW) ................................................................. 57

1.2 Manure-based agricultural waste ...................................................................................................... 60

1.3 Sugar/starch-based agro-industrial waste .................................................................................... 63

1.4 Sewage-based organics .......................................................................................................................... 64

1.5 Treatment of digestates by OLAND: conclusions and perspectives .................................... 72

2 OLAND as mainstream treatment process ............................................................................ 73

2.1 Wastewater as an energy resource ................................................................................................... 74

2.2 Main stream OLAND application: conclusions.............................................................................. 76

3 General conclusions ....................................................................................................................... 76


Chapter 5: Efficient total nitrogen removal in an ammonia gas biofilter through high-

rate OLAND ............................................................................................................................ 79

1 Introduction ..................................................................................................................................... 80


Table of contents

v

2.1 Biofilter set-up and operation ............................................................................................................. 83

2.2 Profile measurements ............................................................................................................................. 83

2.3 Activity batch test ..................................................................................................................................... 83

2.4 Chemical analyses ..................................................................................................................................... 84

2.5 Quantification with real-time PCR ..................................................................................................... 85

3 Results ................................................................................................................................................ 85

3.1 Performance of the biofilter ................................................................................................................. 85

3.2 Vertical distribution of microbial activity ...................................................................................... 89

3.3 Vertical abundance of N species ......................................................................................................... 90

4 Discussion ......................................................................................................................................... 91

4.1 OLAND application for NH3 treatment ............................................................................................. 91

4.2 AnAOB niche in NH3 biofilters ............................................................................................................. 92

4.3 OLAND: gas versus water treatment ................................................................................................ 94

5 Conclusions ....................................................................................................................................... 94


Chapter 6: OLAND is feasible to treat sewage-like nitrogen concentrations at low

hydraulic residence times ...................................................................................................... 95

1 Introduction ..................................................................................................................................... 96

2 Material and methods ................................................................................................................. 100

2.1 OLAND rotating biological contactor (RBC) ................................................................................ 100

2.2 Reactor operation ................................................................................................................................... 100

2.3 Chemical analyses ................................................................................................................................... 100

2.4 Fluorescent in-situ hybridization (FISH) ...................................................................................... 101

2.5 Denaturing Gradient Gel Electrophoresis (DGGE) .................................................................... 101

3 Results .............................................................................................................................................. 102

3.1 Treatment of high nitrogen levels .................................................................................................... 102

3.2 Treatment of low nitrogen levels ..................................................................................................... 102

3.3 Suppression of nitratation at low nitrogen levels ..................................................................... 102

4 Discussion ....................................................................................................................................... 106

4.1 OLAND removal rate and efficiency treating low nitrogen levels ...................................... 106

4.2 Role of DO levels in suppressing nitratation ............................................................................... 106

4.3 OLAND operation at low HRT ............................................................................................................ 107

4.4 Implementation of OLAND in the main stream .......................................................................... 108

5 Acknowledgements ...................................................................................................................... 108

Table of contents

vi

Chapter 7: Cold OLAND on pretreated sewage: feasibility demonstration at

lab-scale ................................................................................................................................. 109

1 Introduction ................................................................................................................................... 110

2 Materials and methods ............................................................................................................... 111

2.1 OLAND rotating biological contactor (RBC) ................................................................................ 111

2.2 RBC operation .......................................................................................................................................... 112

2.3 Detection of AerAOB, NOB and AnAOB with FISH and qPCR ............................................... 112

2.4 Detailed reactor cycle balances ......................................................................................................... 113

2.5 Chemical analyses ................................................................................................................................... 113

3 Results .............................................................................................................................................. 114

3.1 Effect of temperature decrease ......................................................................................................... 114

3.2 Effect of COD/N increase ..................................................................................................................... 118

3.3 Nitratation and NO/N2O emissions ................................................................................................. 121

4 Discussion ....................................................................................................................................... 124

4.1 Effect of temperature decrease ......................................................................................................... 124

4.2 Effect of COD/N increase ..................................................................................................................... 125

4.3 NOB-AnAOB competition at mainstream conditions ............................................................... 126

4.4 OLAND application in the main line ................................................................................................ 127

5 Conclusions ..................................................................................................................................... 127


7 Supplementary data .................................................................................................................... 128

Chapter 8: Environmental assessment of one-stage partial nitritation/anammox

implementation in sewage treatment plants ...................................................................... 133

1 Introduction ................................................................................................................................... 134

2 Materials and methods ............................................................................................................... 136

2.1 Scope definition ....................................................................................................................................... 136

2.2 Plant description ..................................................................................................................................... 137

2.3 Data inventory .......................................................................................................................................... 141

2.4 Impact assessment ................................................................................................................................. 142

3 Results and discussion ................................................................................................................ 143

3.1 Impact of nitrogen removal process on process level ............................................................. 143

3.2 From energy-negative to energy-positive WWTP on system level .................................... 145

3.3 Environmental impact of DEMON implementation on life cycle level .............................. 147

4 Conclusions ..................................................................................................................................... 154

Table of contents

vii


6 Supplementary data .................................................................................................................... 155

PART IV: General discussion

Chapter 9: General discussion and perspectives .............................................................. 159

1 Main outcome and positioning of this work ....................................................................... 159

2 OLAND and sustainability .......................................................................................................... 160

2.1 Balancing energy recovery with sustainability .......................................................................... 160

2.2 Mitigation strategies based on chemical markers ..................................................................... 161

2.3 Mitigation strategies which minimize emission ........................................................................ 164

3 Energy positive WWTP: reality or fantasy? ......................................................................... 164

3.1 Water-energy nexus............................................................................................................................... 164

3.2 Is OLAND an essential treatment step? ......................................................................................... 166

3.3 Decision making for the wastewater engineer ........................................................................... 170

4 Nitrogen removal versus nitrogen recovery ...................................................................... 171

5 Future challenges and opportunities .................................................................................... 173

5.1 Future challenges for mainstream OLAND .................................................................................. 173

5.2 OLAND biofilter application ............................................................................................................... 174

5.3 What are the temperature limits of the OLAND process ........................................................ 175

6 Conclusions ..................................................................................................................................... 176

PART V: Appendices

Abstract ................................................................................................................................. 181

Samenvatting ........................................................................................................................ 185

Bibliography ......................................................................................................................... 189

Curriculum vitae .................................................................................................................. 207

Dankwoord ............................................................................................................................ 215

2

Lab-scale OLAND rotating biological contactor (RBC senior, LabMET)

Chapter 1

3

Chapter 1:

Introduction

1 Autotrophic nitrogen removal

Several new biological nitrogen removal processes have been developed to treat nitrogen-rich

wastewaters devoid in carbon such as digestates (Table 1.1). These processes are mostly

composed out of two main conversion steps; hence a one-step or two-step configuration of the

processes is possible. Performing autotrophic nitrogen removal in two stages implies that both

process steps should be optimized and controlled individually. In contrast, the investment cost

and the difficulty to balance both steps are decreased when operating the process in one step.

Moreover, full-scale application studies showed that for the one-step partial

nitritation/anammox process harmful emission of NO and N2O could be decreased to 1 and

0.001%, respectively (Desloover et al., 2011b). Because of these advantages, the full-scale

applications are all becoming one-step configurations.

Table 1.1: Overview on terminology of one-stage and two-stage autotrophic nitrogen removal

processes based on partial nitritation and anammox and indication of the amount of full-scale plants

operational at this moment. MBBR: moving bed bioreactor, SBR: sequencing batch reactor; RBC:

rotating biological contactor

Process name Stages Patent Plants Reference

SHARON-ANAMMOX 2 Yes 3 (van der Star et al., 2007)

ANAMMOX® 1 No 12 (Abma et al., 2010)

ANITATM

MOX 1 Yes 2 (Christensson et al., 2011)

Deammonification MBBR 1 No 2 (Beier and Schneider, 2008)

Deammonification SBR 1 No 3 (Joss et al., 2009)

CleargreenTM

1 No 1 (Jeanningros et al., 2010)

DEMON® 1 Yes 20 (Wett, 2006)

OLAND RBC 1 No 1 (Kuai and Verstraete, 1998)

Chapter redrafted after: Vlaeminck, S.E., De Clippeleir, H., Verstraete, W., 2012. Microbial

resource management of one-stage partial nitritation/anammox. Microbial Biotechnology.

Microbial Biotechnology, 5, 433-488.

Introduction

4

Therefore, in this work we will focus on the one-step partial nitritation/anammox process,

also known as deammonification but generally referred to as the oxygen-limited autotrophic

nitrification/denitrification (OLAND) process in this thesis. An overview of the terminology

used for pilot and full-scale applications is given in Table 1.1.

Oxygen-limited autotrophic nitrification/denitrification (OLAND) is a one-step nitrogen

removal process based on partial nitritation, performed by aerobic ammonium-oxidizing

bacteria (AerAOB) and anammox, performed by anoxic ammonium-oxidizing bacteria

(AnAOB; Fig. 1.1). The AerAOB, mainly belonging to Nitrosomonas europaea eutropha and

halophila (Vlaeminck et al., 2010), are set so that they oxidize half of the influent ammonium

to nitrite in oxygen-limited conditions (Eq 1; Table 1.2). The AnAOB, mainly members of the

Candidatus genera Kuenenia and Brocadia (van der Star et al., 2007; Vlaeminck et al., 2010),

oxidize the residual ammonium with nitrite to dinitrogen gas under anoxic conditions (Eq. 3,

Table 1.2). Consequently, in the OLAND process ammonium is converted mainly into

nitrogen gas without the use of organic carbon in one reactor. The overall stoichiometry

shows that if the AerAOB and AnAOB activity is well balanced, only 11% of the converted

ammonium is converted to nitrate due to growth of the AnAOB. Higher nitrate formation

(> 11%) implies that nitrite oxidation by nitrite-oxidizing bacteria (NOB) can take place,

probably due to an excess in oxygen (Fig. 1.1). Lower nitrate production

(< 11%) can occur when denitrification can take place due to the presence of organic carbon.

Overall nitrogen removal efficiencies obtained in full-scale one-step process are between 80

and 95% (Table 1.4), depending on the presence of COD.

Figure 1.1: Schematic overview of the balanced and imbalanced output caused by the oxic and anoxic

reactions during OLAND by aerobic and anoxic ammonium-oxidizing bacteria (AerAOB and

AnAOB) and nitrite-oxidizing bacteria (NOB).

Chapter 1

5

2 Cost and energy effectiveness of OLAND

Conventionally nitrogen is biologically removed by nitrification/denitrification (N/DN). This

process converts first all ammonium to nitrate and thereafter denitrifies nitrate with organic

carbon to dinitrogen gas. In case some COD is still present in the wastewater, this endogenous

organic carbon source can be used for denitrification. The typical composition of wastewater

COD is C5H9NO, and 1 mol can reduce 3.36 moles of nitrate (Mateju et al., 1992). Hence, a

strictly anoxic biodegradable COD/N ratio of about 4 is needed for denitrification, or about 5

taking into account some aerobic COD conversion. For wastewaters with lower COD/N

ratios, external addition of a carbon source such as methanol is needed to obtain sufficient

nitrogen removal rates. A cost-saving alternative for the latter is the application of

nitritation/denitritation, saving 40% of the operational costs. The latter is the result of the

decrease of the methanol requirement, sludge production and aeration with 50, 40 and 24%,

respectively (Table 1.2 and 1.3). Moreover, when the fully autotrophic OLAND process is

applied, 84% of the operational costs are saved, with a 100, 89 and 57% decrease in methanol

requirement, sludge production and aeration, respectively (Table 1.3). Note that savings on

methanol might in practice still be somewhat higher because of some aerobic consumption in

the presence of residual dissolved oxygen (DO). The main cause of the low sludge production

is the low biomass yield of the AnAOB. This group of bacteria has a long doubling time of

1-2 weeks (Strous et al., 1998) compared to the AerAOB (1 day) and denitrifying bacteria (in

the order of 1h). In view of energy recuperation by anaerobic digestion, OLAND can offer a

higher net energy gain because it minimizes the energy cost for further digestate treatment. It

should also be mentioned that the choice of reactor technology will also further determine the

operational and investment costs. Reactors with passive aeration, such as rotating biological

contactors for instance, have a 3 times lower energy requirement for aeration compared to

reactors with bubble aeration.

Introduction

6

Table 1.2: Overall stoichiometry for nitrification/denitrification, nitritation/denitritation and OLAND which are based on conversions of AerAOB, NOB,

AnAOB en denitrifiers (Barnes and Bliss, 1983; Mateju et al., 1992; Strous et al., 1998).

Process Equation

nr

Subreaction Stoichiometry

Nitritation (AerAOB) 1 Substrates NH4+ + 1.382 O2 + 0.091 HCO3

-

Products 0.982 NO2- + 1.891 H

+ + 0.091 CH1.4O0.5N0.2 + 1.036 H2O

Nitratation (NOB) 2 Substrates NO2- + 0.488 O2 + 0.003 NH4

+ + 0.013 HCO3

-

Products NO3- + 0.013 CH1.4O0.5N0.2 + 0.008 H2O

Anammox (AnAOB) 3 Substrates NH4+ + 1.32 NO2

- + 0.066 HCO3

- + 0.13 H

+

Products 1.02 N2 + 0.26 NO3- + 0.066 CH2O0.5N0.15 + 2.03 H2O

Denitrification (Denitrifiers) 4 Substrates NO3- + 1.080 CH3OH

Products 0.476 N2 + OH- + 0.760 CO2 + 0.325 CH1.4O0.5N0.2 + 1.440 H2O

Denitritation (Denitrifiers) 5 Substrates NO2- + 0.53 CH3OH

Products 0.48 N2 + OH- + 0.33 CO2 + 0.20 CH1.4O0.5N0.2 + 0.56 H2O

Nitrification/denitrification 1+2+4 Substrates NH4+ + 1.856 O2 + 1.058 CH3OH

Products 0.457 N2 + 1.010 H+ + 0.641 CO2 + 0.421 CH1.4O0.5N0.2 + 2.349 H2O

Nitritation/denitritation 1+5 Substrates NH4+ + 1.382 O2 + 0.52 CH3OH

Products 0.47 N2 + 0.998 H+ + 0.235 CO2 + 0.057 CH1.4O0.5N0.2 + 1.497 H2O

OLAND 1+3 Substrates NH4+ + 0.792 O2 + 0.080 HCO3

-

Products 0.435 N2 + 1.029 H+ + 0.111 NO3

- + 0.052 CH1.4O0.5N0.2 + 0.028 CH2O0.5N0.15 + 1.460 H2O

Table 1.3: Approximation of operational costs of biological nitrogen removal. Calculation factors: 0.32 EUR kg-1

methanol (Mathanex, 2011), dosed at 120%

of stoichiometric requirement to compensate for aerobic breakdown; 0.10 EUR kWhel-1

(Europe’s energy portal); 0.47 EUR kg-1

sludge dry weight (DW)

(Paul et al., 2006); 2 kg O2 kWhel-1

; personnel costs based on a medium-sized plant treating 450 kg N d-1

, requiring 1/2 full-time equivalent staff (FTE) for

operation, maintenance and repair (50 000 EUR FTE-1

yr-1

).

Process Aeration requirement Methanol addition Sludge cost Personnel Total cost Cost savings

kWh

kg-1

N

EUR

kg-1

N

kg

kg-1

N

EUR

kg-1

N

kg

kg-1

N

EUR

kg-1

N

EUR

kg-1

N

EUR

kg-1

N

%

Nitrification/denitrification 2.1 0.21 2.9 0.93 1 0.47 0.15 1.76 0

Nitritation/denitritation 1.6 0.16 1.4 0.46 0.6 0.28 0.15 1.05 40

OLAND 0.9 0.09 0 0 0.1 0.05 0.15 0.29 84

Chapter 1

7

3 OLAND design parameters

3.1 Choice of reactor technology

The application criteria including the complexity of the wastewater, the available footprint

area and the need for high level trained operators are dominating the choice of reactor

technology. Also, the operational costs and particularly the energetic aspects can further

influence the reactor type chosen. In the Table 1.5 a qualitative comparison between different

possible reactor types for applying OLAND is given. Three main categories can be

distinguished i.e. attached, immobilized and suspended growth systems. Due to the lower

complexity and low energy usage (passive aeration), attached growth systems such as rotating

biological contactors (RBC), are preferentially applied at smaller scale (Meulman et al., 2010)

or for complex wastewaters such as landfill leachates (Siegrist et al., 1998). OLAND RBC are

robust and can stably operate for years (LabMET experience). However, the flexibility of the

loading rate is limited and the oxygen balance is hard to control in contrast to suspended

growth systems or systems with carrier material in suspension where oxygen can be regulated

by controlling the aeration rate. Most full-scale OLAND-type of reactors until now are gas-lift

or sequencing batch reactors (SBR), offering efficient DO control mechanisms and high

operational flexibility. Both reactor types (gas-lift and SBR) have high biomass retention

based on well settling sludge allowing to separate the sludge at the top of the reactor with a

three phase separator or during a settling phase, respectively. However, due to the complexity

of the control mechanisms, qualified operators are needed to allow stable and highly efficient

performances in the reactor types based on suspended biomass. The ease of inoculation of

new reactors with cultivated sludge from suspended growth systems is an additional

advantage is this type of reactors and can therefore accelerate the implementation rate of the

anammox-based processes.

Introduction

8

Table 1.4: Full-scale one-stage partial nitritation/anammox applications treating digestates from industrial and municipal origin.

Place Reactor

type

Water

type

N in

(mg N L-1

)

COD/N

in

Bv

(g N L-1

d-1

)

N-Rf

(%)

Vol

(m3)

pH DO

(mg O2 L-1

)

Temp

(°C)

Sludge

(g SS L-1

)

Sludge

aggregate

Olburgen,

NL1

Airlift Potato

processing

250-350 0.6-0.8 1.8 73 600 8.0 2-3 30-35 15 Granules

China1 Airlift Glutamate

factory

600 - 2.0 >80 5000 - - - - Granules


factory

<500 - 2.0 >80 4500 - - - - Granules


factory

<500 - 2.0 >80 5350 - - - - Granules

China1 Airlift Yeast

factory

300-800 - 2.0 - 500 - - - - Granules

China1 Airlift Yeast

factory

300-800 - 2.0 - 3500 - - - - Granules

Poland1 Airlift Distillery 1000 - 2.0 - 600 - - - - Granules

Strass,

Austria2

SBR Sludge

filtrate

1800 0.57 < 1.0 90-95 500 7.0 0-0.35 30-34 3 Small

granules

Heidelberg,

D2

SBR Sludge

filtrate

1300 0.7-1.0 0.60 90-95 800 7.0 0-0.35 25-35 2 Small

granules

Glarnerland,

CH2

SBR Sludge

filtrate

1000 0.8 0.69 > 90 379 7.0 0-0.35 25-35 2 Small

granules

Plettenberg

D2

SBR Sludge

filtrate

800 0.7-1.2 0.50 > 90 134 7.0 0-0.6 25-35 2 Small

granules

Apeldoorn,

NL2

SBR Sludge

filtrate

950 0.7-1.0 0.66 > 90 2914 7.0 0-0.35 25-35 2 Small

granules

Thun, CH2 SBR Sludge

filtrate

1300 0.7-1.0 0.67 > 90 606 7.0 0-0.35 18-30 2 Small

granules

Niederglatt,

CH3

SBR Sludge

reject water

760 - 0.37 - 150 7.8 - 29 4 Flocs

Chapter 1

9

Zurich, CH3 SBR Sludge

reject water

650 - 0.45 - 2 x

1400

7.1 - 30 3.4-3.8 Flocs

Sint Gallen,

CH3

SBR Sludge

reject water

890 - 0.36 - 2 x

300

8.0 - 18-30 5.9-7.7 Flocs

Hattingen,

SE4

MBBR Sludge

reject water

503 - 0.55 63 171 7.8 3 30 13* Biofilm

Hattingen,

SE4

MBBR Sludge

reject water

275 - 1.06 52 67 7.4 3.8 30 13.6* Biofilm

Himmerfjärd

en, SE4

MBBR Sludge

reject water

/ industrial

(9/1)

776 - 0.29 74 699 8.0 - 27 6* Biofilm

Himmerfjärd

en, SE4

MBBR Sludge

reject water

/ industrial

(9/1)

497 - 0.24 59 699 7.1 - 31 5* Biofilm

Sjölunda,

SE5

MBBR Sludge

centrate

855 0.3 1.30 90 4 x 50 6.8-

7.5

0.5-1.5 22-33 - Biofilm

1 personal communication, Tim Hülsen;

2 personal communication, Bernhard Wett;

3 Joss et al. (2009);

4 Beier and Schneider (2008);

5 Christensson et al.

(2011)

* g TS L-1

Kaldness packing material

Introduction

10

Table 1.5: Qualitative comparison of OLAND reactor configurations (advantages indicated in bold). RBC: rotating biological contactor; SBR: sequencing

batch reactor; CSTR: continuous stirred-tank reactor (Vlaeminck et al., 2012)

Biomass growth Attached (biofilm) Immobilized Suspended (flocs and/or granules)

Reactor configuration Trickling filter RBC

†

Fixed/moving

Bed reactor

Fixed/moving Upflow/SBR MBR

Gas-lift

or

upflow

SBR CSTR

with

settler‡ Overall costs Low Low Medium Medium High Medium Medium Medium

Area requirement Medium High Low Low/Medium Medium Low Medium High

Aeration Passive Passive Active Active Active Active Active Active

Ease of DO control Low Medium¶ Medium/High High High High High High

Sludge content Medium Medium Medium Medium High High Low Low

Ease of biomass retention Medium Medium Medium Medium High Low Low Low

Inoculation feasibility♯ Medium Low/Medium Low/High High High High High High

Low HRT feasibility° Yes Yes Yes Yes No Yes No No

Risk for mechanical failure Medium High Low Low Medium Low Low Low

Risk for clogging High Low High/Low Low High Low Low Low

Operational flexibility Low Low Low/Medium Medium/High Medium Medium High* Medium

Operational complexity Low Low Medium Medium/High High Medium High Medium

† Biofilm can grow on rotating discs (fixed), or on carrier material brought in rotating porous cages (moving);

‡ Similar configuration as conventionally used

for activated sludge; ¶ Rotation speed can be controlled by bulk DO level (Meulman et al., 2010a);

♯ Assuming sufficient availability of enriched inoculum,

attached to carrier material if applicable; ° Important for wastewaters with low nitrogen level. For SBR and CSTR, this largely increases required settling time

or settler volume, whereas for MBR this largely increases the amount of membranes required; * Cycle duration can be adjusted to meet effluent requirements

(Siegrist et al., 2008), allowing to respond to changes in wastewater composition

Chapter 1

11

3.2 Key control mechanism to obtain stable performance

3.2.1 Balancing oxygen budget

Oxygen plays a key role in the OLAND process. One hand, enough oxygen should be present

to allow aerobic ammonium oxidation to nitrite. However, if the oxygen input is too high,

further oxidation to nitrate by NOB can take place, decreasing the overall removal efficiency.

Since AnAOB can be reversibly inhibited by oxygen concentration levels of

0.02 – 0.15 mg O2 L-1

(Strous et al., 1997), the two OLAND key players (AerAOB and

AnAOB) have opposite oxygen needs which implies a three-dimensional stratification in the

granule, floc or biofilm (Vlaeminck et al., 2010). The maximum dissolved oxygen (DO) level

experienced by the biomass can be directly controlled in most reactor technologies, except for

RBC and trickling filters. The DO can be kept at a certain setpoint or within a certain range,

with either continuous or intermittent aeration. The effect of the aeration regime on the

OLAND performance is not fully clear yet. Joss et al. (2009) showed that continuous aeration

was preferred over intermittent aeration (75% of the time aerated), because of the lower nitrite

accumulation for these conditions and the better monitoring due to the higher signal to noise

ratio when the aerators were not continuously switched on and off. In contrast to the latter

study at low DO set point (0.5 mg O2 L-1

), Zubrowska-Sudol and co-authors (2011) suggested

that an intermittent regime (66% of time aerated) was optimal at higher DO levels (2, 3, 4 mg

O2 L-1

) obtaining higher nitrogen removal rates but also higher nitrite accumulations. The

optimal DO set point is dependent on the preferred quality of the effluent, mixing conditions

in the reactor and type of biomass (oxygen gradient). For smaller granular (< 1mm, Wett,

2006) or floccular biomass (Joss et al., 2009), DO levels below 0.5 mg O2 L-1

are advisable to

avoid nitrite accumulation and development of NOB. When larger granules (2-3 mm),

allowing higher AnAOB concentrations, are used, higher DO set points can be applied up to

2 mg O2 L-1

(Abma et al., 2010). However, at these higher DO conditions, NOB can more

easily compete with the AerAOB for oxygen and can therefore form a barrier between

AerAOB and AnAOB in the granule (Vlaeminck et al., 2010).

3.2.2 pH control mechanism to obtain balanced performance

In the DEMON process, based on the same microbial conversions as the OLAND process, the

balance between AerAOB and AnAOB is obtained by a dedicated control mechanism based

on pH measurements. As the aerobic ammonium oxidation by AerAOB produces 1.9 mol H+

per mol NH4+ converted, this first reaction causes a decrease in pH which can be correlated

Introduction

12

with nitrite production. The aeration control system in this process is therefore based on a

very tight pH control interval of 0.01 units (Wett, 2006). When a pH decrease of 0.01 units is

measured, aeration is stopped and this allows depletion of the formed nitrite by AnAOB and

some recovery of alkalinity (Table 1.2). Additionally, alkaline influent water is continuously

fed to the system increasing the pH value until the upper value is reached and aeration is

switched on again. This control strategy leads to an intermittent aeration regime with DO

concentrations between 0 and 0.3 mg O2 L-1

while constant feeding is applied (Wett, 2006).

While the OLAND-type of processes are mainly applied at pH ranges between 7 and 8, this

control strategy is applied at pH value between 7.0 and 7.1 (Table 1.4).

3.2.3 Retaining sufficient microbial biomass

Since the doubling time of the AnAOB is 1-2 weeks (Strous et al., 1998), high microbial

biomass retention is a crucial factor to maintain sufficient AnAOB activity in the process. The

microbial biomass retention is most delicate in suspended growth systems where it mainly

depends on the formation of well settling sludge. In a continuously stirred tank reactor

(CSTR) or a SBR, the microbial biomass retention by settling occurs in a separate step

divided in space or time, respectively. In a SBR, biomass loss occasionally occurred due to

small N2 bubbles attached to the flocs (Joss et al., 2009) or due to foaming problems (Wett,

2006). Adjustments of the feeding strategy (Wett, 2006), the settling phase or addition of

flocculants (Joss et al., 2009) could solve this problem. Formation of both well settling flocs

and granules is possible in SBR systems (Wett, 2006; Joss et al., 2009). Formation of granules

is of utmost importance in gas-lift reactors because they depend on the continued presence of

well settling granules (Abma et al., 2010). In attached growth system, biofilm formation

allows for high biomass retention. In general, a total sludge retention time (SRT) of at least

30-45 days is recommended (Wett et al., 2010b; Desloover et al., 2011a; Joss et al., 2011).

3.3 OLAND performance

According to the reported OLAND-type of applications, the size of the reactor can be

dimensioned based on a volumetric loading rate of 0.4 to 2 g N L-1

d-1

(Table 1.4). If the

nitrogen removal rate is monitored directly by an ion-selective ammonium probe or indirectly

via conductivity measurements, the SBR cycle can be adjusted according to the obtained

removal obtaining optimal effluent quality and stable nitrogen removal rates (Joss et al.,

2009).

Chapter 1

13

Figure 1.2: Microbial resource management view on the OLAND process. AerAOB and AnAOB: aerobic and anoxic ammonium-oxidizing bacteria;

NOB: nitrite-oxidizing bacteria; GHG: greenhouse gas; bCOD: biodegradable chemical oxygen demand; GHG: greenhouse gas; DO: dissolved oxygen;

VSS: volitale suspended solids

Introduction

14

4 Microbial resource management (MRM)

The close interaction between the different microbial groups during the OLAND process is

comparable with human beings working together in firms for a shared profit. In this sense, the

concept of human resource management (HRM) can be translated to the microbial

biotechnology as Microbial Resource Management (MRM) and will therefore strive after

maintaining the best performing microbial community for a certain application (Verstraete et

al., 2007). To properly manage complex microbial systems, the engineer needs well-

documented concepts, reliable tools and a set of default values (Verstraete, 2007).

A MRM OLAND framework was elaborated, showing how the OLAND engineer/operator

(1: input) can design/steer the microbial community (2: biocatalyst) to obtain optimal

functionality (3: output), depending on the application domain (0: wastewater) (Fig. 1.2).

Taken this MRM framework in to account, the OLAND engineer can steer the OLAND

process to obtain maximum efficiency (see section 4.1) and higher sustainability (see section

4.2) or to increase the impact of OLAND on the energy balance of wastewater treatment

plants (WWTP) (see section 4.3).

4.1 Maximizing nitrogen removal efficiency

The maximum nitrogen removal efficiency that can be obtained in a balanced OLAND

system without additional denitrification is 89% (Eq. 1, Table 1.2). Lower removal

efficiencies are mainly caused by hampered nitritation resulting in residual ammonium, by an

imbalance between nitritation and anammox resulting in nitrite accumulation, or by increased

nitratation resulting in a higher nitrate production. For most OLAND applications treating

high-strength nitrogenous wastewaters, a post-treatment is obligatory to meet discharge limits.

For sewage sludge reject water treatments (Fux and Siegrist, 2004) or source-separated

black/grey-water systems (Verstraete and Vlaeminck, 2011), the OLAND effluent is sent to

the diluted treatment stream for polishing. For industrial applications, the effluent can be sent

to a sewage treatment plant (Abma et al., 2010), or can be polished by an additional separate

nitrification and denitrification stage (Desloover et al., 2011a; Tokutomi et al., 2011b). The

latter techniques are also used to polish OLAND-treated landfill leachate, and can be

complemented with an activated-carbon stage (Hippen et al., 2001; Denecke et al., 2007). A

possibility which has not been explored so far, is the inclusion of an anoxic reaction phase in

Chapter 1

15

the OLAND reactor to denitrify the nitrate produced with either autochtonous or added COD

to further increase the removal efficiency. Given the low COD/N required to remove the

remaining 11% of the nitrogen load, it is anticipated that denitrifying bacteria would not

outgrow the AnAOB.

AerAOB activity should be high enough to deliver nitrite to the AnAOB, otherwise residual

ammonium prevails (Fig. 1.3). An increase in AerAOB activity can be obtained by adjustment

of the oxygen supply and level, yet care should be taken not to use DO levels above 0.5 mg

O2 L-1

, since this will favour the development of NOB (Wett, 2006; Joss et al., 2009). It

should be noted that in systems with larger aggregates (granules), higher DO setpoints can be

applied (Volcke et al., 2010). Under more extreme conditions, high free ammonia (8 - 120 mg

N L-1

) could decrease AerAOB activity at high ammonium concentrations, high pH and

elevated temperatures, or high nitrous acid concentrations (0.2 - 2.8 mg N L-1

) could be

inhibitory at high nitrite concentrations, low pH and low temperatures (Anthonisen et al.,

1976; Fig. 1.3). However, these conditions are not likely for OLAND reactors.

Figure 1.3: OLAND MRM framework highlighting tools to obtain high nitrogen removal efficiency.

FA: free ammonia; FNA: free nitrous acid; SRT: sludge retention time

Introduction

16

If the AnAOB are not able to consume the formed nitrite or AerAOB leave not enough

ammonium to combine with nitrite, nitrite accumulation will occur, which in a more extreme

case (> 100-250 mg NO2−-N/L; Strous et al., 1999; Egli et al., 2001; Dapena-Mora et al.,

2007) can inhibit AnAOB. Besides lowering the AerAOB activity by operational parameters

such as a lower oxygen supply and level, one of the main factors to discounter the difference

in growth rate between AerAOB and AnAOB is the separation of the sludge retention of

small flocs, containing mainly AerAOB, and larger denser biomass particles, containing

mainly AnAOB (Vlaeminck et al., 2010). Different selection methods are available to

decrease the aerobic activity, depending on the applied reactor technology. Typical critical

settling velocities applied in SBR systems are 0.3 – 3 m h-1

(Chapter 2; Wett, 2006; Joss et al.,

2009). Selection for larger, denser biomass particles can therefore be based on the selective

removal of smaller particles, which have a lower density and hence lower settling velocity. In

granular upflow systems, removal of smaller, nitrifying granules at the top of the sludge bed

led to higher biomass specific conversion rates (Winkler et al., 2011). In floccular systems,

the use of hydrocyclones has been initiated to selectively maintain AnAOB-containing

granules (Wett et al., 2010b). As the AnAOB are the slowest growers in the OLAND system,

they should be maximally maintained in the system and stimulated as much as possible. It has

been shown in several studies that the AnAOB are sensitive for oxygen (Strous et al., 1997;

Egli et al., 2001). The presence of anoxic zones can also be promoted by the use of suspended

carrier material in a MBR (Beier and Schneider, 2008) or by biomass immobilization in a gel

matrix. Moreover, depending on the reactor technology applied, anoxic reactor zones can be

created in space or time. It should be noted that methanol, commonly used as exogenous

carbon source for denitrification, is detrimental for anammox (Güven et al., 2005; Dapena-

Mora et al., 2007). Besides prevention of anammox inhibition, anammox can also be

stimulated with components such as hydrazine, and dodecanoyl homoserine lactone (De

Clippeleir et al., 2011). Other operational conditions that selectively favor AnAOB activity

are not clear yet.

Nitrate accumulation due to NOB should be avoided at all time. For high-strength

wastewaters followed by a post-treatment, NOB can be suppressed in the OLAND system at

high free ammonia concentrations (> 5 mg N L−1

) and low oxygen concentrations (Vlaeminck

et al., 2009b). In the latter case, the AerAOB will have a competitive advantage over the NOB

for substrate and space. In the case of diluted wastewater systems which have to reach

effluent quality standards, free ammonia levels will not be sufficient anymore to suppress

Chapter 1

17

NOB and other methods should be searched especially for application at low temperatures

(Section 4.3). One option is the addition of compounds such as sulphide at concentrations of

20-80 mg S L-1

(Erguder et al., 2008) or chlorate at concentrations of 83-830 mg L-1

(Belser

and Mays, 1980), which have been shown to inhibit NOB activity. However, as long-term

addition of these compounds could result in adaptation and could also affect AerAOB or

AnAOB, this should be avoided as much as possible. Although Nitrospira lacks the common

protection mechanisms for reactive oxygen species (Lücker et al., 2010), the addition of

peroxide (up to 1.0 g H2O2 L−1

) had no influence on the nitratation rate of a nitrifying culture

with Nitrospira. In contrast, already at 0.5 g H2O2 L−1

, the nitritation rate was significantly

inhibited, rendering peroxide addition as an useful strategy to suppress nitritation

(Vanslambrouck, unpublished). A close interaction between AerAOB and AnAOB could also

play a role in avoiding nitratation, as the affinity of the AnAOB for nitrite is higher than the

affinity of NOB for nitrite (Lackner et al., 2008). It should be however noted that until now,

only limited knowledge exists about the genus/species dependency of these inhibition factors

and it is therefore not always straightforward to avoid nitratation.

In general, it is suggested that to obtain a balanced OLAND system with maximum nitrogen

removal efficiency, sufficient DO limitation, and a separation between the SRT of small

aerobic flocs and larger anoxic particles are desired (Fig. 1.3).

4.2 Minimizing harmful gas emissions

In terms of gaseous emissions, sustainability mainly includes minimal emissions of nitric

oxide (NO), an ozone degrader, and nitrous oxide (N2O) and methane (CH4), two potent

greenhouse gases (GHG).

Methane can be expected in the OLAND influent when treating anaerobic digestates

(dissolved at 11 g m-3

at 35°C), and small quantities might be formed in a non-aerated phase if

all oxygen and nitrate are consumed (Desloover et al., 2011a). Aeration causes stripping of

this methane. Although this can have a non-negligible contribution to the overall carbon

footprint of the process (Desloover et al., 2011a), it is difficult to prevent the emission of

dissolved influent methane, unless bubbleless aeration would be used for OLAND, as for

instance in a membrane aerated biofilm reactor (MABR; Pellicer-Nacher et al., 2010).

Introduction

18

In contrast to methane, the formation of N2O and NO occurs in situ (Fig. 1.4). As mentioned

above, for three monitored full-scale OLAND-type of systems, 0.4-1.3% of the nitrogen load

was emitted as N2O (Joss et al., 2009; Kampschreur et al., 2009a; Weissenbacher et al., 2010).

These values can be considered acceptable, since they do not significantly exceed the N2O

emission values from nitrification/denitrification (Kampschreur et al., 2009a). NO emissions

are normally ranging from negligible to 0.01% of N load (Joss et al., 2009; Kampschreur et

al., 2009a; Weissenbacher et al., 2010), but NO is due to its low water solubility easily

emitted when formed. The formation of N2O and NO is complex and often difficult to predict

due to the interplay of many parameters and contributors (Fig. 1.4).

Figure 1.4: OLAND MRM framework elaborated for the risk of N2O and NO emissions in OLAND

systems. q: specific microbial activity

Chapter 1

19

AerAOB are probably the predominant responsibles for N2O/NO emissions in OLAND,

through so-called ‘nitrifier denitrification’. The dominant energy generation method by

AerAOB is via the aerobic metabolic pathways (Chain et al., 2003). However, under oxygen

limitation or anoxic conditions AerAOB, including Nitrosomonas europaea and N. eutropha,

can use NO2- or N2O4 as electron acceptors and NH3 or H2 as electron donors to produce NO

and N2O, but no N2 (Ritchie and Nicholas, 1972; Poth and Focht, 1985; Schmidt et al., 2004).

The oxygen level and regime (i) have profound effects on N2O/NO emissions. At oxygen

concentrations below 1 mg O2 L-1

, N2O productions up to 10% of the nitrogen load were

observed (Goreau et al., 1980). While NO can be produced under both aerobic and complete

anoxic conditions (Ritchie and Nicholas, 1972; Yu et al., 2010), N2O formation by AerAOB

was only detected at aerobic or microaerophilic conditions. The N2O production by AerAOB

mainly occurs at the transition from anoxic to aerobic conditions and is coupled to the

presence of accumulated ammonium (Yu et al., 2010). Besides oxygen, nitrite concentrations

(ii) play an important role in AerAOB NO and N2O emission (Kampschreur et al., 2009b).

Nitrite accumulation is a common malfunctioning in OLAND reactors (Section 4.1), and

significantly increases AerAOB N2O emissions (Colliver and Stephenson, 2000). High N2O

production is additionally linked to high specific activity or alternately high metabolic rates

(iii) during periods with high nitrogen flux through the catabolic pathways (Yu et al., 2010).

Imbalanced enzyme expression in AerAOB performing close to their maximum specific

activity (Yu et al., 2010), would suggest that, according to the Monod kinetics, working with

a AerAOB community with lower substrate affinities (higher Ks) would yield a bigger risk of

N2O emission at lower substrate accumulations. Therefore, process configurations that work

under constant specific activity values, which are related to uniform DO and ammonium

concentrations in the reactor, are expected to produce less N2O. In this content, discontinuous

technologies such as SBR systems have more potential for N2O formation due to more

frequent transitions. Slow and long feeding during the reaction phase would result in more

stable nitrogen concentrations in the liquid phase (Wett, 2006) and could therefore potentially

lower the risk of N2O formation.

Athough ammonium oxidizing archaea (AOA) have recently been shown to produce N2O

(Santoro et al., in press), so far no AOA have been detected in OLAND systems, rendering

their contribution to N2O emissions likely nihil.

Introduction

20

Chemical formation of NO/N2O is another, potentially important pathway. An important

factor is the accumulation of the AerAOB intermediate hydroxylamine. If this compound

accumulates, it can either biochemically by AerAOB (Yu et al., 2010) or purely chemically

(van Cleemput, 1998) react with nitrite and form NO and N2O. Moreover, chemical nitrite

reduction at neutral pH can occur with ferrous iron (van Cleemput, 1998), sulfide (Grossi,

2009) or organic compounds (van Cleemput, 1998) and will also result in the formation of

NO and N2O.

It should be noted that N2O/NO emissions can also be lowered by a decrease of stripping. It

was described that NO and N2O emissions increased with the air flow rate because the

concentration of both gases remained constant in the gas phase. Therefore NO and N2O

emissions can be minimized by minimizing the airflow rate under optimal conditions

(Kampschreur et al., 2008) or by using bubbleless aeration in a MABR (Pellicer-Nacher et al.,

2010).

Although denitrification is limited in OLAND systems, typical OLAND conditions promote

NO/N2O emissions by denitrifiers. A high nitrite concentration during denitrification

suppresses the denitrification rate and therefore leads to NO and N2O accumulation (von

Schulthess et al., 1995). Also COD limitation during denitrification is a known cause for NO

or N2O accumulation (von Schulthess and Gujer, 1996; Chung and Chung, 2000). Moreover,

as oxygen inhibits both the synthesis and activity of denitrifying enzymes and N2O reductase

is the most oxygen-sensitive denitrifying enzyme (Otte et al., 1996), the low DO values

typical for OLAND can lead to N2O emission by denitrifiers.

Although NO is one of the AnAOB intermediates (Kartal et al., 2011), it is unlikely that

AnAOB leak NO, and therefore AnAOB probably do not contribute to NO emissions. Due to

the absence of N2O reductase in the AnAOB genome, N2O production is not expected during

anammox.

Overall, stable conditions allowing for constant specific microbial activities and avoiding

accumulation of nitrite and ammonium likely lead to lower NO and N2O emissions from

OLAND systems (Fig. 1.4). However, the oxygen-limited conditions needed to avoid NOB

activity or caused by well settling sludge remain a risk factor. Note that preliminary

measurements of intermittent versus continuous aeration could not point out lower N2O

Chapter 1

21

emissions for the latter (Joss et al., 2009). It is expected that future long-term, on-line

measurements will reveal the best aeration level and regime to minimize NO/N2O emissions.

4.3 OLAND enabling energy positive sewage treatment

Until now, the OLAND process has been successfully applied for medium and high-strength

nitrogen wastewaters (> 0.2 g N L−1

) such as landfill leachate and digestates from sewage

sludge, specific industrial streams and concentrated black water. For centralized domestic

wastewater treatment, the inclusion of OLAND to treat sludge digestates in the side stream of

a conventional wastewater treatment plant (WWTP) lowered the overall plant energy

requirements with about 50% (Siegrist et al., 2008). Furthermore, Wett et al. (2007)

demonstrated energy autarky by including OLAND in the sidestream of a two-stage activated-

sludge (AS) process (‘AB Verfahren’). In the mainstream, the first AS unit (A stage) has a

very high loading rate (SRT ≈ 0.5 d), and the second AS unit (B stage) has a low loading rate

(SRT ≈ 10 d). Besides these energy saving options with OLAND in a side stream, a novel

treatment scheme was recently proposed, bringing OLAND to the main treatment stream

substituting the previous B stage (Wett et al., 2010b; Verstraete and Vlaeminck, 2011). This

even allows the electrical energy recovery and savings to exceed the electrical energy input.

Moreover, instead of a biological concentration of the sewage, an enhanced physico-chemical

concentration step can be applied, involving enhanced sedimentation, dissolved air flotation

and/or membrane filtration, separating more than 75% of the COD load from the main stream

(Verstraete et al., 2009).

A first difference between treatment of the main or side stream is the lower nitrogen

concentration to be treated by OLAND (Fig. 1.5). Domestic wastewater after advanced

concentration will still contain most of the nitrogen while around 75% of the COD is removed

and sent to the digester, resulting in main stream wastewater with around 30-100 mg N L-1

and 113-300 mg COD L-1

(Metcalf and Eddy, 2003; Tchobanoglous et al., 2003; Henze et al.,

2008). Taking into account the affinity constant of the AerAOB and AnAOB for ammonium

i.e. 2.4 and 0.07 mg N L-1

respectively and the AnAOB affinity constant for nitrite of 0.05 mg

NO2--N L

-1 (Lackner et al., 2008), these low concentrations as such should not be a problem.

However, these low substrate conditions could imply that the microbial community will have

to work at lower metabolic and lower growth rates compared to side stream processes, which

allow higher concentrations in the reactor.

Introduction

22

To obtain high nitrogen removal rates at low concentrations, low hydraulic residence times

are needed for main stream treatment, in the order of hours and hence about 24 times lower

than for side stream treatment (Joss et al., 2009; Weissenbacher et al., 2010). Given the slow

biomass growth of the AnAOB, good biomass retention is a prerequisite for OLAND activity

under low HRT. Sufficient AnAOB retention can be obtained by separating the retention of

small aerobic and larger anoxic particles, which selectively will favour the AnAOB retention

(see section 4.1). On the other hand, by increasing the external settler volume, applying a

granular technology (Abma et al., 2010) or using biofilm-based technology, the total SRT can

be increased.

Besides the survival of the AnAOB under low hydraulic retention times, an important

challenge is to obtain a good microbial balance and activity at low temperature. Some studies

already described the effect of lower temperatures on the separate activity of AnAOB,

AerAOB and NOB. However, limited information exists about the microbial balance of these

three groups under OLAND conditions at low temperature. AerAOB activity decreased with

50% at a temperature interval from 27 to 15°C, yet only limited aerobic ammonium oxidation

could be observed at 5°C (Guo et al., 2010). For AnAOB the critical temperature at which it

was difficult to obtain AnAOB activity was 18°C (Dosta et al., 2008), although several

AnAOB species are found in nature at -1 to 15°C (Dalsgaard et al., 2005). It is not clear

whether other AnAOB species, more related to the cold-temperature marine genus

“Candidatus Scalindua”, will take over from the WWTP types “Candidatus Kuenenia and

Brocadia” at colder temperatures. For inoculation purposes it is important to elucidate if the

same AerAOB and AnAOB species do the job at cold temperatures or other species take over.

In the latter case, the first start-ups will be slower again due to the absence of appropriate

inoculation sources. The possible loss of both AerAOB and AnAOB activities compared to

higher temperatures will result in the accumulation of nitrite and a decrease in oxygen uptake

(Wett et al., 2010a). It will therefore be important to adjust the oxygen regime to impose

oxygen-limited conditions to the AerAOB and by this avoid inhibition of AnAOB by nitrite.

However, due to the decreased total activity, longer HRT or higher biomass concentrations

will be necessary to obtain the same volumetric nitrogen removal rates. Beside the microbial

balance between AerAOB and AnAOB, the lower temperature will have an effect on the

NOB-AnAOB balance. At temperatures lower than 15°C, the growth rate of NOB will

become higher than the growth rate of AerAOB (Hellinga et al., 1998) and it will therefore

not be possible to wash out NOB based on overall or even selective sludge retention. The

Chapter 1

23

main challenge in this application will therefore be the suppression of NOB at low

temperature and low nitrogen concentration (low free ammonia and low nitrous acid).

Figure 1.5: OLAND MRM framework elaborated to elucidate challenges for application of OLAND

in the main stream of a sewage treatment plant.

The last point of attention concerning new inputs in this application domain is the presence of

organics, i.e. moderate levels of bCOD (90-240 mg L-1

) in the wastewater. Depending on the

strength of the raw sewage, COD/N ratios between 2.4 and 3 are expected after the

concentration step, which is on the edge of the described limit for successful OLAND

Introduction

24

(Lackner et al., 2008). On the one hand, the presence of organics will facilitate DO control at

low DO levels due to heterotrophic aerobic activity. On the other hand, competition for nitrite

between heterotrophic denitrification and anammox will take place. These processes have

already been demonstrated to successfully co-exist at a COD/N ratio of 2.2 (Desloover et al.,

2011a). It is anticipated that higher nitrogen sewage levels together with the higher sewage

temperature which will facilitate OLAND treatment in the main stream, will exist in the main

stream due to further dilution preventions (Henze, 1997; Brombach et al., 2005).

Finally, according to this MRM approach (Fig. 1.5), to be able to apply OLAND in the main

stream of the WWTP, the challenges of biomass retention at low HRT and NOB suppression

at low temperature should be resolved first.

5 Objectives and outlines of this research

Altough the first OLAND applications have shown that this technology works in a stable and

efficient way (Table 1.4), the implementation rate of this technology remains dependent on a

few companies. Many potential users hold back because it seems that due to the long start-up

periods for the first reactors and the reported sensitivities, a lot of experience is needed to

keep this process running. To overcome this problem, the output box of the MRM framework

was further studied in detail for high-strength nitrogen containing wastewaters (known

application) in Part II of this work. In Chapter 2 the effect of the hydraulic conditions on the

start-up of the OLAND SBR was studied. Furthermore, strategies to obtain a well-balanced

OLAND system were proposed. As not only the effluent quality, but also the sustainability

can be a competitive factor to choose an environmental technology, the N2O and NO

emissions were studied in a full-scale OLAND reactor in Chapter 3. The relation between the

N2O/NO emission and the accumulation of substrates/intermediates and changes in the

operational conditions was elaborated.

In Part III of this work new opportunities were explored for the OLAND process (box 0 of

MRM framework). In Chapter 4, the impact of OLAND on the total energy balance was

calculated for industrial, communal and agricultural applications to elucidate where

opportunities for OLAND can be found, regarding energy efficiency. From Chapter 4 it

became clear that wastewater treatment of manure-based wastestreams is very complex and

therefore OLAND implementation will depend on specific cases. However, this sector also

Chapter 1

25

has nitrogen-rich gaseous emissions, i.e. ammonia streams, which are mostly treated

inefficiently. Therefore in Chapter 5, the possibility of OLAND to treat gaseous ammonia

streams, instead of water streams, was tested in a biofilter. Another opportunity for OLAND,

based on the energy calcutations of Chapter 4, was the implementation of OLAND in the

mainstream of the municipal WWTP. This application domain was step by step elaborated. In

Chapter 6, the OLAND performance at low nitrogen concentrations and low HRT was tested

in a RBC at 34°C as a first preriquisite for mainstream OLAND. In Chapter 7, the same RBC

was used and was adapted to lower tempatures (up to 15°C) and the presence of organics

(COD/N ratio of 2) to simulate at lab-scale the OLAND performance at mainstream

conditions. Based on a full-scale trial to implement OLAND in the mainstream at a WWTP in

Strass (Austria), a life cycle analysis (LCA) was performed. This analysis was used to

evaluate the effect of OLAND implementation in side and mainstream on a process, plant and

life cycle level (Chapter 8).

In Chapter 9 (Part IV), the results obtained are discussed in the framework of the research

objectives. Conclusions are drawn and perspectives of further research are presented.

Introduction

26

28

Floating hood for greenhouse gas emission measurements (WWTP Strass, Austria)

Chapter 2

29

Chapter 2:

A low volumetric exchange ratio

allows high autotrophic nitrogen

removal in a sequencing batch

reactor

Abstract

Sequencing batch reactors (SBRs) have several advantages, such as a lower footprint and a

higher flexibility, compared to biofilm-based reactors, such as rotating biological contactors.

However, the critical parameters for a fast start-up of the nitrogen removal by oxygen-limited

autotrophic nitrification/denitrification (OLAND) in a SBR are not available. In this study, a

low critical minimum settling velocity (0.7 m h−1

) and a low volumetric exchange ratio (25%)

were found to be essential to ensure a fast start-up. To prevent nitrite accumulation, two

effective actions were found to restore the microbial activity balance between aerobic and

anoxic ammonium-oxidizing bacteria (AerAOB and AnAOB). A daily biomass washout at a

critical minimum settling velocity of 5 m h−1

removed small aggregates rich in AerAOB

activity, and the inclusion of an anoxic phase enhanced the AnAOB to convert the excess

nitrite. This study showed that stable physicochemical conditions were needed to obtain a

competitive nitrogen removal rate of 1.1 g N L−1

d−1

.

Chapter redrafted after: De Clippeleir, H., Vlaeminck, S.E., Carballa, M., Verstraete, W.,

2009. A low volumetric exchange ratio allows high autotrophic nitrogen removal in a

sequencing batch reactor. Bioresource Technology, 100, 5010-5015.

A low volumetric exchange ratio allows high autotrophic N removal in a SBR

30

1 Introduction

Despite the economical advantage of the AnAOB-based processes, such as OLAND, in

comparison with the conventional nitrification/denitrification, these processes are hindered by

the long start-up period due to the slow growth rate of the AnAOB, which have a doubling

time of 7 to 14 days (Strous et al., 1998). Therefore, biomass washout has to be minimized,

e.g. by biofilm formation or granulation. Several biofilm-based reactors, such as a rotating

biological contactor (RBC; Siegrist et al., 1998; Pynaert et al., 2004), a moving bed reactor

(Cema et al., 2006) or a fixed bed reactor (Furukawa et al., 2006) have already been

successfully applied. High biomass retention can also be obtained in a sequencing batch

reactor (SBR) operated at a critical minimum biomass settling velocity. The latter is defined

as the ratio between the settling time and the vertical distance of the water volume decanted

per cycle, and it can also be expressed as the volumetric exchange ratio, i.e. the ratio of the

decanted to the total water volume. Reported minimum biomass settling velocities for

OLAND type SBRs are in the range of 0.3-0.7 m h−1

(Third et al., 2001; Sliekers et al., 2002;

Wett, 2006; Vlaeminck et al., 2009a). Although SBRs have advantages, such as a lower

footprint and a higher flexibility, compared to biofilm based reactors, such as RBC, so far the

nitrogen removal rates obtained in these reactors are almost five times lower (Table 2.1).

Not only efficient biomass retention is required for a successful OLAND process, a good

balance between the AerAOB and AnAOB is needed as well. A higher activity of the

AerAOB in comparison to the AnAOB results in nitrite accumulation in the reactor, which

can inhibit the AnAOB activity at nitrite concentrations of 98 to 350 mg NO2−-N L

−1 (Strous

et al., 1999; Dapena-Mora et al., 2007). While in RBCs the microbial balance is equilibrated

spontaneously due to the limited penetration depth of oxygen in the biofilm, the control of this

microbial balance in SBRs is not straightforward. Two kinds of biomass morphologies, flocs

and granules, were mainly present in suspended growth systems (Innerebner et al., 2007;

Vlaeminck et al., 2010). Granules can be described as compact and dense aggregates with a

high macroscopic circularity that do not coagulate under reduced hydrodynamic shear and

settle significantly faster than flocs (Lemaire et al., 2008). Flocs were found to be enriched in

AerAOB, while AnAOB were dominant in the granules (Nielsen et al., 2005; Vlaeminck et

al., 2009a; Vlaeminck et al., 2010). Therefore, the overall balance between the AerAOB and

AnAOB is dependent on the biomass morphology distribution in the reactor. Morphology

selection on the basis of the settling velocity could therefore improve the microbial balance.

Chapter 2

31

Table 2.1: Overview of the volumetric nitrogen removal rates in OLAND type rotating biological

contactors (RBC) and sequencing batch reactors (SBR).

Reactor

type

Volume

(m³)

Nitrogen removal rate

(kg N m−3

d−1

) Reference

RBC 33 0.4 Siegrist et al. (1998)

RBC 0.044 1.1 Pynaert et al. (2003)

RBC 240 1.7 Schmid et al. (2003)

RBC 0.005 1.8 Pynaert et al. (2004)

SBR 0.002 0.1 Third et al. (2001)

SBR 0.002 0.3 Sliekers et al. (2002)

SBR 0.002 0.5 Vlaeminck et al. (2009a)

SBR 500 0.6 Wett (2006)

SBR 0.002 1.1 This study

Gas-lift 0.002 1.5 Sliekers et al. (2003)

In this study, the microbial balance between the AerAOB and AnAOB was evaluated in an

OLAND SBR. The critical parameters for a fast start-up were determined and strategies to

control the microbial balance and enhance the biomass retention in the reactor were evaluated.

2 Materials and methods

2.1 OLAND SBR

The lab-scale OLAND SBR consisted of a cylindrical vessel with an internal diameter of

14 cm (working volume of 2.5 L). The reactor was inoculated with OLAND biomass

harvested from the reactor described by Pynaert et al. (2003) at an initial biomass

concentration of 2.3 g VSS L−1

. The reactor was fed with synthetic wastewater containing an

initial ammonium concentration of 100 mg N L−1

, 10 mg KH2PO4-P L−1

and 2 mL L−1

of a

trace elements solution (Kuai and Verstraete, 1998). To provide both buffering capacity and

inorganic carbon, 1 mole of bicarbonate was added per mole of nitrogen. If necessary, the

latter ratio was increased temporarily to ensure that the reactor pH did not drop below 7.4. In

addition, the influent ammonium concentration was gradually increased whenever the effluent

concentration was below ca. 25 mg N L−1

. The reactor was mixed with a magnetic stirrer at

245 rpm and aerated at an airflow rate of 40 L h−1

. The temperature and the dissolved oxygen

(DO) concentration were controlled automatically at 33 ± 1°C (temperature controlled room)

and 0.3 to 0.7 mg O2 L−1

(Oxymax W COS31 probe with Liquisis M COM 223 controller;

Endress & Hauser, Reinach, Switzerland), respectively. Three different phases of operation

were carried out: phase 1 with high volumetric exchange ratio (40%) and high critical

minimum settling velocity (2 m h−1

); phase 2 with high volumetric exchange ratio (40%) and


32

low critical minimum settling velocity (0.7 m h−1

); and, phase 3 with low volumetric

exchange ratio (25%) and low critical minimum settling velocity (0.7 m h−1

). During the first

and the second phases, the exchangeable volume was fixed at 1 L, resulting in a volumetric

exchange ratio of 40% (1/2.5). During the third phase, the exchangeable volume was reduced

to 0.5 L (by lowering the working volume to 2 L), and consequently, the volumetric exchange

ratio decreased to 25% (0.5/2). The nitrogen compounds (ammonium, nitrite, nitrate), DO

concentrations and pH were monitored during the whole experiment.

2.2 SBR cycle

The SBR was operated with 1h cycles during the whole experimental period. During the first

phase, 1 L of synthetic medium was fed to the reactor during a 5 minutes filling period. The

reactor was mixed and the DO was controlled both during the feeding and the reaction phase.

Subsequently, the biomass was allowed to settle for 2 minutes, so that the minimum biomass

settling velocity was 2 m h−1

. Finally, an effluent pump removed the supernatant. During the

second and third phases, the settling time was increased to 6 and 3 minutes, respectively,

resulting in a lower selection pressure (critical minimum settling velocity of 0.7 m h−1

).

2.3 Aerobic and anoxic batch tests

The specific activities of AerAOB and AnAOB were determined in aerobic and anoxic batch

tests, respectively, as described in detail by Vlaeminck et al. (2007). Prior to the activity tests,

the biomass was washed with a phosphate buffer (100 mg P L−1

; pH 8) on a sieve (pore size

50 µm) to remove residual dissolved reactor compounds. The aerobic tests were performed in

open Erlenmeyer with ammonium as substrate. For the anoxic tests, biomass incubation

occurred in a gas-tight anoxic serum flask with ammonium and nitrite as substrates. Both tests

were performed on a shaker at 34 ± 1°C.

2.4 Chemical analyses

Nitrite and nitrate were determined on a Metrohm 761 Compact Ion Chromatograph

(Zofingen, Switzerland) equipped with a conductivity detector. Ammonium (Nessler method)

was measured according to standard methods (Greenberg et al., 1992). The pH was measured

with a Consort C532 pH meter (Turnhout, Belgium).

Chapter 2

33

2.5 Physical aggregate characteristics

A mixed liquor sample obtained in a Petri dish was photographed with a high resolution

(10 megapixels) digital camera for particle analyses. The Feret diameter (largest diameter in

irregular particle), the circularity of the biomass aggregates and settling velocity were

determined as described by Vlaeminck et al. (2009a).

3 Results

3.1 OLAND SBR performance

During phase 1 the strategy of the OLAND SBR operation was based on a high critical

minimum settling velocity of 2 m h−1

to induce granulation. These conditions resulted in an

average total nitrogen removal rate of only 20 mg N L−1

d−1

, which was attributed to complete

conversion of ammonium to nitrite (Fig. 2.1B). Moreover, no anammox activity was observed

during this phase. Therefore, it was concluded that a critical minimum settling velocity of 2 m

h−1

was too high.

At the beginning of phase 2 (day 46), 1.6 g OLAND biofilm-VSS L−1

was added and the

critical minimum settling velocity was decreased to 0.7 m h−1

. These actions resulted in a

higher anammox activity since the nitrite production was lower compared to the ammonium

consumption (Fig. 2.1B), and consequently, a higher total nitrogen removal rate (around 135

mg N L−1

d−1

) was obtained (Fig. 2.1A). However, the nitrite production rate was still high

(around 88 mg N L−1

d−1

) and no improvement of the total nitrogen removal over time was

obtained.

In the subsequent phase 3 (day 95), the critical minimum settling velocity was kept constant

(0.7 m h-1

), but the volumetric exchange ratio was decreased from 40 to 25%. Similar to phase

2, extra biomass (1.6 g OLAND biofilm-VSS L−1

) was added. These changes resulted in a

steep and continuous increase of the total nitrogen removal rate and in a stable nitrate

production (Fig. 2.1A). The fraction of nitrate produced in comparison with the net

ammonium consumed was around 11%, which was in accordance with the expected nitrate

production in the OLAND process (Strous et al., 1999). At the end of the experiment, a

competitive removal rate of 1.1 g N L−1

d−1

was obtained.


34

Figure 2.1: Performance of the OLAND SBR, subdivided in three experimental phases. Phase 1: high

critical minimum settling velocity (2 m h−1

) and high volumetric exchange ratio (40%). Phase 2: low

critical minimum settling velocity (0.7 m h−1

) and high volumetric exchange ratio (40%). Phase 3: low

critical minimum settling velocity (0.7 m h−1

) and low volumetric exchange ratio (25%). A Nitrogen

removal rate and relative nitrate production. B Ammonium consumption and nitrite production.

3.2 Biomass morphology

The inoculum of the OLAND SBR consisted of biofilm pieces originating from an OLAND

RBC. Due to mixing, these aggregates disintegrated resulting in a variety of biomass particles

(Table 2.2). To improve the biomass retention in the SBR, granulation was pursued.

Chapter 2

35

Therefore, a stronger selection pressure (2 m h−1

) was applied during phase 1, but this action

did not result in enhanced granulation. In addition, the particle size distribution analyses did

not show significant change during this phase (data not shown). From phase 3 on, the particle

size distribution shifted progressively to the situation of day 135, when granules could be

detected among other smaller biomass aggregates (Table 2.2). The fraction of granules

increased during the operation period to an average of 20% by the end of the experiment.

Table 2.2: Distribution of biomass fractions at the start-up (day 1), when granules were present

(day 135) and when nitrite accumulation occurred (day 161).

Time Biomass fraction per size class (mm)

< 0.5 0.5 – 1 1 – 1.5 >1.5

Day 1 (start-up) 0.10 0.59 0.11 0.20

Day 135 (granule formation) 0.11 0.45 0.24 0.20

Day 161 (nitrite accumulation) 0.22 0.32 0.20 0.25

Two kinds of granules (red and brown) with different characteristics could be distinguished

(Table 2.3). Although the red granules were more uniformly distributed while the brown

granules had a high variety of sizes (data not shown), the average Feret diameter was not

significantly different. The red granules had a high circularity and good settling properties.

Moreover, the red granules were perfectly balanced in activity in contrast with the brown

granules which had an excess AerAOB activity (Fig. 2.2). The equilibration in aerobic and

anoxic activity can also be represented by the nitrite accumulation rate ratio (narr), defined as

the ratio of the net aerobic nitrite production rate to the anoxic nitrite consumption rate

(Vlaeminck et al., 2010). The narr of the red granules, brown granules and the inoculum of

SBR was 1.2, 3.8 and 1.4, respectively.

Table 2.3: Characteristics of the red and brown granules obtained in the OLAND SBR. The

significantly different parameters are indicated with a star (p<0.01).

Red granules Brown granules

Feret diameter (mm) 2.39 ± 0.32 2.49 ± 0.62

Circularity (-) 0.75 ± 0.11 0.65 ± 0.14 *

Settling velocity (m h−1

) 55 ± 10 35 ± 8 *


36

3.3 Control of the microbial balance in the reactor

Parallel to the increase of the total nitrogen removal rate, an increase in nitrite production rate

up to a maximum of 377 mg NO2−-N L

−1 d

−1 on day 116 was observed. As the AerAOB grow

a factor 10 faster than the AnAOB (Jetten et al., 2001), the nitrite production can never be

kept low if no external actions are taken. Particle size analyses indicated that the fraction of

biomass particles smaller than 0.5 mm increased when nitrite accumulation was detected

(Table 2.2), suggesting that these small particles were related to the AerAOB. Since AnAOB

can loose activity at nitrite concentrations in the range of 98 to 350 mg NO2−-N L

−1 (Strous et

al., 1999; Dapena-Mora et al., 2007), a daily selection (every 24h, not every cycle)

corresponding to a critical minimum settling velocity of 5 m h−1

(settling time of 23 seconds)

and an anoxic phase of 10 minutes at the end of the reaction period were included from day

122 on. This occasional higher selection could wash the excess AerAOB and the anoxic phase

could enhance the AnAOB activity. As a consequence, nitrite production decreased and

remained low during the rest of the experimental run (Fig. 2.1).

Figure 2.2: Specific aerobic and anoxic activity and nitrite accumulation rate ratio (narr) of the SBR

inoculum, SBR biomass mixture, the biomass washed out at a critical settling velocity of 10 m h−1

(n=3), the brown and red granules (n=1). Calculation and interpretation of narr is presented in section

‘Results’.

To test the individual effect of the daily selection on the suppression of the nitrite

accumulation, both actions (daily selection and anoxic phase) were left out on day 156,

resulting in an increase of nitrite in the reactor. On day 161 a selection corresponding to a

critical minimum settling velocity of 10 m h−1

was performed and the AerAOB and AnAOB

activity of the washed biomass was determined (Fig. 2.2). It could be observed that the

Chapter 2

37

washout fraction had a higher AerAOB activity (narr 4.1). Together with the small particles,

a small fraction of AnAOB was removed from the reactor as well, since AnAOB activity was

detected in the washed biomass (Fig. 2.2).

The single effect of the anoxic phase could be determined by monitoring the concentrations of

the nitrogen compounds during a SBR cycle (Fig. 2.3). The consumption of ammonium

occurred during the complete cycle and nitrate was simultaneously produced. However, nitrite

accumulated linearly (R²: 0.99) during the reaction period and it was only sufficiently

consumed during the anoxic phase (23% of reaction period).

Figure 2.3: Evolution of ammonium, nitrite and nitrate concentrations during 1 cycle in the OLAND

SBR with 5 periods (day 143): feeding phase (1), reaction phase (2), anoxic phase (3), settling period

(4) and withdrawal (5).

4 Discussion

4.1 OLAND SBR performance

During the first two phases no improvement of the nitrogen removal could be obtained.

However, during phase 3 a steep increase of the nitrogen removal was detected. This increase

in removal rate was attributed to the low volumetric exchange ratio of 25%, because the

addition of extra biomass (1.6 g OLAND biofilm-VSS L−1

) at the beginning of phase 2 had no

effect. To link low volumetric exchange ratio with good start-up, two hypotheses are

formulated. Firstly, the hydraulic retention time was increased from 2.5 to 4 h when lowering

the volumetric exchange ratio. However, these results contradict some results found in


38

literature, showing that nitrogen removal rates did not increase by applying higher hydraulic

retention times (Third et al., 2001; Sliekers et al., 2002; Tsushima et al., 2007b). However, the

different operational and dimensional conditions between the different studies make the

evaluation of the effect of the hydraulic retention time on the start-up and performance of

OLAND type reactors rather difficult. Secondly, a lower volumetric exchange ratio yielded

more stable hydraulic and chemical conditions in the reactor. The variation in the nitrogen

concentrations, metabolic products and shear rates in the reactor between the beginning and

the end of the feeding period were 1.3-fold instead of 1.7-fold in phase 1 and 2. Thus, the

chemical and physical stress was lower, resulting in a more stable and continuous-like

process. These stable physicochemical conditions could explain why higher nitrogen removal

rates up to 2 g N L−1

d−1

can be obtained in continuous reactors, such as RBC and airlift

reactors (Sliekers et al., 2003; Pynaert et al., 2004). Although better performances have been

reported in these continuous OLAND reactors, the nitrogen removal rate obtained in this

study was exceptionally high compared with other lab- and full-scale OLAND-type SBRs

(Table 2.1).

4.2 Biomass morphology

To granulate active or nitrifying sludge, a minimum settling velocity of 4.5 m h-1

is required

(Liu et al., 2005). In this study, granulation was obtained at a critical minimum settling

velocity of only 0.7 m h1. Other researchers detected granulation of OLAND type of biomass

at similar critical settling velocities (Innerebner et al., 2007; Vlaeminck et al., 2009a) and,

similar to Innerebner et al. (2007), granulation was only detected after a good nitrogen

removal had been obtained. This fact suggests that the granulation process of OLAND

biomass is different from the aerobic granulation, where a strong selective settling pressure is

a prerequisite (Liu et al., 2005).

Red and brown granules were detected in the OLAND SBR. These two types of granules had

similar sizes, but different physical and microbial properties. Red granules had a high

circularity and settling velocity, resulting in efficient biomass retention. Moreover, these red

granules had a narr between 0.6 and 1.4, indicating that these aggregates could perform the

OLAND process autonomously (Vlaeminck et al., 2010). The brown granules had an excess

AerAOB activity resulting in a high narr value (3.8). Thus, a high circularity and fast settling

of the biomass aggregates in combination with narr values around 1 is preferable in the

OLAND process.

Chapter 2

39

4.3 Control of the microbial balance in the reactor

Parallel with the steep increase in total nitrogen removal, nitrite accumulation occurred. In

this study, a combination of a daily selection and an anoxic phase could restore the balance.

The effect of the daily selection was confimed in an activity test. This test showed that the

aerobic ammonium oxidizers were dominant in the small fraction (narr of 4.1), but also a

small fraction of anammox bacteria was present. Calculated back to the effect on the reactor

performance, one selection decreased the VSS content, the aerobic and the anoxic activity

with 2, 3 and 0.3%, respectively. Although these percentages are small, the nitrite

accumulation could be suppressed sufficiently. However, this daily selection resulted in sharp

decreases of the nitrogen removal rates (Fig. 2.1), thus indicating that this effect must be

modulated on a long-term basis.

The second action to avoid nitrite accumulation and, consequently the inhibition of the

anammox bacteria was the insertion of an anoxic phase. The anammox reaction was

predominant during the anoxic phase because the nitrate produced per ammonium consumed

(23%) was similar to the relative nitrate production of the anammox reaction, i.e. 26% (Strous

et al., 1999). Therefore, the anoxic phase was effective to control the nitrite concentrations,

but the long-term effect of this action is difficult to asses.

5 Conclusions

A low selection pressure, corresponding to a critical minimum settling velocity of 0.7 m h−1

,

combined with a low volumetric exchange ratio of 25% and an equilibrated microbial activity

were essential to obtain a competitive removal rate of 1.1 g N L−1

d−1

. Besides the better

settling properties, the red OLAND granules were well balanced in activity, and thus more

suitable for a stable operation compared to the brown granules and the small aggregates.

However, without a dominance of red granules, actions should be taken to avoid nitrite

accumulation.

6 Acknowledgements

This research was funded by a PhD grant for Haydée De Clippeleir from the Institute for the

Promotion of Innovation through Science and Technology in Flanders (IWT-Vlaanderen, SB-

81068), by a PhD grant (Aspirant) for Siegfried E. Vlaeminck from the Fund of Scientific


40

Research-Flanders (Fonds voor Wetenschappelijk Onderzoek (FWO) Vlaanderen), and by a

postdoctoral contract for Dr. Marta Carballa from the Xunta de Galicia (Isidro Parga Pondal

program, IPP-08-37). The authors gratefully thank Bart De Gusseme and Peter Aelterman for

the inspiring scientific discussions.

Chapter 3

41

Chapter 3:

Interplay of intermediates in the

formation of NO and N2O during

full-scale partial nitritation/anammox

Abstract

Next to energy- and cost-efficiency, sustainability is evolving as a benchmark for wastewater

treatment. Taking into account the high global warming potential of nitrous oxide (N2O),

minimization of its emission is gaining attention. As the formation of N2O and nitric oxide

(NO) is complex and relies on the interplay of different intermediates, such as nitrite (NO2-)

and hydroxylamine (NH2OH), a detailed monitoring of all nitrogen species in both the gas

and liquid phase was performed in this study. The aim was to find a link between measurable

N components, operational conditions and the NO/N2O emissions from a full-scale OLAND-

type reactor. High loading rates, resulting in highly dynamic cycles with rapid on/off aeration

regimes, resulted in higher NO and N2O emissions, indicating that transient conditions favour

both N2O and NO emission. Therefore, the beginning of a cycle during which most changes

in operational conditions occurred was studied in detail. At the beginning of the cycle a lag

phase in N2O and NO (30 and 15 min., respectively) emission was measured. Sudden peaks in

ammonium oxidation rate up to 335 kg d-1

were accompanied with transient accumulations of

NH2OH (up to 0.001% of NH4+ consumption) and/or NO2

- (up to 0.2% of NH4

+ consumption)

and resulted in N2O and NO emission peaks. Despite the complex interplay of many factors,

this study showed that NH2OH accumulation and NO/N2O emission can be correlated

positively. Therefore, a better understanding of the conditions leading to NH2OH

accumulation could help to find strategies to minimize N2O and NO emission.

Chapter redrafted after: De Clippeleir H., Weissenbacher N., Boeckx P., Chandran K., Boon

N. and Wett B. 2012. Interplay of intermediates in the formation of NO and N2O during full-

scale partial nitritation/anammox. Ecotechnologies for wastewater treatment, Santiago de

Compostela, Spain.

Interplay of intermediates in formation of N2O/NO during OLAND

42

1 Introduction

Besides cost- and energy-efficiency, the sustainability of the process is gaining more and

more attention. Since 1 kg N2O has the global warming potential of 298 kg CO2 on a 100-yr

time horizon (Solomon et al., 2007), the N2O emissions can have a huge impact on the CO2

footprint of a wastewater treatment plant (WWTP). Moreover, NO2 and NO emission

contribute to the formation of tropospheric ozone and can cause acidification (Solomon et al.,

2007). The formation of N2O and also NO occurs in situ. Recently, some studies have

specifically addressed N2O emission from full-scale OLAND-type of systems, showing that

0.4-1.3% of the nitrogen load was emitted as N2O (Joss et al., 2009; Kampschreur et al.,

2009a; Weissenbacher et al., 2010). These values can be considered acceptable, since they do

not significantly exceed the N2O emission values from nitrification/denitrification

(Kampschreur et al., 2009a). NO emissions during OLAND are normally ranging from

negligible to 0.01% of N load (Joss et al., 2009; Kampschreur et al., 2009a; Weissenbacher et

al., 2010). However, NO is due to its low water solubility easily emitted when formed. The

formation of N2O and NO is complex and often difficult to predict due to the interplay of

many parameters, contributors and mechanisms within the contributors (simplified overview

in Fig. 3.1).

It is believed that the decrease in NO and N2O emission can be accomplished by optimization

of the operational parameters. However to do so, a better understanding of the role of the

different NO and N2O producing pathways in in situ conditions, characterized by an interplay

of AerAOB, AnAOB, NOB and denitrifier activities under changing operational conditions, is

needed. In this study, a detailed follow-up of all nitrogen species in liquid and gas phase was

performed with the aim to link the presence of intermediates with NO and N2O formation.

This study was performed on a full-scale OLAND-type of reactor, more specifically a

DEMON SBR in the side line of the WWTP of Strass (Wett, 2006).

Chapter 3

43

NH4+

NO2− NO3

−

O2

N2

NON2O

Nitratation (NOB)

O2

Anammox (AnAOB)

NH2OH

Denitrification (HDN)

Nitritation (AerAOB)

Chemical reaction

NH4+

NO2− NO3

−

O2

N2

NON2O

Nitratation (NOB)

O2

Anammox (AnAOB)

NH2OH



NH4+

NO2− NO3

−

O2

N2

NON2O

Nitratation (NOB)

O2

Anammox (AnAOB)

NH2OH



Figure 3.1: Nitrogen conversion in relation to NO and N2O formation.


2.1 Reactor operation

A full-scale DEMON sequencing batch reactor (SBR, 500 m3) treating sludge digestor

supernatant at the municipal WWTP in Strass, Austria (Wett, 2006) was monitored in this

study. The SBR was operated in cycles of 6 hours of which 75% of the time the oxygen-

limited reaction phase took place. During this phase the reactor was continuously fed and the

balance between AerAOB and AnAOB activity was obtained by a dedicated control

mechanism based on pH measurements. As the aerobic ammonium oxidation by AerAOB

produces 1.9 mol H+ per mol NH4

+ converted, this first reaction causes a decrease in pH,

which can be correlated with nitrite production. The aeration control system in this process is

therefore based on a very tight pH control interval of 0.01 units (Wett, 2006). When a pH

decrease of 0.01 units is measured, aeration is stopped and this allows depletion of the formed

nitrite by AnAOB and some recovery of alkalinity. Additionally, alkaline influent water is

continuously fed to the system increasing the pH value until the upper value is reached and

aeration is switched on again. This control strategy leaded to an intermittent aeration regime

with DO concentrations between 0 and 0.7 mg O2 L-1

while constant feeding is applied (Wett,

2006).

2.2 Emission measurments

To allow continuous off-gas measurements and to control foam formation, a cylinder

(diameter 0.3 m, height 2 m) was vertically placed into the reactor and a defined air stream


44

(0.7 ± 0.6 m s-1

) was blown into the cylinder. Therefore, the in situ emitted gas concentrations

were diluted with at least a factor 2 as during aeration periods maximum gas velocities of

1.8 m s-1

were detected. Gaseous N2O concentrations were measured online at a time interval

of 3 minutes with a photo-acoustic infrared multi-gas monitor (Brüel & Kjær, Model 1302,

Nærem, Denmark). NO was measured online using a chemiluminescense analyzer (APNA

350, Horiba, Japan) and recorded at one minute intervals. For dissolved N2O measurements, a

1 mL filtered (0.45 μm) sample was brought into a 7 mL vacutainer (-900 hPa) and measured

afterwards by pressure adjustment with He and immediate injection at 21°C in a gas

chromatograph equipped with an electron capture detector (Shimadzu GC-14B, Japan).

Ammonium concentration (Nessler method) in the water phase was determined according to

standard methods (Greenberg et al., 1992). Nitrite and nitrate were determined on an ion

chromatograph equipped with a conductivity detector (Metrohm, 761 compact, Zofingen,

Switzerland). Hydroxylamine was determined spectrophotometrically (Frear and Burrell,

1955). The N2O and NO fluxes of the full-scale reactor were based on the measured off-gas

concentration corrected for the background concentration in the defined air stream and

converted to molar concentration with the ideal gas law at the measured temperature and

atmospheric pressure. Multiplication of the measured gas velocity of the air stream and the

cross section area of the outlet of the cylinder (28 cm2) yielded the off-gas flow rate.

Figure 3.2: Picture of the set-up for greenhouse gas emissions in the DEMON reactor

Chapter 3

45

3 Results and discussion

The NO, NO2 and N2O emission from the full-scale OLAND-type SBR was measured at two

different loading rates i.e. 247 kg N d-1

and 107 kg N d-1

(Fig. 3.3). In both cases, the airflow

rate, mixing rate and operational conditions such as DO, pH and influent quality were kept

constant, as no changes were made in set-points and control mechanisms and only the feeding

rate was changed. The effluent quality in both cycles changed from COD, NO3--N, NO2

--N

and NH4+ concentration of 632, 49, 1 and 60 mg L

-1 to 632, 48, 2 and 15 mg L

-1 for the high

and low loading conditions, respectively. As the digestate contained a considerable amount of

inorganic carbon, a lower loading rate caused lower CO2 emissions and CO2 emissions

rapidly followed the aeration regime (Fig. 3.3). To obtain a similar pH and DO pattern at

higher loading rate, a more transient operation was imposed characterized by rapid on/off

aeration regimes.

The full-scale reactor emitted at the high loading rate 3.5 kg N2O-N d-1

, 33 g NO-N d-1

and

6.7 g NO2-N d-1

, which corresponded to 1.4, 0.02 and 0.003% of the nitrogen load,

respectively. These emissions were in the expected range according to literature (Joss et al.,

2009; Kampschreur et al., 2009a; Weissenbacher et al., 2010). As the effluent COD and

nitrate concentration were constant at lower loading rate compared to the higher loading rate

and an increased nitrite and decreased ammonium effluent concentration was observed,

similar or higher relative N2O and NO emissions were expected. However, at the lower

loading rate, characterized by longer anoxic periods, N2O emissions decreased until 0.37 kg

N2O-N d-1

or 0.3% of N load. NO was more easily stripped out, but because longer anoxic

phases were applied, the total emission was lowered until 6 g NO-N d-1

or 0.01% of the N

load. As a constant aeration flow rate was maintained during the SBR cycles, the lower

emission was caused by a lower concentration of N2O and NO concentration that was emitted.

The aeration during the SBR cycle was controlled by measuring a decrease in pH during

aerobic phases, which was linked to AerAOB activity and a similar increase in pH during

anoxic phases caused by addition of digestate (Wett, 2006). As a lower amount of digestate

was added during the 2nd

cycle, the pH increase took longer and caused an increase of the

anoxic periods during the cycle with 25% compared to the first cycle at high loading. In

addition, the individual aeration phases were 50% longer at lower loading. Although NO and

N2O are mainly emitted during the aeration phases, the decreased total aeration time could not

fully explain the decrease of 79 and 50% for the N2O and NO emission, respectively. NO2


46

emission was less dependent of the aeration regimes and the emission increased with almost a

factor 2 until 0.005% of N load. As NO2 emission is strongly linked to nitrite concentration in

the liquid phase (Weissenbacher et al., 2007), the lower emission of N2O and NO were

probably not caused by lower nitrite fluctuations.

In both cases, a lag phase in the N2O and NO emission compared to CO2 emission was

observed between the start of the aeration in the beginning of the cycle (Fig. 3.3). Therefore,

the question arose whether this occurred because the formation of NO and N2O did not start

from the beginning or because this was just a matter of stripping and the formation a

gas/liquid equilibrium. Because of the higher N2O/NO dynamics at high loading, a detailed

follow-up of the first two hours of this cycle, characterized by highly changing operational

conditions, was performed to answer this question and to try to understand mechanisms

responsible for the emission under transient conditions better (Fig. 3.4).

Chapter 3

47

Figure 3.3: Emission of CO2, N2O, NO and NO2 of a SBR cycle at high (top) and low loading rate

(bottom).


48

Figure 3.4: Top: Concentrations of CO2, N2O, NO and NO2 measured in the defined air stream.

Middle: Intermediate (NO2- and NH2OH), NH4

+ and N2O concentrations in the liquid phase.

Bottom: Dissolved oxygen (DO), NO3- and COD concentration in liquid phase.

Chapter 3

49

During the settling phase (end of previous cycle), oxygen concentrations were depleted and

emissions in the gas phase were limited (Fig. 3.4). However, ammonium consumption (1.2 kg

N d-1

) followed by a NH2OH peak (0.2% of NH4+ consumption) could take place due to the

sudden presence of a limited amount of oxygen (0.2 mg O2 L-1

). The NH2OH peak, together

with a decrease in ammonium and nitrite concentration, was accompanied with an increase in

N2O(l+g) and NO(g) concentration, representing 0.67 kg N2O-N d-1

and 0.001 g NO-N d-1

or 56

and 0.08% of the NH4+ consumptions rate, respectively (Fig. 3.4, Table 3.1). In this phase

several mechanisms could have played a role i.e. nitrifier denitrification, biological or

chemical reaction of NO2- with NH2OH or nitrification-dependent NO and N2O formation

(Chandran et al., 2011). Moreover, a second actor could have been responsible as a COD

removal rate of 3.1 kg COD d-1

was observed at that time, which could indicate that

denitrification could occur at a maximum rate of 1 kg N d-1

. The latter could be another cause

of the N2O and NO emission as the small amount of oxygen present (Fig. 3.4) could probably

inhibit the N2O reductase during heterotrophic denitrification (Otte et al., 1996). Based on the

calculated denitrification rate, this would mean that 67 and 0.1% of denitrified nitrogen ended

up as N2O and NO, respectively. In this context, it should also be mentioned that autotrophic

NO to N2O conversion is not inhibited by oxygen, which is a major departure from known

pathways of heterotrophic denitrification. Taking into account the small increase in nitrate

(0.28 kg N d-1

) and decrease in nitrite (0.08 kg N d-1

), it is plausible that all consumed

ammonium was oxidized to nitrate and further reduced during denitrification. It is hard to

distinguish the mechanisms at this point because of the interplay of several actors (AerAOB,

AnAOB and denitrifiers) during this phase and because the in situ oxygen availability was

unclear.

The sudden pulse of NH4+ together with the start of the aeration at the beginning of the cycle,

resulted in an initial increase of the NO and N2O production. However, the N2O emission in

the gas phase showed a lag phase of about 30 minutes while the lag phase for NO was only

15 minutes and no lag phase was detected for CO2 emission (Fig. 3.4). The difference in lag

phase is on one hand caused by the difference in water solubility and on the other hand a

result of the sequential formation of N2O from NO. A first sharp increase in the ammonium

oxidation rate from 1.2 to 320 kg N d-1

was directly followed by NO emission and N2O

formation of which 13% remained in the liquid phase (Table 3.1, Fig. 3.4). During the

following minutes, nitrite and NH2OH accumulation was observed together with a 1.3 and 5.4

fold increase of the NO and N2O emission, respectively (Table 3.1). In literature, it was


50

described that the imposition of excessive NH3 loads triggers a higher ammonium oxidation

rate, and potentially also a higher amo gene expression (Chandran et al., 2011). The latter

could in turn result in NH2OH accumulation, which is in agreement with our observations.

The same effect was observed when going from anoxic to oxic conditions (Chandran et al.,

2011). AerAOB potentially need to oxidize the accumulated NH2OH to NO in addition to

NO2- to prevent self-inhibition and more effectively derive energy. The latter can both be

done biologically or chemically. Thus, the steep increase in N2O formation during this phase

could mainly be explained by higher specific AerAOB activities resulting in oxidative

formation of NO out of NH2OH.

During the anoxic phase where only feeding was supplied, nitrite consumption occurred while

ammonium concentrations gradually increased. Ammonium consumption rate during this

phase was comparable to the other phases in the cycle (on average 304 kg N d-1

), but

decreased until 290 kg N d-1

at the end of this phase (Table 3.1). A build-up of NH2OH, as a

result of increasing DO concentrations (from 0 to 0.08 mg O2 L-1

) was observed while NO

and N2O(l,g) concentrations were stable around 0.8 and 0.01% of the NH4 oxidation rate,

respectively. From the nitrite consumption, an anoxic ammonium oxidation rate of only

0.31 kg NH4+-N d

-1 was estimated. However, as oxygen still seemed present, aerobic

oxidation of ammonium should have taken place combined with AnAOB activity to explain

the total nitrogen loss of around 300 kg N d-1

.

Table 3.1: Accumulation rate of the different intermediates and anoxic products relatively based on

the total ammonium oxidation rate under different conditions during the SBR cycle. The ammonium

oxidation is a combination of AerAOB and AnAOB activity, which seemed very balanced and stable

over the cycle as no substantial nitrite accumulation took place.

Conditions NH4+ Rv Accumulation rate (% of NH4

+ removal rate)

Feed Aeration DO (kg N d-1

) NO2- NH2OH N2O (l) N2O

(g)

NO Total

No No 0.2 1.2 - 0.2 29.4 26.7 0.08 56

Yes* Yes 0 320 - - 0.1 0.7 0.03 0.8

Yes**

Yes 0 312 0.2 0.001 - 4.3 0.04 4.6

Yes No 0 -0.1 304 - - 0.001 0.8 0.01 0.8

Yes On/off 0 -0.7 308 - 0.0001 - 1.6 0.004 1.6 * First 10 minutes of feeding and aeration;

** Second 10 minutes of feeding and aeration

During the subsequent transient phase with rapidly on/off aeration regimes (on average

6 minutes aeration, 6 minutes without aeration), DO levels sharply increased for the first time

(up to 0.7 mg O2 L-1

). Moreover due to continuous NH4+ feeding, the aerobic oxidation had to

Chapter 3

51

start at higher NH4+ concentrations (170 mg N L

-1 instead of 74 mg N L

-1). From the moment

oxygen was present, aerobic ammonium oxidation could start which resulted in a peak in the

total ammonium oxidation of 335 mg N d-1

, which was even higher than during the start of the

cycle. The consequent NH2OH accumulation and NO and N2O emission could be the result of

both oxidative and reductive formation of NO (Chandran et al., 2011). Both the ammonium

peak and as a consequence specific activity peak and the transition from anoxic to oxic

conditions favor oxidative formation of NO (Chandran et al., 2011), indicating that the

NH2OH availability should have been important. Indeed, as all nitrogen compounds in the

liquid phase remained constant expect for the NH2OH concentration, the latter could be linked

to the increasing NO emission (Fig. 3.4). Due to the low water solubility of NO, the

subsequent increased N2O formation was probably avoided.

As already suggested (Kampschreur et al., 2009b), also in our study AerAOB seem to be the

major contributor to the NO and N2O emission. Sudden pulses of O2 always resulted in

increased NO and N2O formation (Fig. 3.4) and these emissions were accompanied with

NH2OH accumulation. Moreover, peaks in the ammonium oxidation rate (start of the cycle

and start of transition phase, Fig. 3.4) increased the NO and N2O emission. As in both cases

NH2OH accumulation was observed, these measurements indicated a strong link between

NH2OH concentration in aerobic conditions and emissions of these harmful gases. This could

indicate that, under these highly dynamic conditions, nitrifier denitrification played only a

minor role in the system and that the NO and N2O formation mainly followed the NH2OH

route (either chemically or biologically, see Fig. 3.1). This suggests that a better

understanding of the conditions that lead to transient NH2OH accumulation could help to

develop operational strategies to reduce NO and N2O emission from one-stage partial

nitritation/anammox systems.

4 Conclusions

Three main conclusions could be drawn from these measurements:

Highly transient conditions, implying peaks in aerobic ammonium oxidation rates

resulted in increased NO and N2O emissions.

Peaks in NO and N2O emission were always accompanied with NH2OH accumulation.

Therefore, it seemed that biological or chemical production of NO and N2O from


52

NH2OH is the most important cause for the emission during transient one-stage partial

nitritation/anammox.

Operation at more stable conditions and avoidance of NH2OH accumulation could be

key parameters to decrease the NO/N2O emissions.

5 Acknowledgements

H.D.C. is a supported by a PhD grant from the Institute for the Promotion of Innovation by

Science and Technology in Flanders (IWT-Vlaanderen, number SB-81068). The

investigations at the Strass treatment plant were also supported by the Austrian Federal

Ministry of Environment.

54

Sludge settler (WWTP Strass, Austria)

Chapter 4

55

Chapter 4:

OLAND maximizes net energy gain

in technology schemes with anaerobic

digestion

1 Treatment of digestates by OLAND

Autotrophic nitrogen removal processes are the economically preferred method for nitrogen

containing wastewaters low in organic carbon. Landfill leachate, urine and industrial

wastewaters from coke-ovens (Toh and Ashbolt, 2002), tanneries (Abma et al., 2007), semi-

conductor plants (Tokutomi et al., 2011a) and the fertilizer industry (Alberta Environment,

1999) have these characteristics as such, while others obtain an optimal COD/N ratio after

anaerobic digestion. Since this chapter focuses on the impact of OLAND on the energy

balance of systems with energy recovery by anaerobic digestion, the above-mentioned

streams are not covered in this chapter. In what follows, four important combinations of

anaerobic digestion with subsequent OLAND-based nitrogen removal are discussed i.e. the

treatment of (I) the organic fraction of municipal solid waste (OFMSW), (II) manure-based

agricultural waste, (III) starch/sugar-based agro-industrial wastes and (IV) sewage-based

organics. These four application domains in which OLAND minimizes energy use for

digestate treatment, will be discussed in detail in a first section. In a second section, OLAND

application as an active step in minimizing the energy usage of a wastewater treatment plant

with anaerobic digestion as central treatment step is discussed.

Chapter redrafted after: De Clippeleir, H., Vlaeminck, S.E., Courtens, E., Verstraete, W.,

Boon, N., in press. Oxygen-limited autotrophic nitrification/denitrification maximizes net

energy gain in technology schemes with anaerobic digestion Renewable Energy Sources.

Academy Publish, Wyoming, U.S.A.

OLAND maximizes net energy gain in systems with anaerobic digestion

56

BOX 1: General assumptions for energy calculations

The energy values expressed as kWh are considered electrical energy values. If total or

thermal energy is considered the indices ‘tot’ and ‘th’ were used, respectively.

Discharge limits

A discharge limit of 135 mg COD L-1

and 15 mg N L-1

for starch-based agro-industrial and

OFMSW biogas plants, and 5 g COD IE-1

d-1

and 2.5 g N IE-1

d-1

for sewage plants (Siegrist

et al., 2008), was taken into account. For manure-based agricultural waste, a specific case

study from literature was considered (Karakashev et al., 2008), in which 4% of the original

COD and 13% of the original nitrogen present in the digestate was send back to the land.

Energy production through anaerobic digestion

For anaerobic digestion of OFMSW, starch-based agro-industrial waste, source separated

black water and primary/A-stage and secondary sludge, COD removal efficiencies of 87, 90,

80, 60/60 and 35%, respectively, were taken into account (Vaz et al., 2008; Abma et al., 2010;

de Graaff et al., 2010; Hernández Leal et al., 2010). Moreover in every case, a constant biogas

production (0.5 m3 kg

-1 COD removed) and CH4 content of the biogas (65%) was considered.

For the electrical energy recovery, a total energy content 10 kWhtot m-3

CH4 and electrical

recovery efficiency through combined heat and power (CHP) unit of 38% (Wett et al., 2007)

was used. Together these factors lead to an electrical energy recovery of 1.24 kWh kg-1

COD

removed. The composition of the liquid fraction of the digestate was calculated taken into

account an anaerobic yield factor of 0.05 kg COD in biomass kg-1

COD removed and a

COD/N ratio in the sludge of 15 (van Haandel and van der Lubbe, 2007). For each

application, it was considered that nitrogen which was not assimilated, ended up in the

digestate. For COD the assumption differed depending on the application. In sewage-based

systems, COD which was not converted was considered to end up in the solid fraction of the

digestate, while in the other applications, due to a lack of full-scale data, a worst case scenario

was calculated in which all COD that was not converted to biogas or sludge ended up in the

liquid fraction of the digestate.

Energy consumption for the different aerobic/anoxic treatment steps

From the stoichiometry in Table 1.2 (Chapter 1), the oxygen demand for

nitrification/denitrification and OLAND could be calculated and was 4.34 kg O2 kg-1

N

removed and 1.81 kg O2 kg-1

N removed, respectively. Furthermore assuming an oxygen

Chapter 4

57

demand for COD removal of 1 kg O2 kg-1

COD removed, an actual electrical oxygen transfer

efficiency of 1 kg O2 kWh-1

(van Haandel and van der Lubbe, 2007), the resulting energy

requirements for the different treatment schemes were calculated. In case post-denitrification

was considered the energy consumption was calculated based on a volumetric nitrogen

removal rate of 1 kg N m-3

d-1

and an energy demand for mixing of 10 W m-3

reactor. Post-

denitrification could be performed by either adding an external carbon source or by using the

raw waste stream as carbon source, which results in an additional cost or a lower energy

recovery, respectively. The assimilation due to heterotrophic growth was calculated taken into

account a yield factor of 0.5 kg COD in biomass kg-1

COD removed and assuming a COD/N

ratio in the sludge of 15, the sludge production was calculated (van Haandel and van der

Lubbe, 2007). For autotrophic conversions sludge production was neglected.

Additional assumptions for sewage plants

A loading of 135 g COD IE-1

d-1

and 10 g N IE-1

d-1

was used in all sewage treatment schemes

(Verstraete and Vlaeminck, 2011). It was considered that during primary settling 20% of the

COD and 10% of the total N could be separated from the water flow. During enhanced

primary settling the COD removal efficiency could be improved to 40%. For the harvesting of

primary sludge, no additional energy demand was taken into account. For the application of a

highly loaded activated sludge step (A-step) an electrical energy requirement of 0.11 kWh

kg-1

COD removed was considered (Salomé, 1990) and a COD removal efficiency of

conventionally 60 and 50% was considered for centralized and decentralized systems. Due to

a lack of data for the other domains of application, sludge dewatering was only taken into

account for the sewage-based applications at an electrical energy demand of 0.15 kWh kg-1

COD (Zessner et al., 2010).

1.1 Organic fraction of municipal solid waste (OFMSW)

1.1.1 State of the art

The Food and Agriculture Organization (FAO) of the United Nations showed that one third of

all food produced for human consumption, namely 1.3 billion tons wet biomass each year

(84-115 g COD IE-1

d-1

), ends up as municipal solid waste (Monson et al., 2007; Vaz et al.,

2008; FAO et al., 2011). This waste stream consists of kitchen (food) waste (22%), paper and

cardboard (23%) and garden waste (16%) (Burnley, 2007). These organic wastes are

classified as organic fraction of municipal solid waste (OFMSW). If the OFMSW would be

digested anaerobically with a methane yield of 0.3 m³ CH4 kg-1

COD (Monson et al., 2007;


58

Vaz et al., 2008; FAO et al., 2011), an electrical energy production of 101-131 Wh IE-1

d-1

or

1.1-1.3 kWh kg-1

COD is expected. This represents about 2% of the total global electrical

energy utilization in 2008, which amounted to 16 1012

kWh (IEA, 2010). Besides electrical

energy recovery, digesting food waste lowers the amounts of solids send to landfills and thus

reduces greenhouse gas emissions and transportation costs. The high potential of energy

recovery out of those wastes together with the need of new waste management strategies gave

rise to OFMSW specialized biogas producing plants. Energy production is the main objective

of these plants, and one therefore wants to maximize the ‘energy index’, i.e. the ratio of the

produced and consumed electrical energy.

Municipal solid waste generally has a higher risk to contain toxic and inhibitory compounds

than wastewater. These compounds can upon entering the reactor, diffuse quickly in the

diluted slurry and hereby negatively affect the microorganisms (Vandevivere et al., 2002). It

is important to make the distinction between so-called ‘dry’ and ‘wet’ anaerobic digesters

within the OFMSW biogas plants, treating respectively ‘semi-solid’ and ‘liquid’ waste

streams. ‘Dry’ digesters are fed with OFMSW characterized by a low water content (< 15%;

De Baere, 2006), and operate in a semi-solid way resulting in biogas and a solid digestate that

generally is converting into high quality compost or directly applied for agricultural purposes.

The DRANCO, Valorga and Kompogas processes are the most common technologies for this

type of digestion (Six and Debaere, 1992; Wellinger et al., 1993; de Laclos et al., 1997).

Semi-solid systems have become prevalent in Europe, making up 60% of the single-stage

digester capacity installed (De Baere, 2006). On the other hand, ‘wet’ digesters deal with

separately collected food waste with a high moisture content (74-90%, Zhang et al., 2007).

For these streams, the upflow anaerobic sludge blanket (UASB) technology and the

continuous stirred-tank reactors (CSTR) are the most applied technologies. The latter plants

produce next to biogas a solid and a liquid digestate, that can be spread on agricultural land

after pasteurization, in the simplest case. In most cases further treatment of the nitrogen-rich

digestate is required, which has a big influence on the overall electrical energy balance. The

electrical energy demand for pasteurizing is 12%, while 69% is needed to cool the digestate

and subsequently treat by conventional nitrification/denitrification (De Sousa and Vaz, 2009).

The deficit in organic carbon to allow full denitrification implies that less can be digested

(lower energy recovery) or that methanol or another carbon source needs to be added to the

liquid digestate (higher operational costs; Table 1.3 Chapter 1). Until now, full-scale

OLAND-type reactors are not yet applied for the treatment of liquid digestates from OFMSW

Chapter 4

59

biogas plants. However, successful OLAND treatment has been demonstrated on digestates

from co-digestion plants with food waste in centralized (Wett et al., 2007) or decentralized

(Zeeman et al., 2008) sewage plants, suggesting the feasibility of OLAND implementation in

OFMSW plants.

1.1.2 Implication of OLAND application on the energy balance

The main goal of OFMSW biogas plants is to produce energy from biodegradable municipal

solid waste. As mentioned above, nitrogen treatment will only have an impact on the total

energy balance of ‘wet’ digesters, which produce a liquid digestate that cannot be discharged

as such.

BOX 2: Energy calculation from an OFMSW biogas plant

Assuming a COD loading rate to the OFMSW digestor of 81 g COD IE-1

d-1

and a

digestibility of 87% (Vaz et al 2008), a biogas production of 0.02 m3 CH4 IE

-1 d

-1 and

electricity gain of 87 Wh IE-1

d-1

or 1 kWh kg-1

COD can be achieved. The digestate still

contains 7 g COD IE-1

d-1

(maximum value) and 6 g N IE-1

d-1

, which should be treated

biologically by OLAND/post-denitrification or by conventional nitrification/denitrification.

Discharge limits are assumed at 0.0025 g N IE-1

d-1

and 0.005 g COD IE-1

d-1

, and the volume

at 0.2 Lsolid municipal waste IE-1

d-1

. For the OLAND scenario, 11.7 Wh IE-1

d-1

is needed for

ammonium oxidation. For aerobic COD degradation, an electrical energy demand of

4.9 Wh IE-1

d-1

is expected. Because OLAND oxidizes 11% of the ammonium to nitrate,

0.1 Wh IE-1

d-1

is needed for mixing in the post-denitrification reactor to meet discharge

limits. Hence, the electricity consumption of the OLAND scenario amounts to 17 Wh IE-1

d-1

.

Using conventional nitrification/denitrification, addition of an external carbon source is

needed or raw waste should be used for denitrification (lower biogas recovery). In the latter

case, the electrical energy recovery will decrease to 67 Wh IE-1

d-1

and a total energy

consumption of 22 Wh IE-1

d-1

is needed to fully polish the digestate.

As organic waste is highly digestible (efficiency around 80-90%, BOX 1), in general an

electrical energy recovery of 1.1 kWh kg-1

CODin can be expected. The liquid digestate is rich

in nitrogen and devoid in organic carbon (4800-9700 mg COD L-1

, 1119-1500 mg N L-1

,

COD/N 4-11). For further digestate polishing by OLAND with post-denitrification and by

conventional nitrification/denitrification, an electricity consumption of 17 and 22 Wh IE-1

d-1

or 0.19 and 0.25 kWh kg-1

COD, respectively, was estimated (BOX 2). It can be concluded


60

that by applying OLAND for this application a net electrical energy gain of 0.8 kWh kg-1

CODin compared to 0.5 kWh kg-1

CODin for conventional treatment can be obtained.

1.2 Manure-based agricultural waste


Animal waste streams are characterized by a high nitrogen and organic carbon content.

Although the latter implies a high energy potential, particularly for pig manure the relatively

high free ammonia concentrations can significantly disturb biogas formation (Chen et al.,

2008). To avoid this, an acidifying digestion (pH 6-7; lower free ammonia levels) is

sometimes performed, converting the more complex organics into volatile fatty acids (Chen et

al., 2008). Biogas production is limited in this step. By separation of the solid and liquid

fraction afterwards, the volatile fatty acids in the liquid fraction can be converted to biogas in

an anaerobic digester (e.g. UASB). In this step around 57% of the present COD can be

converted to biogas in the presence of about 4 g NH4+-N L

-1 (Karakashev et al., 2008). Given

the high organic carbon content of pig manure (around 70 g COD L-1

), the digestate still

contains a considerable amount of soluble organics (10 g COD L-1

) (Karakashev et al., 2008).

The latter levels and some specific compounds might be inhibitory for AnAOB (Dapena-

Mora et al., 2007) and their removal will enhance the success of OLAND treatment. A

separate oxidation step can decrease COD levels to about 3.5 g COD L-1

and COD/N ratios to

about 2. Given the choice of a separate oxidation, most studies on autotrophic nitrogen

removal on digested manure focused on two separate nitrogen removal steps (partial

nitritation – anammox) (Van Hulle et al., 2010), incorporating the COD oxidation in the

partial nitritation stage. Removal efficiencies obtained with AnAOB-based technologies for

digested manure are generally around 70% (Van Hulle et al., 2010), and thus lower than for

less complex digestates such as sludge reject water (Table 1.4, Chapter 1). Research showed

that the removal efficiency of AnAOB-based processes was not only dependent on the

COD/N ratio obtained after digestion or post-digestion, but also on the absolute COD levels.

COD concentrations above 142 and 242 mg COD L-1

for instance ceased the AnAOB activity

for post-digested manure and partially oxidized digestate, respectively (Molinuevo et al.,

2009). These values should however not be generalized, and are likely dependent on the test

conditions and acclimatization.

Treatment schemes with autotrophic nitrogen removal in a one-stage (Karakashev et al.,

2008) or two-stage (Hwang et al., 2006) setting have thus far only been tested in batch or in

Chapter 4

61

continuous lab-scale reactors. The lack of pilot- and full-scale investigations does not yet

allow to discuss their environmental and economical sustainability (Karakashev et al., 2008).

Due to the stringent environmental regulations concerning the application of the manure

nutrients as direct fertilizer on agricultural land, treatment of digestates is more and more

becoming a necessity. At this moment, the amount of added nitrogen may not exceed

170 kg N ha-1

yr-1

(Oenema, 2004), and this amount is even likely to be decreased in the

future. This should provide a stimulus to validate schemes with autotrophic nitrogen removal

on a larger scale.


Because of the relatively low biodegradability of manure COD and the relatively high

nitrogen content, electrical energy recovery through anaerobic digestion is more difficult than

for other streams. For pig manure for instance, only 29 kWh m-3

raw manure or 0.4 kWh kg-1

CODmanure can be recovered as electricity (BOX 3). The implementation of OLAND in the

treatment scheme requires an electrical input of 17.6 kWh m-3

raw pig manure or 0.25 kWh

kg-1

CODmanure (BOX 3), taking into account the energy need for solid/liquid separation, COD

oxidation and OLAND treatment. The electrical energy demand for conventional

nitrification/denitrification is somewhat higher, i.e. 0.27 kWh kg-1

CODmanure (BOX 3).

Therefore, the net electrical energy gains are 11.4 and 10.2 kWh m-3

raw pig manure, or 0.16

and 0.15 kWh kg-1

CODmanure for the treatment with OLAND and nitrification/denitrification,

respectively (BOX 3). It should be mentioned that external carbon will be required for

nitrification/denitrification, given the relatively low BOD/COD ratio of manure (Lemmens et

al., 2007). So, although additional energy gain by implementation of OLAND seems minor

(BOX 3), the cost for an external organic carbon source can be avoided in this way.

Direct biological treatment of the liquid manure fraction is often applied to avoid the cost of

an external carbon source. This approach does not recover energy and has an average net

energy consumption between 16 and 22 kWh m-3

raw manure (Lemmens et al., 2007), which

is 26 and 33 kWh m-3

raw manure higher than the electricity consumption of the schemes with

anaerobic digestion.


62

BOX 3: Energy calculation of a manure-based agricultural plant

As an example, a treatment scheme was chosen for piggery manure with thermophilic

anaerobic digestion followed by a centrifugation step. The liquid fraction (~90 vol% of the

raw manure) is subsequently treated in a UASB reactor to produce biogas. The digestate is

then treated in a separate COD oxidation step, followed by an OLAND reactor. The mass

balances of COD and N are shown in Fig. 4.1. If conventional nitrification/denitrification

would be applied, the COD oxidation step can be left out of the scheme.

From the proposed treatment scheme 25 kg COD m-3

raw manure can be recovered as biogas,

hence the electrical recovery is 29 kWh m-3

raw manure. A solid/liquid separation step, often

using decanter centrifugation, consumes on average 4 kWh m-3

raw manure (Lemmens et al.,

2007). An electricity consumption of 11.8 kWh m-3

raw manure was calculated to convert

COD to CO2 and sludge in the oxidation stage. It was assumed that nitrogen losses were

minimal, and mainly caused by volatilization of ammonia and nitrogen assimilation. The

subsequent OLAND reactor has an electrical energy demand of 1.8 kWh m-3

raw manure. The

total electrical energy requirement for this system is therefore 17.7 kWh m-3

raw manure.

When conventional nitrification/denitrification is applied, the total electrical energy demand

only slightly increases to 18.8 kWh m-3

raw manure due to the absence of the aerobic COD

degradation. However, external organic carbon source addition will be needed in this scheme,

increasing the operational costs. Addition of raw manure as carbon source for denitrification

seems not a good option due to the low biodegradability and high nitrogen content.

Figure 4.1: Possible innovative treatment scheme for pig manure, with COD and N quantities

expressed for 1 m3 of manure (Karakashev et al., 2008).

Chapter 4

63

1.3 Sugar/starch-based agro-industrial waste


Several full-scale OLAND-type processes are used to treat digestates in food industry, yeast

factories and distilleries (Table 1.4, Chapter 1). For example wastewater from potato

processing can be treated by anaerobic digestion to recover energy from the present organics

(57 kg COD m-3

), followed by struvite precipitation to recover the phosphorous and OLAND

treatment to remove the residual nitrogen (0.7 kg N m-3

digestate) (Abma et al., 2010). Both

one-step and two-step autotrophic nitrogen removal processes for potato processing plants are

operational at full-scale (Abma et al., 2010; Desloover et al., 2011a). Another example can be

found in Asian food culture, where monosodium glutamate is a popular flavor. It is produced

by fermentation of rice, starch and molasses, which finally creates a wastewater rich in

suspended solids (SS; 200-1000 mg SS L-1

), COD (1500-60000 mg L-1

), NH4+ (200-15000

mg N L-1

) and sulfate (3000-70000 mg L-1

) (Zhang et al., 2008). This wastewater is

traditionally treated by physico-chemical and biological methods decreasing the wastewater

content to around 200-270 mg SS L-1

, 1000-1400 mg COD L-1

and 150-350 mg N L-1

(Zhang

et al., 2008). In case of a low biodegradability of the residual COD, denitrifier overgrowth of

AnAOB can be avoided, making the OLAND process feasible. Nitrogen removal efficiencies

above 80% were obtained at full-scale OLAND plants treating effluent from monosodium

glutamate wastewater in China (Table 1.4, Chapter 1).

In general, anaerobic digestion of carbon-rich, digestible industrial wastestreams can

significantly decrease the BOD/N ratio below 3-4. The digestate then comes into the scope of

subsequent OLAND treatment.


Most industrial full-scale OLAND applications are situated in the food industry (Table 1.4,

Chapter 1). As wastewater from several food processing companies is directly digestible, the

electrical energy recovery from this waste stream is expected to be around 1.1 kWh kg-1

CODin. Nitrogen removal through OLAND and post-denitrification has an electricity demand

of 0.11 kWh kg-1

CODin. This leads to an overall net electrical energy gain of 1.0 kWh kg-1

CODin (BOX 4). In the other scenario, with nitrification/denitrification and bypassing some

organic carbon source to the denitrification reactor, the net electrical energy gain is reduced

with 27% to 0.73 kWh kg-1

CODin (BOX 4). OLAND implementation on digestible industrial


64

wastewater can hence decrease the electricity consumption with a factor 2, having a

significant impact on the overall energy balance.

BOX 4: Energy calculation of a starch-based agro-industrial plant

A potato factory producing 17000 kg COD d-1

and 1000 kg N d-1

(at 3000 m3 d

-1) obtained a

COD removal efficiency during anaerobic digestion of 90% (Abma et al., 2010). The

electrical energy recovery from biogas production is 19000 kWh d-1

or 1.1 kWh kg-1

CODin.

In the worst-case assumption, the COD that is not converted to biogas ends up in the

digestate, i.e. 316 mg COD L-1

, next to 316 mg N L-1

. The latter should be treated biologically

by OLAND/post-denitrification or by nitrification/denitrification. Discharge limits of

15 mg N L-1

and 135 mg COD L-1

were used. For the scheme with OLAND, electrical energy

demands of 1700 kWh d-1

, 142 kWh d-1

and 11 kWh d-1

are needed for the OLAND reactor,

aerobic COD removal and final denitrification, respectively. Thus, the total electrical energy

need for digestate amounts to 1900 kWh d-1

or 0.11 kWh kg-1

CODin. In contrast to this

scenario, a part of the raw waste can bypass the digestor and serve as carbon source for

denitrification. In this case the electrical energy recovery decreases to 16000 kWh d-1

(0.9 kWh kg-1

COD) and the electrical energy consumption increases to 3500 kWh d-1

(0.20 kWh kg-1

COD). The overall electricity gain of this scenario is hence 0.73 kWh kg-1

COD.

1.4 Sewage-based organics

1.4.1 State of the art: centralized treatment

As the main treatment step during sewage treatment is based on heterotrophic, aerobic

conversions, sludge production during conventional activated sludge (CAS) treatment is about

0.5 kg COD converted to sludge biomass per kg COD converted. The daily specific sludge

production varies between 40 and 60 g DM IE-1

d-1

, with the lowest values for CAS systems

with nitrogen treatment and the highest production for CAS systems with additional P

removal (Zessner et al., 2010). Sludge treatment by anaerobic digestion in combination with

land application is the most sustainable approach due to the low emission and low energy

consumption (Suh and Rousseaux, 2002). The digestate formed as a result of sludge digestion,

normally only accounts for 1% of the influent water flow but for 15-20% of the nitrogen load

of the CAS system. Therefore, it should be further treated before discharge (Fux and Siegrist,

2004).

Chapter 4

65

Figure 4.2: Sewage treatment schemes based on CAS (A, B) and the A/B process (C, D) with and without OLAND implementation in the side stream.


66

Digestates from municipal sludge digestion are characterized by a high nitrogen content

(0.4-1 g N L-1

) while a high proportion of the biodegradable COD is already digested,

obtaining in most cases biodegradable COD (bCOD)/N ratios of 1-2. Due to the optimal

composition of the sludge reject water, and provided the presence of qualified operators at

CAS systems and the possibility to treat the OLAND effluent in the existing treatment

system, around 70% of the full-scale OLAND-type reactors are applied in this area. Long

start-up periods (order of years) were needed for the first full-scale applications (van der Star

et al., 2007). At present, due to the possibility to acquire active biomass from other

installations, start-up periods are ranging from 1-2 months for suspended growth systems

(personal communication, Bernhard Wett) and 3-6 months for systems based on attached

growth (Christensson et al., 2011). Since similar total nitrogen removal efficiencies are

obtained in the different reactor types (Table 1.4, Chapter 1), the choice of technology mainly

depends on the footprint area availability and the importance of energy efficiency. Moving

bed bioreactors (MBBR) consume almost 5 times more electrical energy to remove the same

amount of nitrogen (5.63 compared to 1.13 kWh kg-1

N) and moreover require lower

volumetric loading rates to obtain optimal performance (Wett et al., 2007; Christensson et al.,

2011). In most cases however, the reactor choice is connected with the constructor choice for

implementation of a full-scale OLAND reactor, as every constructor is known for his

operation methodology and design parameters.

As the implementation of anaerobic digestion in municipal WWTP is still growing, with

Sweden as one of the leading countries, digesting already 83% of the sludge (Lantz et al.,

2007), the OLAND application in this area is immerging. It is slowly becoming a standard

treatment method in municipal WWTP as high and stable nitrogen removal performance has

been demonstrated and a positive influence on the energy balance of the WWTP has been

established.

1.4.2 State of the art: decentralized treatment

Since 70-80% of the costs of municipal sewage management derive from the sewerage system

(Bieker et al., 2010), and since additionally during the long transport around 30% of the

dissolved COD (potential energy) in wastewater is lost (Huisman et al., 2004), decentralized

concepts are suggested, aiming at a maximum recovery of energy and nutrients. Source

separation can prevent pollutant dilution, and hence renders recovery feasible. At the

household level, three main streams can be separated (Table 4.1). Urine or yellow water

Chapter 4

67

contain most of the soluble nutrients and due to the high concentration, recovery of N and P is

possible (Otterpohl, 2002). From brown water (faeces), energy can efficiently be recovered

due to the concentrated organic carbon content by anaerobic digestion, which simultaneously

allows hygienisation towards an organic soil improver. In some cases, urine and faeces are

collected together as black water which can be even concentrated by the use of vacuum toilets

(Zeeman et al., 2008). Grey water, a less concentrated stream from showers, bath, washing

machines, kitchens etc can be treated with small efforts as its temperature is elevated and the

nutrient content is low enough to require nutrient elimination to reach service water quality

(Cornel et al., 2011). The grey water can for example by recycled as toilet flushing water,

saving 30% of the potable water consumption (Cornel et al., 2011).

Table 4.1: Distribution of the daily COD and nitrogen loads in different wastewater streams

(Henze, 1997; Otterpohl, 2002).

Sewage Grey water Brown water

(faeces)

Yellow water

(urine)

Volume

(L IE-1

d-1

)

25-100 25-100 0.05 0.5

Compound (g IE-1

d-1

) % % %

COD 135 41 47 12

N 10 3-8 8-10 77-87

P 2.5 10-20 20-40 50-60

At present, few source-separated schemes incorporate a complete treatment scheme in

operation. A semi-centralized approach (20 000 IE) including separate grey and black water

treatment and biogas production from bio-waste was applied in Qingdao (China). Due to the

incorporation of bio-waste digestion, this concept could provide the electric energy demand

within this semi-centralized treatment process (Cornel et al., 2011). Another pioneering

project in Sneek (The Netherlands), where grey water and concentrated black water are

collected from 32 houses, has shown to be feasible and profitable (Zeeman et al., 2008;

Verstraete and Vlaeminck, 2011). In the latter concept concentrated black water vacuum

collection consumes in terms of electricity 10-27 Wh IE-1

d-1

, but it uses 7 times less water

(WRS, 2001), thus allowing to treat a very concentrated black water stream that is directly

suitable for anaerobic digestion. Co-digestion of bio-waste and sludge from the grey water

treatment (A/B system) can further increase the biogas output. However, due to the small

scale of this project so far, electrical power production from biogas formation is not feasible

and thus only heat is used by applying a co-combustion, which switches between biogas and

natural gas. The digestate, containing 90% of the nitrogen load of a household is treated by an


68

energy-friendly OLAND RBC, decreasing the overall energy need of the plant. Compared to

the centralized approach where around 15-20% of the nitrogen load ends up in the digestate,

the importance of nitrogen removal technology choice increases significantly in the

decentralized approach. Therefore, OLAND technology can drastically change the energy

consumption and operational costs in the latter case. Due to the concentration of the nutrients

in black water, recovery of phosphorous and part of the nitrogen can also be achieved by

struvite precipitation.


Current sewage treatment systems are mainly designed to remove organics from wastewater

although the latter can be regarded as a source of energy. Nitrogen removal in the

conventional activated sludge (CAS) system requires a lot of electrical energy to obtain full

nitrification (Table 1.3, Chapter 1, Fig. 4.2). Moreover, it uses organic carbon to denitrify

nitrate to nitrogen gas. Aeration constitutes over 60-70% of the electrical energy consumption

of CAS systems with and without anaerobic digestion, respectively (Zessner et al., 2010). In

an average CAS system the total electrical energy consumption is around 96 Wh IE-1

d-1

,

which can be covered for 44% by electrical energy recovery through anaerobic digestion of

the primary and secondary sludge (Table 4.2, Fig. 4.2). Application of a two-stage activated

sludge system or A/B Verfahren system, can increase the role of electrical energy recovery by

anaerobic digestion to 90% of the electrical energy consumption (Table 4.2). Implementation

of OLAND for digestate treatment in the side line of the CAS and A/B Verfahren system,

without other changes, can not increase the total electrical energy gain (Table 4.2). The main

reason for this is that the decrease in electrical energy requirement for nitrogen removal can

not counteract the increase in aerobic COD removal due to the lower denitrification rate in the

main line. Therefore on first side, the implementation of the OLAND process does not seem

to have a significant effect on the total energy balance. However, OLAND in the side line

allows higher organic carbon removal through sludge by enhanced primary settling (e.g. by

addition of flocculants) or improved highly loaded activated sludge treatment (A-step of A/B

system), because less organic carbon is needed for final nitrogen polishing through

denitrification and can thus be recovered as biogas. OLAND implementation for digestate

treatment in CAS systems can lower the overall plant energy requirements with about 50%

(Table 4.2; Siegrist et al., 2008) due to a higher electrical energy recovery and a decrease in

aerobic COD degradation. Furthermore, according to the theoretical calculations (Table 4.2)

and the in practice experiments of Wett et al. (2007), energy autarky by including OLAND in

Chapter 4

69

the sidestream of a two-stage activated-sludge (AS) process (‘A/B Verfahren’) can be

obtained if the A-step efficiency is high enough. In general, OLAND implementation in the

side line of centralized WWTP can allow a higher net electrical energy gain if at the same

time a higher organic carbon recovery through biogas production is applied to make profit

from the lower organic carbon requirements for denitrification.

Table 4.2: Energy demand and gain of municipal WWTP schemes. CAS: conventional activated

sludge treatment; OL

: OLAND reactor in side line, treating sludge reject water; ° enhanced primary

settling applied; AX/B: A/B system with a COD removal efficiency of X in the A-stage; OL: OLAND

in the main stream; *corrected for COD removal through denitrification

Oxygen and energy demand Electrical energy gain (Wh IE-1

d-1

)

Case 1 Case 2 Case 3 Case 3 Case 4 Case 5

CAS CASOL

CASOL°

A60/B A60/BOL

A75/BOL

COD removal* -37.6 -41.9 -25.9 -9.7 -15.6 -3.6

N removal main line -32.0 -24.7 -28.6 -33.2 -23.2 -26.2

OLAND in side line / -3.1 -2.4 / -4.2 -3.7

Energy consumption A step / / / -8.9 -8.9 -11.1

Pumping/mixing -20.0 -20.0 -20.0 -20.0 -20.0 -20.0

Sludge dewatering -6.1 -6.1 -6.2 -6.4 -6.4 -6.5

Total energy consumption -95.7 -95.8 -83.1 -78.3 -78.3 -71.1

Biogas-based energy production +42.3 +42.3 +56.4 +70.6 +70.6 +81.2

Net energy gain -53.4 -53.5 -26.7 -7.7 -7.7 +10.1

BOX 5: Sensitivity analysis of energy calculations

The assumptions made regarding the energy calculations of the centralized wastewater

treatment schemes, have a high impact on the obtained energy gain. The following parameters

showed the highest impact on the total energy balance of the systems:

The actual oxygen transfer efficiency

The actual oxygen transfer efficiency is dependent on the type of aerator (surface, diffusers),

the reactor design and the wastewater and operational properties. A high oxygen transfer

efficiency can drastically decrease the total electrical energy consumption (Fig. 4.3).

Moreover, in A60/B systems at oxygen transfer efficiencies above 1.8 kg O2 kWh-1

a net

energy production independent of the digestibility of primary sludge can be obtained. In CAS

systems it is clear that a net electrical energy gain is hard to obtain without increasing the

primary sludge production.

The digestibility of primary sludge

Digestion efficiencies of mixtures of primary and secondary sludge are reported around 50%.

However, a difference in digestion efficiencies between both types of sludge exists. In the


70

main calculations an efficiency of 60 and 35% was considered for primary and secondary

sludge, respectively. Depending on the type of wastewater treated and the method of primary

sludge production (settler, enhanced settler, A-stage), the digestibility of primary sludge can

differ. Applying this deviation to the calculation, it could be shown that the net energy gain

can be significantly influenced by this parameter (Fig. 4.3). For A60/B systems this influence

starts earlier and is higher because of the higher proportion of primary sludge production

compared to CAS systems.

Figure 4.3: Net electrical energy gain (Wh IE-1

d-1

) of CAS system (top) and A60/B system (bottom)

in function of the actual oxygen transfer efficiency and primary sludge digestibility.

Chapter 4

71

Energy for pumping

In all schemes the electrical pumping energy was considered constant (20 Wh IE-1

d-1

) and

represented around 22-31% of the total electrical energy consumption. Depending on the

treatment scheme, the amount of nitrogen that should be denitrified differs. The latter is

mainly steered by applying a certain recirculation ratio in the activated sludge system in the

main line, which means that a lower denitrification rate implies a lower recirculation ratio and

thus a lower pumping cost. A lower electrical pumping energy due to the absence of

recirculation with for example 25% (to 15 Wh IE-1

d-1

; Kartal et al., 2010a), can decrease the

total energy consumption with 5-8%.

Aerobic yield of heterotrophs

The aerobic COD yield of heterotrophs in the activated sludge system was considered 0.5 kg

COD in cellular microbial biomass kg-1

COD removed in the calculation. A deviation of this

factor with 20% to 0.6 or with 40% to 0.7 kg COD in biomass kg-1

COD removed could

increase the net electrical energy gain in a CAS system with 15 and 65%, respectively. The

latter indicated that for a CAS system and A60/B system an increase in the yield to 0.8 and

0.52 kg COD in biomass kg-1

COD removed, respectively, could result in an energy self-

sufficient system in terms of electrical energy.

On decentralized level, a size of 50 000 to 100 000 IE is recommended to obtain electrical

energy recovery from biogas (Bieker et al., 2010). Current research gives reasons to believe

that investment costs and income from energy are going to balance after about 15 to 20 years

(so far integrated decentralized systems may be more expensive in investments) while the

operation costs of these systems seem to be only a fraction of the costs of centralized systems.

The main reason for the latter is the energetic use of solid waste and sewage sludge within the

decentralized approach, while this is more difficult in the conventional centralized approach

due to the high dilution (Bieker et al., 2010). Since more than 75% of the nitrogen load is

present in the liquid fraction of digested concentrated black water, which was separated from

grey water (Table 4.1), the application of OLAND vs. nitrification/denitrification has a very

high impact. The implementation of OLAND in source-separated systems with black water

(from vacuum toilets) and grey water, makes the crucial difference between energy negative

and energy-positive treatment (Table 4.3). Decentralized schemes based on source separation

with nitrification/denitrification consume 43 and 24 Wh IE-1

d-1

less than the classical

centralized system, when digestion of only black water or also additional sludge of the A/B

treatment of the grey water line and of the nitrogen treatment in the black water line is


72

considered, respectively. However, both schemes with nitrification/denitrification are energy

negative. The implementation of OLAND makes both schemes energy positive (Table 4.3),

due to a lower electrical energy consumption of 86 and 125 Wh IE-1

d-1

for both schemes

respectively, compared to CAS systems. Therefore, an electrical energy saving with a factor

2-5, depending on the digestion options, can be established by implementation of OLAND in

source separated treatment schemes.

Table 4.3: Electricity consumption (Wh IE-1

d-1

) and energy (E) index (-) comparing different options

for decentral sewage treatment schemes with the central, conventional activated sludge (CAS). For

‘decentral 1’, only black water is digested, whereas also sludge from A and B stages (grey water line)

and nitrogen removal (black water line, through OLAND or nitrification/denitrification; N/DN) is

digested for ‘Decentral 2’. Calculation assumptions are mentioned in Box 1. Organic carbon for

denitrification was provided by raw black water.

CAS Decentral 1:

AD black water

Decentral 2:

AD black water+all sludge

OLAND N/DN OLAND N/DN

COD

removal*

-37.5 -15.9 -15.9 -15.9 -15.9

Biogas 42.3 78.7 56.5 112.1 34.6

N removal -32.0 -13.1 -32.8 -11.2 -34.1

Pumping -20.0 -20.0 -20.0 -20.0 -20.0

Dewatering -6.4 -9.1 -10.2 -5.3 -6.1

Water* +12 +12 +12 +12

Total gain -53.4 32.3 -10.7 71.5 -29.6

E - index 0.4 1.6 0.7 2.4 0.6

* Incorporating savings aerobic COD oxidation in case of mainstream CAS treatment * The drinking water production of 25 L IE

-1 d

-1 is avoided, requiring 0.47 Wh L

-1 (Verwin, 2006)

1.5 Treatment of digestates by OLAND: conclusions and perspectives

As OLAND decreases the energy need for nitrogen removal up to a factor 2, OLAND has the

potential to decrease the overall electrical energy consumption significantly. However, during

OLAND treatment, 11% of the converted nitrogen ends up in the effluent as nitrate, which

especially for high-strength wastewaters such as digestates needs further polishing before

discharge is permitted. The significance of OLAND implementation on the net electrical

energy gain of a treatment system depends firstly on the proportion of the nitrogen load send

to the OLAND reactor and secondly on the composition of the digestate itself (Table 4.4).

Due to the suboptimal composition of liquid manure digestates, OLAND implementation has

no strong effect on the electrical energy balance compared to conventional treatment.

However, OLAND can offer a cost-effective treatment method because the cost of an external

organic carbon source is avoided. In OFMSW plants and sugar/starch-based agro-industrial

Chapter 4

73

treatment plants, OLAND application could significantly enhance electrical energy recovery

from waste by minimizing the energy consumption for digestate treatment mainly because

around 95% of the nitrogen load is treated by OLAND and optimal BOD/N ratios are

obtained in the digestate. Moreover, also in the latter applications, the cost for external

organic carbon source addition is avoided in contrast to conventional treatment. In municipal

treatment plants the effect of OLAND implementation in the side line, without other

adjustments, is negligible because only 15-20% of the nitrogen load is treated through

OLAND and the decrease in electrical energy demand for nitrogen removal can not counteract

the increased demand for aerobic COD removal (Table 4.4). Despite the low nitrogen load

treated by OLAND in municipal WWTP, energy autarky is possible through the

implementation of enhanced primary sludge production and thus increased electrical energy

recovery (Table 4.4, Wett et al., 2007; Siegrist et al., 2008). At decentralized level, due to

source separation, OLAND can make the difference between energy-positive and energy-

negative operation.

Table 4.4: Comparison of energy index (energy production/energy consumption) of the treatment of

anaerobic digestion digestates with the conventional treatment through nitrification/denitrification

(N/DN) and the alternative treatment through OLAND.

Application domain Energy index

N/DN

Energy index

OLAND

OFMSW plant* 3 6

Manure-based agricultural plant* 1.6 1.7

Starch-based agro-industrial plant* 5 10

Sewage with CAS 0.4 0.4/0.7°

Sewage with A60/B 0.9 0.9/1.1°°

Decentral 1 (AD black water) 0.9 1.6

Decentral 2 (AD black water + additional sludge) 0.6 2.4

*Pumping energy is considered negligible; °With enhanced primary settling; °° With an improved

A-step (75%)

2 OLAND as mainstream treatment process

OLAND implementation combined with improved primary sludge production in municipal

WWTP allows energy autarky as discussed in the previous section. However, if a higher

proportion of the nitrogen present in sewage can be send to the OLAND system, for example

by implementation of OLAND in the main line of the municipal WWTP, a higher decrease of

the electrical energy consumption should be possible. Moreover, more organics could be

recovered as electrical energy since no additional organic carbon is needed to meet nitrogen


74

discharge limits. The consequences of applying OLAND as a main treatment step in

municipal WWTP are discussed in this section (Fig. 4.4).

Figure 4.4: Schematic overview of implementation of OLAND in mainline of WWTP (A/OL).

2.1 Wastewater as an energy resource

The potential energy in the form of organics available in the raw sewage exceeds the

electricity requirements of the treatment process. Based on calorimetric measurements a

specific energy input of 14.7 kJ per g COD can be calculated (Shizas and Bagley, 2004). In

the conventional activated sludge (CAS) system, 38% of the incoming COD is aerobically

converted to CO2 or anaerobically used for denitrification (Table 4.2, BOX 1). So, this means

that 754 kJ IE-1

d-1

is spilled through metabolic reactions and not recovered as electrical

energy resource. As a consequence the CAS systems can only produce the electrical

equivalent of 42 Wh IE-1

d-1

which can not cover the electrical total energy costs 96 Wh IE-1

d-1

(Table 4.2). To fully recover the potential energy in the raw sewage, not only electrical

energy usage minimization should be accomplished but more important, the electrical energy

recovery by anaerobic digestion should be maximized. Therefore, the CAS system should be

replaced by a first concentration step, bringing as much as COD as possible to the solid

fraction, and a second biological step removing the residual nitrogen and COD with a

minimal energy demand (Fig. 4.4).

A highly loaded activated sludge (A-step) compartment in the mainline can act as a

concentration step. This A-step works at low hydraulic (15-30 minutes) and sludge retention

Chapter 4

75

(0.5 d) times, allowing the bacteria to work at their maximum growth yield. Therefore,

organics are only incorporated in the biomass resulting in COD removal efficiencies to the

sludge phase of 60-70% (Salomé, 1990). Higher efficiencies up to 80% can be accomplished

by an increased loading rate, addition of flocculants, iron etc (Xia et al., 2005), allowing even

higher biogas production rates. In Austria, an A-step is operated at COD removal efficiencies

of 60% allowing conventional nitrification/denitrification in the main stream provided that the

nitrogen in the side stream is separately treated in an OLAND step (Wett et al., 2007). This

concept as such allows almost electrical energy autarky and can lead with co-digestion of

kitchen waste to electrical energy neutral operation (Wett et al., 2007). An overall net

electrical energy gain can be accomplished by applying a more efficient A-step (75% COD

removal instead of 60%) (Table 4.5) allowing higher energy recovery through anaerobic

digestion. Taking into account an average domestic wastewater composition of 30-100 mg N

L-1

and 450 – 1200 mg COD L-1

(Tchobanoglous et al., 2003; Henze et al., 2008), rendering

COD/N ratios between 12 and 15, COD separation efficiencies in the A-step of 75-80% will

result in COD/N ratios which are too low to allow full nitrification/denitrification. Therefore

to obtain the same overall removal efficiencies in the WWTP, an external organic carbon

source should be added to allow conventional nitrification/denitrification. This additional cost

for an organic carbon source counteracts the advantages of the higher net electrical energy

gain and therefore, in this case, the OLAND process could offer a cost-effective and energy

friendly alternative. Theoretically, the implementation of OLAND in the main line at an A-

step COD efficiency of 75%, does not further increase the net electrical energy gain (Table

4.5). Applying OLAND in the main line implies that the remaining COD should be removed

aerobically and cannot be totally removed by denitrification, increasing the electrical energy

need for COD removal with a factor 4. Therefore the decrease of the electrical energy

requirement for nitrogen removal with 35%, can not counteract the increase in electrical

energy demand for COD removal (Table 4.5). The main advantage of the implementation of

OLAND in the main line compared to conventional nitrification/denitrification is the cost

savings for external organic carbon addition. Since the improved CAS system with enhanced

primary settling and OLAND in the side line requires an electrical input of 27 Wh IE-1

d-1

and

this new concept can potentially gain an electrical equivalent of 10 Wh IE-1

d-1

(Table 4.5),

the OLAND application in the main stream should be a further step towards a more energy

friendly wastewater treatment. In the following sections the challenges (as indicated in

Chapter 1) to accomplish this concept are studied in detail (Chapter 6-8).


76

Table 4.5: Electrical energy demand and gain of municipal WWTP schemes. CAS: conventional

activated sludge treatment; OL

: OLAND reactor in side line, treating sludge reject water; ° enhanced

primary settling applied; AX/B: A/B system with a COD removal efficiency of X in the A-stage;

OL: OLAND in the main stream; *corrected for COD removal through denitrification

Oxygen and energy demand Electrical energy gain (Wh IE-1

d-1

)

CASOL°

A60/B A75/BOL

A75/OL

COD removal* -25.9 -9.7 -3.6 -14.4

N removal main line -28.6 -33.2 -26.2 -19.5

OLAND in side line -2.4 / -3.7 /

Energy consumption A step / -8.9 -11.1 -11.1

Pumping/mixing -20.0 -20.0 -20.0 -20.0

Sludge dewatering -6.2 -6.4 -6.5 -6.5

Total energy consumption -83.1 -78.3 -71.1 -71.5

Biogas-based energy production +56.4 +70.6 +81.2 +81.2

Net energy gain -26.7 -7.7 +10.1 +9.7

2.2 Main stream OLAND application: conclusions

The implementation of OLAND in the main line of the WWTP shows high potential to

further increase the energy gain from sewage. A net electrical energy gain of 10 Wh IE-1

d-1

is

expected to be possible. However, the challenges of microbial biomass retention at low HRT

(< 1d) and NOB suppression at low temperature (10-20°C) should be first resolved before

successful operation will be possible. Also the balance between energy gain and CO2 footprint

of the WWTP should be considered when selecting the most sustainable solution. By this

time, interest in this concept is rising and resulted in already a full-scale trial in Strass

(Austria, in WERF project) and pilot set-up in Rotterdam (Paques).

3 General conclusions

The need to minimize the use of fossil fuel energy in the treatment of sewage, manures, agro-

industrial wastes and municipal solid waste organics will continue to increase in the future.

Anaerobic digestion allows to recover the chemical energy present as organic carbon and to

convert it to electrical energy. The latter can make the overall treatment plant self-sufficient in

electrical energy (energy index ranging from 1 to 5; Table 4.4) in the case of treatment of

manures, agro-industrial wastes and municipal solid waste organics. Yet, problems with

excess nitrogen and deficit of organic carbon in digestates to remove nitrogen by conventional

nitrification/denitrification warrant the development of a new process design.

Chapter 4

77

OLAND, integrated with anaerobic digestion, avoids external organic carbon addition and

allows to increase the energy index to 6-10 for the treatment of agro-industrial wastes and

municipal solid waste organics (Table 4.4).

OLAND application for sewage treatment only significantly lowers the electrical energy

demand if the amount of organic carbon normally needed for denitrification is captured in the

primary sludge by applying an enhanced primary settling or an improved A-stage. OLAND

application in the main stream of sewage plants will allow an energy index above 1 (Table

4.5).

OLAND, as a downstream process of anaerobic digestion is gradually becoming a mature

technology which in different configuration (SBR, MBBR, RBC, airlift, etc) can be reliable

operated at full-scale. Several full-scale cases are discussed here.

4 Acknowledgements

H.D.C. was recipient of a PhD grant from the Institute for the Promotion of Innovation by

Science and Technology in Flanders (IWT-Vlaanderen, number SB-81068). E.C and S.E.V.

were supported as a doctoral and postdoctoral fellow from the Research Foundation Flanders

(FWO-Vlaanderen), respectively. The authors gratefully thank Bernhard Wett and Tim

Hülsen for providing full-scale plant data and thank Tim Lacoere for technical support. This

work was supported by Ghent University Multidisciplinary Research Partnership (MRP) –

Biotechnology for a sustainable economy (01 MRA 510W).


78

Chapter 5

79

Chapter 5:

Efficient total nitrogen removal in an

ammonia gas biofilter through

high-rate OLAND

Abstract

Ammonia gas is conventionally treated in nitrifying biofilters, however addition of organic

carbon to perform post-denitrification is required to obtain total nitrogen removal. Oxygen-

limited autotrophic nitrification/denitrification (OLAND), applied in full-scale for wastewater

treatment, can offer a cost-effective alternative for gas treatment. In this study, the OLAND

application was broadened towards ammonia loaded gaseous streams. A down flow, oxygen-

saturated biofilter (height of 1.5 m; diameter of 0.11 m) was fed with an ammonia gas stream

(248 ± 10 ppmv) at a loading rate of 0.86 ± 0.04 kg N m-3

biofilter d-1

and an empty bed

residence time of 14 s. After 45 days of operation a stable nitrogen removal rate of 0.67 ±

0.06 kg N m-3

biofilter d-1

, an ammonia removal efficiency of 99%, a removal of 75-80% of

the total nitrogen and negligible NO/N2O productions were obtained at water flow rates of

1.3 ± 0.4 m3 m

-2 biofilter section d

-1. Profile measurements revealed that 91% of the total

nitrogen activity was taking place in the top 36% of the filter. This study demonstrated for the

first time highly effective and sustainable autotrophic ammonia removal in a gas biofilter and

therefore shows the appealing potential of the OLAND process to treat ammonia containing

gaseous streams.

Chapter redrafted after: De Clippeleir H., Courtens E., Mosquera M., Vlaeminck S.E., Smets

B.F., Boon N. and Verstraete W. 2012. Efficient total nitrogen removal in an ammonia gas

biofilter through high-rate OLAND. Environmental Science and Technology, Environmental

Science and Technology, 46(16), 8826-8833.

Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND

80

1 Introduction

Ammonia (NH3) is a colorless and reactive air pollutant that is an important cause of

acidification of soils and waters, and high levels of nitrate in surface and drinking waters. It is

commonly emitted from both industrial and agricultural activities such as wastewater

treatment plants, chemical and manufacturing industries, composting plants, and livestock

farming (Chung et al., 1996; Busca and Pistarino, 2003; Kim et al., 2007). In contrast to the

operational complexity and high costs of physico-chemical treatment processes, biological

treatment can offer cost-effective and straightforward purification of gas streams. The latter

biofiltration systems are mainly based on nitrification, transforming ammonia into nitrite and

nitrate, and as a result end up with a highly loaded percolate mixture of ammonium (NH4+),

nitrite (NO2-) and nitrate (NO3

-) (Baquerizo et al., 2009). To obtain dischargeable effluent,

post-denitrification with the addition of an external organic carbon source is applied or the

effluent is send to a central wastewater treatment facility (Sakuma et al., 2008; Cabrol, 2010).

Anoxic autotrophic nitrogen removal by anoxic ammonium-oxidizing or anammox bacteria

(AnAOB), able to combine nitrite with ammonium to N2 gas, can offer a solution in this

nitrogen rich biofilter environment devoid in organic carbon. Oxygen-limited autotrophic

nitrification/denitrification (OLAND) is a one-stage realization of partial

nitritation/anammox, the economically preferred nitrogen removal technology for

wastewaters with a biodegradable COD/N ratio below 3 (Kuai and Verstraete, 1998). This

process is based on the cooperation between aerobic ammonium-oxidizing bacteria

(AerAOB), which oxidize part of the ammonium to nitrite in the outer aerobic zones of the

biofilm, and AnAOB, which subsequently convert nitrite and ammonium to nitrogen gas in

the inner, anoxic zones. As a result, nitrogen is converted autotrophically in one step to

nitrogen gas. This autotrophic nitrogen removal process has been established in full-scale for

several wastewater treatment applications (Wett, 2006; Joss et al., 2009; Abma et al., 2010).

However, this process was thus far not applied for the treatment of gaseous ammonia-rich

streams.

Application of an OLAND biofilter would allow a total nitrogen removal, defined as a total

nitrogen loss based on gas and water composition, in the biofilter itself due to N2 gas

production by AnAOB. Although most ammonia gas biofilters are based on nitrification, a

total nitrogen removal efficiency is commonly observed ranging from 10 to 50% (Table 5.1).

Chapter 5

81

Total nitrogen removal can occur in the inert form of N2 or in the unsustainable form of NO

or N2O. However, the contribution of NO and N2O production to the total nitrogen removal

and the operational factors inducing higher total nitrogen removal rates are unclear. Until now

the total nitrogen removal was attributed to denitrification by heterotrophic and/or nitrifying

bacteria, both needing oxygen limitation, or by bacterial growth. The contribution of AnAOB

as a cause for total nitrogen removal in biofilters was not considered before despite the

presence of ammonium and nitrite and the occurrence of anoxic activity in these filters.

Heterotrophic denitrification is possible when organic compounds in the gas or water phase

are available and can lead to both N2 and NO/N2O formation (Juhler et al., 2009). Nitrifier

denitrification by aerobic ammonium-oxidizing bacteria (AerAOB) implicates nitrogen

removal by NO and N2O formation instead of N2 production (Chandran et al., 2011), and

hence negatively affects the sustainability of the technology. It was reported that almost 20%

of the NH3 loading can be converted to N2O by autotrophic and/or heterotrophic

denitrification (Maia et al., 2012). Finally, nitrogen can also be incorporated into the biomass

and used for growth. It was estimated that the nitrogen incorporation in biomass accounts for

7% of the nitrogen input (Cabrol, 2010). The stimulation of AnAOB in the biofilter and thus

application of the OLAND process for the treatment of ammonia containing gas streams

could offer two advantages. Firstly, AerAOB inhibition by free ammonia or free nitrous acid

is commonly observed in biofilters and results in ammonium to nitrate ratios in the percolate

of around 1 (Smet et al., 2000; Chen et al., 2005; Baquerizo et al., 2009; Cabrol, 2010). The

lower ammonium consumption rate by AerAOB can be compensated during the OLAND

process by ammonium consumption by AnAOB. Secondly, the higher the AnAOB activity in

the filter, the higher the nitrogen gas production rate, and thus the higher the total nitrogen

removal rate in the filter will be. Together, these two facts will decrease the need for post-

treatment of the percolate and consequently the cost for external organic carbon source

addition. The goal of this study was to demonstrate the possibility to obtain fully autotrophic

total nitrogen removal in an ammonia gas biofilter through a combination of AerAOB and

AnAOB activity, also referred to as the OLAND process. This is the first study showing

anammox as the main nitrogen removal process in ammonia gas biofilters.


82

Table 5.1: Overview of the operational parameters and nitrogen losses in ammonia gas biofilters. DF/UF: down flow/upflow reactors, H/D: height over

diameter ratio; EBRT: empty bed residence time

N loss DF/UF Packing

material

Height H/D EBRT NH3 in Loading rate (++)

Water

flow rate(+)

Temperature Reference

% m s ppm kg N m-3

biofilter d-

1

kg N m-2

biofilter

section

d-1

m3 m

-2

biofilter d-

1

°C

0 UF Slow release 1.0 10 20-36 90-260 0.1-0.6 0.1-0.6 0.1° 24 (Baquerizo et al. 2009)

(2009) 0-30 DF Slow release 0.3 3 14 270-

700 1.3-3.0 0.4-0.9 3.9 22-25 (Sakuma et al. 2008)

15 DF Slow release 0.6 4 32-85 10-150 0.1-0.2 0.07-0.1 1.6 30 (Kim et al. 2007)

16-32 UF Slow release 1.0 7 30-35 50-200 0.1-0.6 0.1-0.6 0° 25-30 (Chen et al. 2005)

52 DF Slow release 1.5 9 54 35 0.03 0.05 0.4 20-25 (Cabrol 2010)

30-60 DF Slow release 1.4 14 50 35-170 0.1-0.3 0.1-0.5 0.08 20-30 (Malthautier et al. 2003)

(2003) 98* DF Inert 0.6 11 60 100-

600 0.1-0.5 0.05-0.3 72 20-25 (Moussavi et al. 2011)

75-80 DF Inert 1.6 14 14 250 ±

10 0.9 ± 0.1 1.3 ± 0.1 1.2 ± 0.4 20-25 This study

*External organic carbon source addition in filter to obtain simultaneous nitrification/denitrification

°The air flow was humidified before entering the biofilter

The water to N ratio expressed as L water g-1

Nin can be calculated by dividing (+) by (++)

Chapter 5

83


2.1 Biofilter set-up and operation

The biofilter consisted of a PVC cylindrical column with a height of 1.57 m and an internal

diameter of 0.11 m. The section surface of the filter was thus 95 cm2. The column was packed

with Kaldnes K1 packing material (AnoxKaldnes, Lund, Sweden) and on 50% of the carriers

OLAND biomass from a stably working OLAND rotating contactor was added (Pynaert et al.,

2003), resulting in an initial biomass concentration of 3.8 g VSS L-1

biofilter. The total

contact surface based on the specific surface of the Kaldnes rings was estimated at 800 m2 m

-3

total reactor. The inlet ammonia stream was supplied at the top as a mixture of compressed air

and pure ammonia, and was controlled by two digital mass flow controllers (Bronkhorst, The

Netherlands) to ensure a stable inlet concentration of 248 ± 10 ppmv, a gas velocity of 0.1 m

s-1

and a gas empty bed residence time (EBRT) of 14 seconds. The biofilter was humidified

by discontinuously spraying (1 second every 5 minutes) tap water at an initial flow rate of 0.8

m3 m


-1 on top of the filter. The filter was operated at room temperature

(23±1°C). Daily, samples were taken from the gas in- and outlet (200 mL) and from the water

phase (10 mL) to determine the ammonia, ammonium, nitrite and nitrate concentration.

Nitrous oxide and nitric oxide concentration were only measured during the profile

measurements.

2.2 Profile measurements

On days 90 and 99, gas and water samples were taken at 0, 7, 32, 57, 82, 107, 132 and 157 cm

depth from the top for the detection of NH3, O2, NO, N2O and NH4+, NO2

- and NO3

-. In all

water samples, the pH was also measured. These measurements allowed obtaining vertical

activity profiles.

2.3 Activity batch test

On day 125, the specific activities of AerAOB, AnAOB and nitrite oxidizing bacteria (NOB)

in the different zones of the biofilters (see profile measurements) was determined in separate

activity tests in aqueous media at 22°C and at initial nitrogen concentration of 100 mg N L-1

,

as described by Vlaeminck et al. (2007).


84


NH3 was measured in the gas phase with colorimetric gas detection tubes (RAE, Hoogstraten,

Belgium), using 100 mL of gas sample. The NH3 detection tubes had a detection limit of 1

ppmv NH3 (0.62 mg NH3-N L-1

). The N2O and O2 concentrations in the gas phase were

analyzed with a Compact GC (Global Analyser Solutions, Breda, The Netherlands), equipped

with a Porabond precolumn and a Molsieve SA column. The thermal conductivity detector

had a detection limit of 1 ppmv for each gas component. NO measurements were done based

on the principle of chemiluminescence using Eco Physics CLD 77 AM (Eco Physics AG,

Duernten, Switzerland) with a detection limit of 1 ppbv. Ammonium (Nessler method) and

VSS (after removing the biomass from the carriers) were measured according to standard

methods (Greenberg et al. 1992). Nitrite and nitrate were determined on a Metrohm 761

Compact Ion Chromatograph (Zofingen, Switzerland) equipped with a conductivity detector.

Dissolved oxygen (DO) and p were measured with, respectively, an 0d DO meter

( ach Lange, D sseldorf, Germany) and an electrode installed on a C833 meter (Consort,

Turnhout, Belgium).

Table 5.2: Overview of the primers sets and conditions used for determination of the abundance of

AerAOB, AOA, AnAOB and NOB with qPCR.

Functional

group

Target

gene

Primers Sequences (5´-3´) Melting

temp

(ºC)

Ref.

AerAOB amoA gene amoA- 1F

amoA-2R

GGGGTTTCTACTGGT

GGT

CCCCTCKGSAAAGCC

TTCTTC

55 (1)

AOA Creanarchaeal

amoA gene

CrenamoA23f

Creanamo A616r

ATGGTCTGGCTWAG

ACG

GCCATCCATCTGTAT

GTCCA

56 (2)

Nitrospira sp. 16S rRNA Nspra675f GCG GTG AAA TGC

GTA GAK ATC G

67.2 (3)

Nitrospira sp. 16S rRNA Nspra746r TCA GCG TCA GRW

AYG TTC CAG AG

65.3 (3)

AnAOB 16S rRNA Amx809f GCC GTA AAC GAT

GGG CAC T

67.1 (4)

AnAOB 16S rRNA Amx1066r ATG GGC ACT MRG

TAG AGG GGT TT

67.4 (4)

(1): Rotthauwe et al. (1997); (2) Tourna et al. (2008); (3) Graham et al. (2007); (4) Tsushima et al.

(2007a)

Chapter 5

85

2.5 Quantification with real-time PCR

Biomass samples (approx. 5 g) for nucleic acid analysis were taken from the OLAND rotating

contactor (inoculum of the biofilter) and at 7, 32, 57, 82, 107, 132 and 157 cm depth after 125

days of operation. DNA was extracted using FastDNA® SPIN Kit for Soil (MP Biomedicals,

LLC), according to the manufacturer’s instructions. The obtained DNA was purified with the

Wizard® DNA Clean-up System (Promega, USA) and its final concentration was measured

spectrophometrically using a NanoDrop ND-1000 spectrophotometer (Nanodrop

Technologies). The SYBR Green assay (Power SyBr Green, Applied Biosystems) was used to

quantify the 16S rRNA of bacterial anammox bacteria and Nitrospira sp. and the functional

amoA gene for AerAOB and ammonium-oxidizing archaea (AOA). The primers for

quantitative polymerase chain reactions (qPCR) used in this study are listed in Table 5.2.

Plasmid DNAs carrying AerAOB, AOA functional AmoA gene and Nitrospira and anammox

16SrRNA gene, respectively, were used as standards for qPCR.

3 Results

3.1 Performance of the biofilter

In a biomass free control test with inert Kaldnes K1 packing material, all nitrogen inserted via

the gas phase as NH3 could be found back in the effluent gas and water phase, excluding

nitrogen removal by leakages. After inoculation of the biofilter with active OLAND biofilm

on the Kaldnes K1 packing material, the biofilter was immediately fed with an ammonia gas

stream, without acclimatization of the biomass by water recirculation, at a loading rate of 0.88

± 0.04 kg N m-3

biofilter d-1

. After 31 days of operation, the ammonia gas removal efficiency

remained stable around 99 ± 0.7%, independent of the operational conditions (Fig. 5.1).

Although a high pH value around 8.3 ± 0.6 was measured during the start-up period (phase I,

Fig. 5.1), only 20 ± 5% of the nitrogen load was detected in the percolate as ammonium and

the total nitrogen removal accounted already for 53 ± 11% of the total nitrogen load. During

phase II (Fig. 5.1), an ammonium decrease and nitrite and nitrate increase in the percolate

together with higher total nitrogen removal efficiencies up of 70 ± 5% were accompanied by a

pH decrease from 8.3 ± 0.6 (Phase I) to 6.6 0.4 (Phase II). From day 45 onwards, the

decrease in pH was stabilized by addition of coccolith lime on the biofilter top (on average

0.7 kg m-3

biofilter d-1

) resulting in a pH value of 6.9 ± 0.3 during the following operation

period (end of phase II and phase III). During phase III, the influence of the water flow rate


86

on the biofilter performance was tested as the latter influences the NH3 dissolution and the

nitrogen concentration at which the bacteria are exposed to. A small increase in the water

flow rate from 1.2 ± 0.6 to 1.7 ± 0.2 m3 m


-1 combined with the stable pH

conditions allowed higher total nitrogen removal efficiencies of 79 ± 6% between day 73 and

day 90. The latter was mainly due to higher ammonium removal efficiencies (Fig. 5.1). A

decrease from day 91-105 of the water flow rate to 1.2 ± 0.4 m3 m


-1 did

not have a significant effect on the removal performance. Moreover, during the increase of the

water flow rate up to 2.4 ± 0.7 m3 m


-1 (day 106-125), the removal

efficiency remained stable around 79 ± 7% and only the absolute concentration in the

percolate decreased due to dilution. A NH3 gas inlet failure (day 114), resulting in 1 day

without NH3 addition, had no significant influence on the performance afterwards.

Chapter 5

87

Figure 5.1: Nitrogen loading and removal rates (top) and corresponding contribution of the different

nitrogen species in the emitted air and water flow, as a percentage of the incoming nitrogen (bottom).

The nitrogen removal is considered to be due to N2 formation, since profile measurements showed

negligible amounts of NO (0.5% of incoming N) and N2O (below detection limit) production. Three

main periods are distinguished: a start-up period (phase I); a pH stabilization period (phase II) and a

water flow rate optimization period (phase III).


88

Table 5.3: Operational conditions measured directly in the filter and microbial activities measured in separate aqueous activity tests (n=3) at different top

down biofilter zones. n.d.: not detected; AerAOB: aerobic ammonium oxidizing bacteria; AnAOB: anoxic ammonium oxidizing bacteria; NOB: nitrite

oxidizing bacteria

Water flow rate Top down biofilter zone

(m3 m

-2 biofilter

section d-1

)

7-32 cm 32-57 cm 57-82 cm 82-107 cm 107-132 cm 132-157 cm

pH (-) 1.4 8.6 7.2 7.1 6.5 6.3 7.1

1.1 8.1 6.6 6.8 6.0 6.3 7.1

Free ammonia (mg N L-1

) 1.4 61 0.4 0.4 0.07 0.03 0.2

1.1 51 1.0 0.4 0.07 0.09 0.5

NO2- (mg N L

-1) 1.4 200 147 119 122 30 71

1.1 441 411 248 254 21 121

Microbial group

Total anoxic nitrogen removal rate

(mg N g-1

VSS d-1

)

AnAOB 13 ± 3 9 ± 3 13 ± 4 2 ± 2* n.d.* n.d.*

Aerobic ammonium oxidation rate

(mg N g-1

VSS d-1

)

AerAOB 142 ± 52 252 ± 27 389 ± 10 226 ± 54 149 ± 59 244 ± 37

Aerobic nitrate production rate

(mg N g-1

VSS d-1

)

NOB 1 ± 1 n.d. 19 ± 14 157 ± 63 140 ± 26 136 ± 46

*Anoxic nitrite consumption without anoxic ammonium consumption was observed, but this should not be considered as AnAOB activity.

Chapter 5

89

3.2 Vertical distribution of microbial activity

The vertical profile measurement during phase III showed that the highest microbial activity

occurred in the top 0.57 m of the biofilter (Fig. 5.2), while at all heights oxygen was saturated

in the gas phase. Ammonia dissolved for 80-95%, depending on the water flow rate, in the

first 32 cm of the biofilter (Fig. 5.2). In the first 7 cm, only dissolution and no microbial

activity occurred. In the subsequent zones the highest total nitrogen removal rates were

observed. In these zones, ammonium and nitrite were consumed without an equivalent nitrate

production (Fig. 5.2). In this upper 57 cm of the filter, 91% of the total nitrogen removal was

taking place (Fig. 5.2A) and according to the stoichiometry, the absence of organic carbon

and the absence of NO or N2O production, this was mainly attributed to the AnAOB. In the

lower two thirds (> 57 cm depth, Fig. 5.2A) some denitrification (9% of the total nitrogen

removal), probably using organics from bacterial decay, occurred. Although the total nitrogen

removal rate in the biofilter remained constant when the water flow rates was decreased from

1.4 to 1.1 m3 m


-1 (Fig. 5.2B), a downward shift of the OLAND activity

from 0.07-0.57 m to 0.32-0.82 m was observed. This was probably attributed to the inhibition

of the AnAOB activity by higher nitrite concentrations in the upper section of the filter and

the slower dissolution of NH3 (Table 5.3).

Biomass samples were taken at 6 biofilter zones and the AnAOB, AerAOB and NOB activity

was tested in aqueous medium. Despite the 4-fold lower total nitrogen removal rate compared

to the biofilter performance (0.2 compared to 0.8 kg N m-3

biofilter d-1

), the vertical profile

distribution of the AnAOB activity confirmed the direct biofilter profile measurements (Table

5.3). AnAOB activity was measured in the zone 7-82 cm and decreased rapidly in the lower

compartments (Table 5.3). In the lower zones (> 82 cm), (nitrifier) denitrification could take

place because anoxic nitrite consumption was taking place while no difference in ammonium

concentration was observed (data not shown). AerAOB were active over the total depth of the

biofilter, while NOB started to show activity at the lower part of the biofilter (> 82 cm). The

total biomass concentration, measured after emptying the biofilter, increased during 125 days

of operation from 3.8 g VSS L-1

biofilter to 19 g VSS L-1

biofilter, with the highest

concentration at 7-32 cm depth.


90

Figure 5.2: Vertical profile measurement expressed as NH3, NH4+, NO2

- and NO3

- productions based

on the gas and water phase analysis at day 90 (A) and day 99 (B) operated at water flow rates of 1.4

and 1.1 m3 m


-1, respectively. Total NO production was negligible (0.5% of

nitrogen input) and N2O production was not detected.

3.3 Vertical abundance of N species

The biofilter was inoculated with biomass containing 2 102 AerAOB-AmoA copies, 9 10

3

AnAOB-16SrRNA copies and 2 102 Nitrospira-16SrRNA copies ng

-1 DNA, which was

homogeneously distributed over the filter. AOA were not detected in the inoculum. However

after 125 days of operation, up to 2 102 AOA-AmoA copies ng

-1 DNA were detected (Fig.

5.3). AnAOB abundance remained constant over the filter. AerAOB showed a peak

concentration at a depth of 57-82 cm, which correlated well with the activity test (Table 5.3).

Chapter 5

91

The Nitrospira abundance increased significantly to 2 105 Nitrospira-16SrRNA copies ng

-1

DNA below a depth of 82 cm. The observed decrease in inhibition factors such as free

ammonia and the NOB activity measured at these zones (Table 5.3) confirmed the abundance

measurements.

Figure 5.3: Abundance of AerAOB, AOA, AnAOB and Nitrospira, expressed as copies of AerAOB-

AmoA, AOA-amoA, AnAOB-16SrRNA and Nitrospira-16SrRNA ng-1

DNA, respectively, in the

inoculum and in the different biofilter zones after 125 days of operation.

4 Discussion

4.1 OLAND application for NH3 treatment

This study showed for the first time that AnAOB activity can be obtained in an oxygen-

saturated biofilter treating a NH3 gas stream. The application of the OLAND process instead

of nitrification in the biofiltration technology would allow higher total nitrogen removal in the

biofilter itself (up to 80%), significantly decreasing the cost for an external carbon source

addition needed for post-denitrification of the percolate. Moreover, this study showed that by

implementing the OLAND process, a sustainable total nitrogen removal can be obtained

without NO and N2O formation. The unsustainable nitrogen removal during conventional

NH3 treatment is in most studies neglected and not measured (Table 5.1), but is expected to be

high (up to 20% of the nitrogen loading rate can be emitted as N2O; Maia et al., 2012).


92

4.2 AnAOB niche in NH3 biofilters

Total nitrogen removal in NH3 fed biofilters has been reported in several studies (Table 5.1).

However, the total nitrogen removal efficiency was mainly lower than 60%, while in this

study a total nitrogen removal efficiency of almost 80% was obtained (Table 5.1). The total

nitrogen removal rates obtained in the NH3 fed OLAND biofilter were in the same range as

OLAND application for wastewater treatment (Vlaeminck et al., 2012). Generally, the total

nitrogen removal in ammonia gas biofilters can be attributed to several processes: (i)

denitrification; (ii) nitrifier denitrification; (iii) nitrogen biomass incorporation, (iv) chemical

reactions and as shown in this study (v) anammox. Because inert packing material was used in

this study and no organic carbon source was present in the gas or water flow, the contribution

of denitrification to the total nitrogen removal was considered to be minor. In contrast to

several studies suggesting that AerAOB were responsible for the total nitrogen removal due to

nitrifier denitrification (Chen et al., 2005; Kim et al., 2007), the latter pathway could be

excluded in this study because no N2O and very low NO emissions (0.5% of N loading) were

detected. Also chemical reactions leading to NO or N2O formation could be neglected in this

study (Chandran et al., 2011; Vermeiren et al., 2012). Nitrogen incorporation in the biomass

could probably explain for a part the 9-15% nitrogen loss that was measured during the

profile measurements but that was, based on the stoichiometry not caused by AnAOB

activity. The total nitrogen removal to N2 in the top part of the filter (> 82 cm) was attributed

to AnAOB activity as 85-91% of the nitrogen removal took place at the biofilter zones where

ammonium and nitrite consumption was observed (Fig. 5.2) and as the specific activity tests

confirmed AnAOB activity in the top part of the filter (Table 5.3). Moreover, if denitrification

had been responsible for the total nitrogen removal, at least 2 kg COD m-3

biofilter d-1

should

have been consumed, corresponding with 1.5 kg VSS m-3

biofilter d-1

, or around 40% of the

inoculated biomass organics, allowing no biomass growth in the filter.

Data on the microbial community structure in NH3 biofilters are still relatively scarce,

compared to other engineered systems such as bioreactors for wastewater treatment. Studies

performed on NH3 fed biofilters discuss mainly overall diversity and dynamics (Cabrol et al.,

2010) or focus only on the AerAOB (Juhler et al., 2009; Yin and Xu, 2009; Yasuda et al.,

2010) or ammonium-oxidizing Archaea (Yasuda et al., 2010). The anaerobic ammonium

oxidation was not considered before in this application domain. Moreover, due to a lack of

information about the relation between the community structure and the total nitrogen balance

Chapter 5

93

in these biofilters (Table 5.1), there was no evidence of the presence of Planctomycetes and

more specifically AnAOB in NH3 fed biofilters. However, this study showed that AerAOB in

close relationship with AnAOB can cause high nitrogen removal rates in gaseous biofilters.

Both substrates ammonia and nitrite are commonly present in biofilters due to the high

AerAOB activity and lower NOB activity (Baquerizo et al., 2009), which indicate a niche

environment for AnAOB, provided that anoxic conditions are created.

As the biofilter was fed under fully aerobic conditions, anoxic zones should have been present

to allow AnAOB activity. It could be calculated that anoxic zones could be obtained in the

biofilm itself when the thickness of an oxygen-consuming biofilm was higher than 84 μm

(Perez et al., 2005) given oxygen saturation in the gas phase over the whole depth of the

biofilter. On the other hand, preferential gas and water flow due to a low ratio between the

biofilter reactor diameter and packing material diameter (11 < 12), could probably occur

allowing oxygen gradients in the filter (Beavers et al., 1973).

Due to the high free ammonia concentration and consequently NOB inhibition (Anthonisen et

al., 1976) at the top of the biofilters (Table 5.3), total nitrogen removal by AnAOB was

mainly taking place between 7-57 cm depth despite the high nitrite levels (around 200 mg

NO2--N L

-1). AnAOB can irreversibly be inhibited by nitrite. However the reported inhibition

range (100-350 mg N L-1

) is broad and the effect depends on the AnAOB species (Strous et

al., 1999; Egli et al., 2001; Dapena-Mora et al., 2007). In this study, inhibition of AnAOB was

only observed at nitrite levels above 411 mg N L-1

(Table 5.3), and this effect seemed

reversible in case the water flow rate was adjusted. Therefore, the water flow rate relative to

the nitrogen gas loading of the system, calculated as the ratio between the water flow rate

(m3 m


-1) and the nitrogen loading rate (kg N m


-1),

determined the degree of AnAOB activity over NOB activity in the system as well as the

AnAOB over NOB abundance. Water to N ratios lower than 1 L g-1

Nin resulted in both

AnAOB and NOB inhibition in the top layers (Fig. 5.2B), while AnAOB were favored above

NOB at higher water to N ratios (around 1 L g-1

Nin). In most studies reporting minor total

nitrogen removal efficiencies (Table 5.1), this ratio was high (>>1 L g-1

Nin) decreasing free

ammonia concentrations in the filter (higher dilution), and consequently allowing NOB

activity. As a result, AnAOB probably did not have a competitive advantage compared to

NOB in the top layers and could not significantly invade the filter in contrast to the OLAND

biofilter in this study. So, to obtain high nitrogen losses without significant N2O emissions


94

and thus a niche for AnAOB, high nitrogen levels together with low water to N ratios (around

1 L g-1

Nin) are advised.

4.3 OLAND: gas versus water treatment

OLAND is considered an established technology for the treatment of digestates in several

application domains (Chapter 4) as it can provide high and stable performance and decreased

operational cost by decreasing the energy consumption and avoiding the addition of external

organic carbon (Vlaeminck et al., 2012). Compared to these water applications, the OLAND

biofilter for gas treatment could offer an additional advantage. The optimal balance between

AerAOB, AnAOB and NOB activity is more easily obtained without complicated control

systems as needed during wastewater treatment. In the latter application, NOB activity is

avoided by a combination of DO control, free ammonia levels and specific SRT control of

aerobic flocs (Joss et al., 2011), which significantly increases the operational complexity.

Moreover, the control of the microbial balance becomes more difficult when treating low

nitrogen concentration (Chapter 7). In the OLAND biofilter, the nitrogen gas flow, although

containing low NH3 concentrations, is concentrated in the water film on top of the biofilm,

allowing more easily NOB inhibition by free ammonia or even free nitrous acid. Therefore,

besides saving costs for further percolate treatment, the OLAND biofilter can be stably

operated at minimal operational complexity.

5 Conclusions

This study demonstrated for the first time highly effectient (up to 80%) and sustainable

(negligible NO/N2O emission) autotrophic ammonia removal, based on AnAOB activity in a

gas biofilter. Therefore, this study shows the appealing potential of the OLAND process to

treat ammonia containing gaseous streams.

6 Acknowledgements

H.D.C. was a supported by a PhD grant from the Institute for the Promotion of Innovation by

Science and Technology in Flanders (IWT-Vlaanderen, number SB-81068). E.C and S.E.V.

were supported as doctoral candidate (Aspirant) and a postdoctoral fellow, respectively, from

the Research Foundation Flanders (FWO-Vlaanderen). The authors thank Samuel Bodé for

kind assistance with NO analyses and Joachim Desloover, Tom Hennebel and Frederiek-

Maarten Kerckhof for inspiring scientific discussions.

Chapter 6

95

Chapter 6:

OLAND is feasible to treat sewage-

like nitrogen concentrations at low

hydraulic residence times

Abstract

Energy-positive sewage treatment can in principle be obtained by maximizing energy

recovery from concentrated organics, and by minimizing energy consumption for

concentration and residual nitrogen removal in the main stream. To test the feasibility of the

latter, sewage-like nitrogen influent concentrations were treated with oxygen-limited

autotrophic nitrification/denitrification (OLAND) in a lab-scale rotating biological contactor

(RBC) at 25°C. At influent ammonium concentrations of 66 and 29 mg N L−1

and a

volumetric loading rate of 840 mg N L−1

d−1

yielding hydraulic residence times (HRT) of 2

and 1 h, respectively, relatively high nitrogen removal rates of 444 and 383 mg N L−1

d−1

were obtained, respectively. At low nitrogen levels, adapted nitritation and anammox

communities were established. The decrease in nitrogen removal was due to decreased

anammox and increased nitratation, with Nitrospira representing 6% of the biofilm. The latter

likely occurred given the absence of dissolved oxygen (DO) control, since decreasing the DO

concentration from 1.4 to 1.2 mg O2 L−1

decreased nitratation by 35% and increased

anammox by 32%. Provided a sufficient suppression of nitratation, this study showed the

feasibility of OLAND to treat low nitrogen levels at low HRT, a prerequisite to energy-

positive sewage treatment.

Chapter redrafted after: De Clippeleir, H., Yan, X., Verstraete, W., Vlaeminck, S.E., 2011.

OLAND is feasible to treat sewage-like nitrogen concentrations at low hydraulic residence

times. Applied Microbiology and Biotechnology, 90, 1537-1545.

OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT

96

1 Introduction

Biological nitrogen removal is economically preferred above physicochemical nitrogen

recovery for wastewaters containing less than 5 g N L−1

(Mulder, 2003). Furthermore, if the

ratio of biodegradable chemical oxygen demand (bCOD) to nitrogen is relatively low

(typically ≤ 2-3), nitrogen removal with partial nitritation and anammox saves about 60% of

the aeration, 90% of the sludge handling and transport, and 100% of the organic carbon

addition compared to conventional nitrification/denitrification (Mulder, 2003). Overall some

30-40% of the overall nitrogen removal costs can be saved (Fux and Siegrist, 2004). Oxygen-

limited autotrophic nitrification/denitrification (OLAND) is a one-stage configuration of this

process (Kuai and Verstraete, 1998), in which aerobic ammonium-oxidizing bacteria

(AerAOB) oxidize about half of the ammonium to nitrite in the outer, aerobic zones of the

biomass (partial nitritation), while the anoxic ammonium-oxidizing bacteria (AnAOB)

subsequently convert nitrite and the residual ammonium to mainly nitrogen gas (89%) and

some nitrate (11%) in the inner, anoxic zones (anammox; Pynaert et al., 2003; Vlaeminck et

al., 2010). Oxygen plays a key role in balancing the microbial activities (Fig. 6.1A), with on

the one hand an oxygen requirement of 1.8 g O2 g−1

N to achieve sufficient ammonium

oxidation while avoiding excess nitrite production by AerAOB. On the other hand,

sufficiently low dissolved oxygen (DO) levels (e.g. 0.3 mg O2 L−1

) are needed to suppress

excess nitrate production by nitrite-oxidizing bacteria (NOB) (Joss et al., 2009).

Conventional activated sludge (CAS) systems for sewage treatment have low volumetric

carbon and nitrogen loading rates (around 1 g COD L−1

d−1

and 0.08 g N L−1

d−1

) and are

energy-negative. The aeration required for organic carbon and nitrogen removal constitutes

about 60-70% of the total energy consumption of a sewage treatment plant (Zessner et al.,

2010). However, if enhanced primary settling is applied to increase physico-chemical sludge

production and if OLAND is used for nitrogen removal from the digestate of primary and

secondary sludge (Fig. 6.1B), the aeration requirements of the CAS step can be decreased

with 25% (Siegrist et al., 2008). Over the last five years several OLAND-type treatments

were developed to treat sewage sludge digestates (Joss et al., 2009; Jeanningros et al., 2010).

Furthermore, if primary settling is replaced by a highly loaded activated sludge step, where

organic matter is converted to biomass at maximal yield, energy-neutrality is achievable given

the even higher conversion of bCOD by anaerobic digestion into biogas and hence electricity

(Wett et al., 2007). Given the high energetic content of the sewage bCOD, energy-positive

Chapter 6

97

sewage treatment should be possible (Siegrist et al., 2008; Verstraete et al., 2009; Kartal et al.,

2010a). This requires an advanced biological or physicochemical bCOD concentration step to

further increase energy recovery from anaerobic digestion of concentrated organics and a low

energy demand for the concentration step and the removal of residual nitrogen (and some

bCOD) in the main stream (Fig. 6.1B). The energy requirement for OLAND is influenced by

the reactor configuration: active aeration in sequencing batch reactors requires 1.3 kWh kg−1

N (Wett et al., 2010b), whereas passive aeration in rotating biological contactors (RBC)

requires down to 0.4 kWh kg−1

N (Mathure and Patwardhan, 2005). Depending on the

dilution, sewage is typically composed of 30-100 mg N L−1

and 450-1200 mg COD L−1

rendering a COD/N ratio of about 12 to 15 (Metcalf and Eddy, 2003; Tchobanoglous et al.,

2003; Henze et al., 2008). An advanced concentration step is expected to separate up to 75-

80% of the COD (Verstraete et al., 2009) and about 20% of the sewage nitrogen, mainly

consisting of colloidal and particulate organic nitrogen, from which the anaerobically

hydrolyzed part is returned to the main stream as ammonium (Fig. 6.1B). Hence, the OLAND

stage would receive nitrogen as ammonium at a COD/N ratio below 4, which is theoretically

low enough to avoid the risk that heterotrophs overgrow AnAOB (Lackner et al., 2008).

Until now, the OLAND process has been applied for medium and high-strength nitrogen

wastewaters (> 0.2 g N L−1

) such as landfill leachate and digestates from sewage sludge,

specific industrial streams and concentrated black water at relatively high hydraulic residence

times (HRT, Table 6.1). To obtain reasonably high nitrogen removal rates (400 mg N L−1

d−1

),

the treatment of low nitrogen levels (< 80 mg N L−1

) has to occur at low HRT, in the order of

some hours, rendering biomass retention an important requirement. In this study, a first and

exploratory step towards the implementation of OLAND in the new sewage treatment scheme

was tested, feeding sewage-like ammonium influent concentrations without COD addition.

Given its low energy consumption for aeration, a RBC was chosen as lab-scale reactor, and

operated at 25°C, simulating the maximum sewage temperatures in summer (Breda, NL;

Mollen, personal communication). This is one of the first tests on the OLAND treatment of

low nitrogen concentrations at such low HRT, a prerequisite to energy-positive sewage

treatment.


98

Figure 6.1: A. Conversion of nitrogen species, oxygen and protons in oxygen-limited autotrophic nitrification/denitrification (OLAND), showing balanced

and imbalanced contributions of three bacterial groups, i.e. aerobic ammonium-oxidizing, nitrite oxidizing and anoxic ammonium-oxidizing bacteria

(AerAOB, NOB and AnAOB, respectively); B. Conventional and redesigned sewage treatment schemes with OLAND in the side and main line, respectively.

In the redesigned scheme, energy-positive sewage treatment can in principle be obtained by maximizing energy recovery through anaerobic digestion of

concentrated organics in the side stream, and by minimizing energy consumption for the physicochemical and/or biological concentration step and the residual

nitrogen removal step, applying OLAND.

Chapter 6

99

Table 6.1 Overview of typical average nitrogen concentrations, volumetric loading/removal rates and hydraulic residence times (HRT) for existing one-step

partial nitritation/anammox processes and for the low nitrogen concentration and HRT application in this study. RBC: rotating biological contactor;

SBR: sequencing batch reactor

Wastewater Influent

concentration

(mg NH4+-N L

−1)

N loading rate

(g N L−1

d−1

)

N removal rate

(g N L−1

d−1

)

HRT

(d)

Reactor

type

Reference

Digested black water 1023 0.94 0.71 1.33 RBC (Vlaeminck et al., 2009b)

Sewage sludge digestate 800 0.74 0.67 0.93 SBR (Jeanningros et al., 2010)

Sewage sludge digestate 650 0.54 0.51 1.20 SBR (Joss et al., 2009)

Industrial digestate 300 2.0 1.17 0.18 Gas-lift (Abma et al., 2010)

Landfill leachate 209 0.38 0.38 0.55 RBC (Hippen et al., 1997)

Landfill leachate 250 0.67 0.41 0.51 RBC (Siegrist et al., 1998)

Sewage-like nitrogen concentrations 66 0.86 0.44 0.08 RBC This study

Sewage-like nitrogen concentrations 31 0.84 0.38 0.04 RBC This study


100

2 Material and methods

2.1 OLAND rotating biological contactor (RBC)

The lab-scale RBC was based on an airwasher LW14 (Venta, Weingarten, Germany) with a

rotor consisting of 40 discs interspaced at 3 mm, resulting in a disk contact surface of 1.32 m2.

The reactor had a liquid volume of 3.6 L, immersing the discs for 64%. The reactor

temperature was set at 25°C and the pH was adjusted to be higher than 7.3 by the addition of

NaHCO3. The DO concentration was not directly controlled. For continuous rotation the

rotation speed was fixed at 3 rpm and in the intermittent rotation mode, rotation at the same

rotation speed occurred only 1/3 of the time, equally spread over time (1 min on, 2 min off).

2.2 Reactor operation

The influent of an OLAND lab-scale rotating biological contactor (RBC), as used by

Vlaeminck et al. (2009b) to treat digested black water (Table 6.1), was switched to synthetic

wastewater consisting of (NH4)2SO4, NaHCO3, KH2PO4 (10 mg P/L) and 2 mL L−1

of a trace

element solution (Kuai and Verstraete, 1998). After a long term stable operation of the reactor

treating 537 mg N L−1

, the influent ammonium concentration was stepwise decreased to 278,

146, 66 and 31 mg N L−1

over 41, 48, 52 and 60 days, respectively, maintaining a constant

loading rate (about 840 mg N L−1

d−1

) by a stepwise decrease in hydraulic residence time

(HRT) (Table 6.2). Each nitrogen influent concentration was applied for 1.5 to 2 months to

obtain enough data points and stabilization for statistical comparison between the phases.

Reactor pH, DO and temperature were daily monitored and influent and effluent samples

were taken at least thrice a week for ammonium, nitrite and nitrate analyses.


Ammonium (Nessler method) was determined according to standard methods (Greenberg et

al., 1992). Nitrite and nitrate were determined on a 761 compact ion chromatograph equipped

with a conductivity detector (Metrohm, Zofingen, Switzerland). DO and pH were measured

with respectively, an electrode installed on a C833 meter (Consort, Turnhout, Belgium) and a

HQ30d DO meter (Hach Lange, Düsseldorf, Germany).

Chapter 6

101

2.4 Fluorescent in-situ hybridization (FISH)

At the start (537 mg N L−1

) and at the end (29 mg N L−1

) of experiment biomass samples were

taken from the discs and bottom of the reactor for identification of the autotrophic nitrogen

removing species present. At both time points FISH quantification of AerAOB, AnAOB and

NOB was performed. A paraformaldehyde (4%) solution was used for biofilm fixation and

FISH was performed according to Amann and coworkers (1990). Relevant target groups were

gathered from a recent nitrogen cycle review (Vlaeminck et al., 2011), and probe sequences

and formamide concentrations were applied according to (Lücker, 2010) for Nitrotoga and

probeBase for the other targets (Loy et al., 2003): Amx820 for the AnAOB

Kuenenia/Brocadia; a mixture of NSO1225 and NSO190 for the b-proteobacterial AerAOB

Nitrosomonas/Nitrosospira; and NIT3 (+ competitor), Ntspa662 (+ competitor) and Ntoga221

for the NOB Nitrobacter, Nitrospira and Nitrotoga, respectively. The AnAOB, AerAOB and

NOB abundance was evaluated by combining the specific probe with an equimolar mixture of

EUB338I, II and III, targeting all bacteria, and 4'-6-diamidino-2-phenylindole (DAPI),

targeting all DNA-containing cells. Image acquisition was done on a Zeiss Axioskop 2 Plus

epifluorescence microscope (Carl Zeiss, Germany). For quantification, 20 randomly taken

images were analyzed with ImageJ software, and the percentage of the specific group was

calculated as the ratio of the specific area to the total DNA-containing area. The EUB338

signals served as a control.

2.5 Denaturing Gradient Gel Electrophoresis (DGGE)

At the start and at the end of experiment, biomass was harvested to compare the community

structure (AerAOB, Planctomycetes and total bacteria) while treating high (537 mg N L−1

)

and low (29 mg N L−1

) nitrogen concentrations, respectively. DNA extraction, nested PCR

and DGGE were performed according to Pynaert et al. (2003), based on the primers

CTO189ABf, CTO189Cf, and CTO653r for b-proteobacterial AerAOB; PLA40f and P518r

for Planctomycetes, the bacterial phylum harbouring AnAOB; and GC338 and 518r for all

bacteria. The obtained DGGE patterns were subsequently processed with BioNumerics

software (Applied Maths, Sint-Martens-Latem, Belgium) and similarities were calculated as

the Pearson correlation coefficient.


102

3 Results

3.1 Treatment of high nitrogen levels

Following the influent shift from digested black water (Vlaeminck et al., 2009b) to synthetic

wastewater, the OLAND RBC was operated for 96 days at an influent concentration of

537 mg N L−1

. Over the last 21 days of this period, the nitrogen removal rate was 642 mg N

L−1

d−1

and the nitrogen removal efficiency was 79% (Table 6.2). The contributions of the

different nitrogen pathways were quantified (Fig. 6.2), using the measured dissolved nitrogen

species and assuming that (i) negligible denitrification occurred given the absence of bCOD

in the influent, (ii) nitrogen gas was the product of the removed dissolved species and (iii)

AnAOB produced 0.11 g NO3−-N per g NH4

+-N converted to nitrogen gas. Initially, OLAND

converted 90% of the influent nitrogen, and nitrite and nitrate accumulation were negligible at

high nitrogen levels (Fig. 6.2). Indeed, no NOB could be detected in the biofilm with FISH.

The average DO and free ammonia levels are relatively high at 1.4 mg O2 L−1

and 0.9 mg N

L−1

, respectively (Table 6.2). The AnAOB and AerAOB communities made up 5% and 23%

of the biofilm, respectively (Table 6.3), and were composed of several species (Fig. 6.3).

3.2 Treatment of low nitrogen levels

The nitrogen influent concentration and HRT were gradually decreased over the phases II-Va

while maintaining a constant nitrogen loading rate. Over these changes, the pH and DO levels

remained in the same range, but the total nitrogen removal rate and hence the efficiency

decreased significantly (p<0.05; Table 6.2). The proportion of ammonium oxidized remained

relatively stable but nitrite and nitrate accumulated in the effluent (Table 6.2; Fig. 6.2),

indicating that decreased anammox and increased nitratation were responsible for the

decreased efficiency. Excess nitrate production by NOB significantly increased from period I

to period II, remained constant for periods III and IV, and was followed by a significant

increase in period Va (Table 6.2, Fig. 6.2). The NOB could be identified as Nitrospira, and

composed 6% of the biofilm community at the lowest nitrogen concentration (Table 6.3). The

free ammonia concentration decreased sharply over time whereas DO levels were stable, but

relatively high (Table 6.2).

3.3 Suppression of nitratation at low nitrogen levels

In an attempt to decrease the DO level in phase Vb in order to decrease nitratation and

increase the total nitrogen removal efficiency, discontinuous rotation was introduced. This

Chapter 6

103

resulted in a significant decrease of the oxygen concentration from 1.4 to 1.2 mg O2 L−1

.

Consequently, nitratation decreased with 35% and anammox increased with 32%, restoring

the total nitrogen removal efficiency which was previously obtained in phase IV (Fig. 6.2;

Table 6.2). The lower DO resulted in a total nitrogen effluent concentration of 20 mg N L−1

and a nitrogen removal rate of 383 mg N L−1

d−1

, removing 46% of the influent nitrogen

(Table 6.2). The used RBC set-up did not allow to further decrease of the DO levels since

rotating more intermittently resulted in a strong decrease of the nitritation rate (data not

shown). The relatively stable nitritation indicated few or no influence by the decreasing HRT

(Fig. 6.2; Table 6.2). The decreasing AnAOB activity at lower HRT could be partly

counteracted by a lower DO level and resulting lower nitratation in period Vb (Fig. 6.2),

indicating the influence of DO level on NOB/AnAOB competition rather than a negative

effect of low HRT on anammox

Figure 6.2: Contributions of microbial conversions to reactor nitrogen products: nitrogen gas

production (nitrogen in – nitrogen out) by anoxic ammonium-oxidizing bacteria (AnAOB) from

influent ammonium (upper part) and from nitrite produced by aerobic ammonium oxidizing bacteria

(AerAOB; lower part); nitrate production by AnAOB (11% of ammonium converted by balanced

OLAND); nitrate production by nitrite oxidizing bacteria (NOB: nitrate out – nitrate in – nitrate

production by AnAOB); excess nitrite production by AerAOB (nitrite out – nitrite in); and residual

ammonium (ammonium in – ammonium out).


104

Table 6.2: OLAND rotating biological contactor conditions and performance (average ± standard deviation) over the periods with stepwise decreases of the

ammonium influent concentration and hydraulic residence time (HRT). In periods I-Va, rotation was continuous, whereas this was intermittent in period V

b.

For the eight bottom rows, statistical analyses were performed and the phases that were not significantly different (p>0.05) are indicated with the number of

the similar phase. d: days; h: hours; DO: dissolved oxygen level; prod.: production; cons.: consumption

Period I II III IV Va V

b

Duration (d) 21 41 48 52 31 29

Number of samples (-) 14 18 29 36 23 12

Influent NH4+ level (mg N L

−1) 537 13 278 ± 11 146 ± 21 66 ± 5 29 ± 8 31 1

Influent flow rate (L d−1

) 5.4 0.2 10.5 0.3 20.5 1.5 42.9 2.3 82.6 2.0 83.6 0.7

HRT (h) 16.0 ± 0.5 8.3 ± 0.3 4.2 ± 0.4 2.0 ± 0.1 1.0 ± 0.0 1.0 0.0

N loading rate (mg N L−1

d−1

) 819 ± 30 840 ± 49 832 ± 68 855 ± 56 851 ± 66 840 20

DO level (mg O2 L−1

) 1.4 ± 0.2III,Va

1.2 ± 0.2IV,Vb

1.4 ± 0.2I,Va

1.2 ± 0.1II,Va

1.4 ± 0.4I,III

1.2 0.1II,IV

pH (-) 7.6 ± 0.1 7.5 ± 0.1 7.3 ± 0.2IV,Va,Vb

7.4 ± 0.1III

7.3 ± 0.2III,Vb

7.3 0.0III,Va

Free ammonia (mg N L−1

)* 0.91 1.58II,III

0.40 0.15III,I

0.40 0.17I,II

0.10 0.03 0.04 0.02Vb

0.04 0.01Va

N removal rate (mg N L−1

d−1

) 642 ± 72 565 ± 42 471 ± 88IV

444 ± 84III,Vb

303 ± 75 383 52IV

N removal efficiency (%) 79 ± 9 67 ± 3 58 ± 9 51 ± 8Vb

35 ± 7 46 6IV

NH4+ removal efficiency (%) 94 10 91 3

IV,Va,Vb 72 26 89 4

II,Va,Vb 77 31

II,IV,Vb 91 5

II,IV,Va

NO3−prod./NH4

+ cons. (%)** 12 ± 2 22 ± 2

IV 18 ± 8

IV 21 ± 6

II,III 45 ± 11 32 6

Effluent NH4+ (mg N L

−1) 19 ± 10

II 25 ± 10

I,III 29 ± 12

II 7.4 ± 2.7 3.7 ± 1.5

Vb 3.0 ± 1.0

Va

Effluent NO2- (mg N L

−1) 19 ± 5

III 10 ± 12 16 ± 6

IV,I 14 ± 3

III 5.2 ± 1.3

Vb 5.1 ± 0.4

Va

Effluent NO3- (mg N L

−1) 65 ± 4 59 ± 5 21 ± 9 14 ± 2

Va 15 ± 2

IV 11 ± 0.9

* Calculated from the measured ammonium level, temperature and pH (Anthonisen et al. 1976)

** Values exceeding 11% indicate excess nitrate production by nitrite oxidizing bacteria (NOB)

*** Sum of ammonium, nitrite and nitrate

Chapter 6

105

While treating 29 mg N L−1

, the AnAOB and AerAOB abundances in the biomass were 8 and

25%, respectively (Table 6.3), which was quite comparable to the abundance while treating

high nitrogen concentrations. In the final operation period, part of the biomass was found at

the bottom of the reactor, but neither FISH nor DGGE could detect important differences in

the microbial composition of settled biomass versus biofilm on the discs (Table 6.3, Fig. 6.3),

indicating that the biomass was probably the result of detachment of the biofilm from the

discs. For the AnAOB communities, the DGGE profiles showed only small changes in the

abundant species (88% similarity), while the AerAOB patterns changed more (23%

similarity) (Fig. 6.3). The shift in the most abundant AerAOB could also be observed in the

DDGE profiles of all bacteria.

Figure 6.3: DGGE gels for -proteobacterial aerobic ammonium-oxidizing bacteria (AerAOB) and

Planctomycetes (Plancto), the phylum harbouring anoxic ammonium-oxidizing bacteria (AnAOB).

Biomass samples were taken at the end of the treatment period at 537 mg N L−1

(high N; sample from

the disc biofilm) and at 31 mg N L−1

(low N; sample from the disc biofilm and from settled biomass).

Similarities were calculated as the Pearson correlation coefficient, and plus symbols highlight the three

AerAOB and Planctomycetes bands with the highest intensity, indicating shifts of the most dominant

species.

Table 6.3: Abundances of aerobic and anoxic ammonium-oxidizing bacteria (AerAOB and AnAOB)

and nitrite oxidizing bacteria (NOB) in OLAND biomass, as determined from quantitative fluorescent

in-situ hybridization (FISH). The NOB genera Nitrobacter and Nitrotoga could not be retrieved.

ND: not detected; NM: not measured

Influent (mg L-1

) 537 29

Biomass sample Biofilm Biofilm Settled

AerAOB Nitrosomonas/Nitrosospira (%) 23 ± 18 22 ± 12 30 ± 16

AnAOB Kuenenia/Brocadia (%) 5 ± 5 7 ± 4 8 ± 6

NOB Nitrospira (%) ND 6 ± 5 5 ± 5


106

4 Discussion

4.1 OLAND removal rate and efficiency treating low nitrogen levels

In this study, operation of the OLAND RBC on sewage-like nitrogen concentrations (66 and

29 mg N L−1

) at low HRT (2 and 1 h) resulted in nitrogen removal rates of 383-444 mg N L−1

d−1

, which are reasonably high (Table 6.1). In the energy-positive sewage treatment scheme

(Fig. 1.B), 20-25% of the sewage COD remains in the main line following an advanced

concentration step (Verstraete et al., 2009) as well as about 80% of the sewage nitrogen,

partly derived from returning the digestate to the main line. Assuming raw sewage

compositions of 825 mg COD L−1

and 65 mg N L−1

, the OLAND step would hence receive

165 mg COD L−1

and 52 mg N L−1

. Given the overall required removal efficiencies of 50-

60% of COD and 75% of the nitrogen according to European standards (European

Commision, 1991), the OLAND step should remove an additional 36 mg N L−1

and hence the

desired OLAND removal efficiency should be around 70%. In this study, nitrogen removal

efficiencies during treatment of low nitrogen levels were 46-51%, and hence not sufficiently

high to comply with the required standards. The obtained nitrogen removal percentages were

lower than previously reported for this type of reactors (Pynaert et al., 2003; Schmid et al.,

2003; Pynaert et al., 2004), mainly due to additional nitratation. Also in absolute terms, the

effluent nitrogen concentrations of around 20 mg N L−1

were slightly above the discharge

requirements (> 15 mg N L−1

; European Commision, 1991). Since AerAOB and AnAOB

have high affinities for their nitrogen substrates, with half-saturation constants of 0.05-2.4 mg

N L-1

(Lackner et al., 2008), the microbial capacity should allow further optimization.

4.2 Role of DO levels in suppressing nitratation

The DO levels in the RBC were 1.2-1.4 mg O2 L−1

(Table 6.1) and therefore not low enough

to suppress NOB growth (Bernet et al., 2001; Joss et al., 2009), resulting in a substantial

nitratation (Fig. 6.2). Indeed, NOB could not be detected treating 537 mg N L−1

, but the NOB

genus Nitrospira colonized the biomass at lower nitrogen influent levels, leading to a final

abundance of 5-6%. In contrast to Nitrobacter and Nitrotoga, Nitrospira is typically found in

systems under oxygen-limited conditions, relatively low nitrite levels and moderate

temperature (Lücker et al., 2010). Free ammonia levels between 0.08-0.8 mg N L−1

can inhibit

nitratation (Anthonisen et al., 1976), and could have played a role primarily in period I

(0.9 mg NH3-N L−1

). A DO decrease by 0.2 mg O2 L−1

during phase Vb lowered nitratation

with 35% (Fig. 6.2), demonstrating the link between DO level and NOB activity. It is

Chapter 6

107

anticipated that controlled operation at a sufficiently low DO setpoint (e.g. 0.3 mg O2 L−1

)

will effectively suppress NOB at long term, as demonstrated for treatment of higher strength

OLAND applications (Joss et al., 2009). In an OLAND RBC, it is less straightforward to

control the DO experienced by the biomass than in systems based on active aeration in which

the biomass is either suspended (e.g. Joss et al., 2009) or attached to submerged carrier

material (e.g. Szatkowska et al., 2007). Lower RBC DO levels can generally be obtained by

decreasing the rotor speed (e.g. Meulman et al., 2010), which was not possible on the RBC in

this study, or by increasing the immersion level of the disks. These two actions influence the

biofilm exposure time to atmospheric oxygen and the input turbulence of air in the bulk liquid

by rotation. An additional control of the oxygen level in gas phase of the RBC could further

optimize the microbial balance. In practice, the higher oxygen demand in the presence of

organics (165 mg COD L−1

), will also yield lower bulk DO levels at a similar rotor speed as in

the presence of ammonium only.

4.3 OLAND operation at low HRT

Compared to described OLAND systems, the applied HRT in this study were very low (Table

6.1) but this did not seem to have an adverse effect on AerAOB or AnAOB activity. It is not

clear whether an expected higher biomass washout at lower HRT could have been responsible

for shifts in the microbial community. The sequentially decreasing nitrogen concentrations in

the reactor (Table 6.2) possibly had a stronger influence on establishing an adapted OLAND

microbiome which was likely more oligotrophic.

Compared to treatment at high HRT (e.g. 24 h), applying lower HRT (e.g. 1 h) at high

volumetric loading rates will have an influence on the design parameters, depending on the

type of OLAND reactor. In suspended growth systems, biomass retention is based on settling.

In case of an external settler, a lower HRT will result in a higher sludge surface load, and

hence needs a relative increase of the settler volume compared to the reactor volume to

maintain the same sludge residence time. In case of a sequencing batch reactor, decreased

HRT will require an increased minimum biomass settling velocity from 1 m h−1

(Joss et al.,

2009) to 24 m h−1

to maintain an acceptably low ratio of settling to reaction time, and

therefore will require granules rather than flocs (Chapter 2; Vlaeminck et al., 2010). In

contrast to suspended growth configurations, HRT in biofilm-based systems is expected to

have only a minor influence on the biomass retention, allowing for compact reactors.


108

4.4 Implementation of OLAND in the main stream

OLAND treatment of pretreated sewage should achieve sufficiently high nitrogen removal

rates and efficiencies at low hydraulic residence times and nitrogen concentrations at minimal

energy requirements, given the overall aim of energy-positive sewage treatment. Overall,

several decision factors will determine the desirable reactor technology. Passive versus active

aeration will determine energy requirements, but also the ease of controlling the microbial

activity balance, and suspended versus attached biomass growth will determine the ease of

maintaining a high biomass retention at low HRT.

The next research challenges for the implementation of OLAND in the main stream of the

sewage treatment relate firstly to a decrease of the process temperatures from the maximum

summer temperature (25°C) over the average year temperature (17°C) to the minimum winter

temperature (8°C) (Breda, NL; Mollen, oral communication). This will elucidate whether

OLAND requires a distinct oligotrophic and cold-tolerant autotrophic community and

physiology. Secondly, the continued OLAND performance will have to be shown in the

presence of moderate bCOD levels (90-240 mg L−1

), with COD/N ratios between 2.4 and 3.

The latter will likely facilitate DO control at low DO levels due to heterotrophic aerobic

activity. However, also competition for nitrite will take place between heterotrophic

denitrification and anammox. These processes have however been demonstrated already to

successfully co-exist at a COD/N of 2.2 (Desloover et al., 2011). It is anticipated that due to

future dilution preventions (Henze, 1997; Brombach et al., 2005), higher nitrogen sewage

levels together with the higher sewage temperature will facilitate OLAND treatment in the

main stream. Finally, high OLAND performance will have to be shown under realistic

temporal variations in sewage composition and in performance of the preceding advanced

concentration.

5 Acknowledgements

H.D.C. received a PhD grant from the Institute for the Promotion of Innovation by Science

and Technology in Flanders (IWT-Vlaanderen, SB-81068) and S.E.V. was supported as a

postdoctoral fellow from the Research Foundation Flanders (FWO-Vlaanderen). The authors

gratefully thank Hans Mollen (Waterschap Brabantse Delta, NL) for sharing temperature data,

Siska Maertens for molecular analyses, and Nico Boon, Tom Hennebel, Jan Arends, Yu

Zhang, Sebastià Puig and Samik Bagchi for inspiring scientific discussions.

Chapter 7

109

Chapter 7:

Cold OLAND on pretreated sewage:

feasibility demonstration at lab-scale

Abstract

Energy-positive sewage treatment can be achieved by implementation of oxygen-limited

autotrophic nitrification/denitrification (OLAND) in the main water line, as the latter does not

require organic carbon and therefore allows maximum energy recovery through anaerobic

digestion of organics. To test the feasibility of mainstream OLAND, the effect of a gradual

temperature decrease from 29°C to 15°C and a COD/N increase from 0 to 2 was tested in an

OLAND rotating biological contactor (RBC) operating at 55-60 mg NH4+-N L

-1 and a

hydraulic retention time of 1 hour. Moreover, the effect of the operational conditions and

feeding strategies on the reactor cycle balances, including NO/N2O emissions were studied in

detail. At 15°C (9 months) high anoxic and aerobic ammonium oxidation activities were

maintained. However, nitratation (NOB activity) occurred at temperatures below 20°C.

Operation at COD/N ratios of 2 and 15°C (2 months) still allowed for high nitrogen removal

rates of 0.5 g N L-1

d-1

, which are in the same range as high temperature applications. The

main challenge to allow high removal efficiencies in this application was the suppression of

NOB at low free ammonia (< 0.25 mg N L-1), low free nitrous acid (<0.9 μg N L

-1) and higher

DO levels (3-4 mg O2 L-1

). This study showed that high NO levels had the potential to favor

anammox above NOB activity. It should be evaluated if the increased NO/N2O emission can

be compensated with a decreased energy consumption to justify OLAND mainstream

treatment.

Chapter redrafted after: De Clippeleir H., Vlaeminck S.E., De Wilde F., Daeninck K.,

Mosquera M., Boeckx P., Verstraete W. and Boon N. Cold one-stage partial

nitritation/anammox on pretreated sewage: feasibility demonstration at lab-scale. Submitted.

Cold OLAND on preteated sewage: feasibility demonstration on lab-scale

110

1 Introduction

Currently, around 40 full-scale 1 stage partial nitritation/anammox plants are implemented to

treat highly loaded nitrogen streams devoid in carbon (Chapter 1). This process, known under

the acronyms OLAND (Kuai and Verstraete, 1998), DEMON (Wett, 2006), CANON (Third

et al., 2001) etc, showed highly efficient and stable performance when treating digestates

from sewage sludge treatment plants and industrial wastewaters (Wett, 2006; Abma et al.,

2010; Jeanningros et al., 2010). From an energy point of view, the implementation of the

OLAND process for the treatment of sewage sludge digestate decreased the net energy

consumption of a municipal wastewater treatment plant (WWTP) with 50% (Siegrist et al.,

2008). Moreover, when co-digestion of kitchen waste was applied, an energy neutral WWTP

was achieved (Wett et al., 2007). To fully recover the potential energy present in wastewater,

a ‘ZeroWasteWater’ concept was proposed which replaces the conventional activated sludge

system by a highly loaded activated sludge step (A-step), bringing as much as organic carbon

(COD) as possible to the solid fraction, and a second biological step (B-step) removing the

residual nitrogen and COD with a minimal energy demand (Verstraete and Vlaeminck, 2011).

Subsequently, energy is recovered via anaerobic digestion of the primary and secondary

sludge. For the B-step in the main line, OLAND would potentially be the best choice as this

process can work at low COD/N ratio, allowing maximum recovery of COD in the A-step.

Moreover, it was calculated that if OLAND is implemented in the main water treatment line

and a maximum COD recovery takes place in the A-step, a net energy gain of the WWTP of

10 Wh inhabitant equivalent (IE)-1

d-1

is feasible (Chapter 4).

To allow this energy-positive sewage treatment, OLAND has to face some challenges

compared to the treatment of highly loaded nitrogen streams (> 250 mg N L-1

). A first

difference is the lower nitrogen concentration to be removed by OLAND. Domestic

wastewater after advanced concentration will still contain around 30-100 mg N L-1

and 113-

300 mg COD L-1

(Tchobanoglous et al., 2003; Henze et al., 2008). High nitrogen conversion

rate (around 400 mg N L-1

d-1

) by the OLAND process can be obtained at nitrogen

concentrations of 30-60 mg N L-1

and at low hydraulic retention times (HRT) of 1-2 hours

(Chapter 6). A second challenge is the low temperature at which OLAND should be operated

(10-15°C compared to 34°C). Several studies already described the effect of temperature on

the activity of the separate microbial groups (Dosta et al., 2008; Guo et al., 2010; Hendrickx

et al., 2012). However, limited knowledge exists about the microbial balances of these

Chapter 7

111

3 groups under OLAND conditions at low temperature (< 20°C). At temperatures around

15°C, maintaining the balance between nitrite-oxidizing bacteria (NOB) and anoxic

ammonium-oxidizing bacteria (AnAOB) and the balance between NOB and aerobic

ammonium-oxidizing bacteria (AerAOB) will get more challenging since the growth rate of

NOB will become higher than the growth rate of AerAOB (Hellinga et al., 1998). Therefore,

it will not be possible to wash out NOB based on overall or even selective sludge retention.

The third and main challenge in this application will therefore be the suppression of NOB at

temperature ranges of 10-20°C and at nitrogen concentration ranges of 30-60 mg N L-1

(low

free ammonia and low nitrous acid). A final fourth challenge will include the higher input of

organics at moderate levels of 90-240 mg bCOD L-1

in the wastewater. Depending on the raw

sewage strength, COD/N ratios between 2 and 3 are expected after the concentration step,

which is on the edge of the described limit for successful OLAND (Lackner et al., 2008). The

presence of organics could result in an extra competition of heterotrophic denitrifiers with

AerAOB for oxygen or with AnAOB for nitrite.

In this study an OLAND RBC at 29°C was gradually adapted over 24, 22 and 17°C to 15°C

under synthetic wastewater conditions (60 mg N L-1

, COD/N of 0). Additionally, the COD/N

ratio of the influent was increased to 2 by supplementing NH4+ to diluted sewage to simulate

pretreated sewage. The effect of the operational conditions and feeding strategies on the

reactor cycle balances, including gas emissions and microbial activities were studied in detail.

An alternative strategy to inhibit NOB activity and as a consequence increase AnAOB

activity at low temperatures was proposed.


2.1 OLAND rotating biological contactor (RBC)

The lab-scale RBC described in Chapter 6 was further optimized at 29°C by an increase in the

influent nitrogen concentration from 30 to 60 mg N L-1

and a limitation of the oxygen input

through the atmosphere by covering the reactor before this test was started. The reactor was

based on an air washer LW14 (Venta, Weingarten, Germany) with a rotor consisting of 40

discs interspaced at 3 mm, resulting in a disc contact surface of 1.32 m2. The reactor had a

liquid volume of 2.5 L, immersing the discs for 55%. The latter was varied over the time of

the experiment. The reactor was placed in a temperature-controlled room. The DO


112

concentration was not directly controlled. For continuous rotation the rotation speed was fixed

at 3 rpm.

2.2 RBC operation

The RBC was fed with synthetic wastewater during phases I to VII. From phase VIII

onwards, the COD/N was gradually increased (Phase VIII-X) to 2 (phase XI-XIII). The

synthetic influent of an OLAND RBC, consisted of (NH4)2SO4 (55-60 mg N L-1

), NaHCO3

(16 mg NaHCO3 mg-1

N) and KH2PO4 (10 mg P L-1

). Pretreated sewage was simulated by

diluting raw sewage of the communal WWTP of Gent, Belgium (Aquafin). The raw

wastewater contained 23-46 mg NH4+-N L

-1, 0.2-0.4 mg NO2

--N L

-1, 0.4-2.7 mg NO3

--N L

-1,

23-46 mg Kjeldahl-N L-1

, 3.8-3.9 mg PO43-

-P L-1

, 26-27 mg SO42-

-S L-1

, 141-303 mg CODtot

L-1

and 74-145 mg CODsol L-1

. The raw sewage was diluted by a factor 2-3 to obtain COD

values around 110 mg CODtot L-1

and by addition of (NH4)2SO4 to obtain final COD/N values

around 2. The reactor was fed in a semi-continuous mode: 2 periods of around 10 minutes per

hour for phases I-XI, 1 period of 20 minutes per hour for phases XII and XIII. Reactor pH,

DO and temperature were daily monitored and influent and effluent samples were taken at

least thrice a week for ammonium, nitrite, nitrate and COD analyses.

2.3 Detection of AerAOB, NOB and AnAOB with FISH and qPCR

For NOB and AnAOB, a first genus screening among the most commonly present organisms

was performed by fluorescent in situ hybridization (FISH) on biomass of day 1 (high

temperature) and day 435 (low temperature and COD presence). A paraformaldehyde (4%)

solution was used for biofilm fixation, and FISH was performed according to Amann et al.

(1990). The Sca1309 and Amx820 probes were used for the detection of Cand. Scalindua and

Cand. Kuenenia & Brocadia, respectively, and the NIT3 and Ntspa662 probes and their

competitors for Nitrobacter and Nitrospira, respectively. This showed the absence of

Nitrobacter and Scalindua (Table S7.1). Biomass samples (approx. 5 g) for nucleic acid

analysis were taken from the OLAND RBC at days 1, 60, 174, 202, 306, 385, 399 and 413 of

the operation. DNA was extracted using FastDNA® SPIN Kit for Soil (MP Biomedicals,

LLC), according to the manufacturer’s instructions. The obtained DNA was purified with the

Wizard® DNA Clean-up System (Promega, USA) and its final concentration was measured

spectrophotometrically using a NanoDrop ND-1000 spectrophotometer (Nanodrop

Technologies). The SYBR Green assay (Power SyBr Green, Applied Biosystems) was used to

quantify the 16S rRNA of AnAOB and Nitrospira sp. and the functional amoA gene for

Chapter 7

113

AerAOB. The primers for quantitative polymerase chain reactions (qPCR) for detection of

AerAOB, NOB and AnAOB were amoA-1F – amoA-2R, NSR1113f-NSR1264r and

Amx818f – Amx1066r, respectively. For bacterial amoA gene, PCR conditions were: 40

cycles of 94°C for 1 min, 55°C for 1 min and 60°C for 2 min. For the amplification of

Nitrospira sp 16S rRNA gene 40 cycles of 95°C for 1 min, 50°C for 1 min and 60°C for 1

min were used while for AnAOB 16S rRNA the PCR temperature program was performed by

40 cycles of 15 seg at 94°C and 1 min at 60°C. Plasmid DNAs carrying Nitrospira and

AnAOB 16SrRNA gene and AerAOB functional AmoA gene, respectively, were used as

standards for qPCR. All the amplification reactions had a high correlation coefficient (R2>

0.98) and slopes between -3.0 and -3.3. A paraformaldehyde (4%) solution was used for

biofilm fixation, and FISH was performed according to Amann et al. (1990). The Bfu613

probe was used for the detection Brocadia fulgida (Kartal et al., 2008) and EUB I,II,II for

detection of all bacteria.

2.4 Detailed reactor cycle balances

For the measurements of the total nitrogen balance, including the NO and N2O emissions, the

OLAND RBC was placed in a vessel (34 L) which had a small opening at the top (5 cm2). In

this vessel a constant upward air flow (around 1 m s-1

) was generated to allow calculations of

emission rates. On the top of the vessel (air outlet), the NO and N2O concentration was

measured, off- and online, respectively. In the water phase, ammonium, nitrite, nitrate,

hydroxylamine (NH2OH), N2O and COD concentrations were measured. Moreover, DO

concentration and pH values were monitored. The air flow was measured with Testo 425 hand

probe (Testo, Ternat, Belgium).


Ammonium (Nessler method) was determined according to standard methods (Greenberg et

al., 1992). Nitrite and nitrate were determined on a 761 compact ion chromatograph equipped

with a conductivity detector (Metrohm, Zofingen, Switzerland). Hydroxylamine was

measured spectrophotometrically (Frear and Burrell, 1955). The chemical oxygen demand

(COD) was determined with NANOCOLOR® COD 1500 en NANOCOLOR® COD 160 kits

(Macherey-Nagel, Düren, Germany). The volumetric nitrogen conversion rates by AerAOB,

NOB and AnAOB were calculated based on the measured influent and effluent compositions

and the described stoichiometries (Vlaeminck et al., 2012). DO and pH were measured with

respectively, a HQ30d DO meter (Hach Lange, Düsseldorf, Germany) and an electrode


114

installed on a C833 meter (Consort, Turnhout, Belgium). Gaseous N2O concentrations were

measured online at a time interval of 3 minutes with a photo-acoustic infrared multi-gas

monitor (Brüel & Kjær, Model 1302, Nærem, Denmark). Gas grab samples were taken during

the detailed cycle balance tests for NO detection using Eco Physics CLD 77 AM (Eco Physics

AG, Duernten, Switzerland), which is based on the principle of chemiluminescence. For

dissolved N2O measurements, a 1 mL filtered (0.45 μm) sample was brought into a 7 mL

vacutainer (-900 hPa) and measured afterwards by pressure adjustment with He and

immediate injection at 21°C in a gas chromatograph equipped with an electron capture

detector (Shimadzu GC-14B, Japan).

3 Results

3.1 Effect of temperature decrease

During the reference period (29°C), a well-balanced OLAND performance (Fig. 7.1, Table

7.1) was reached with minimal nitrite accumulation (2%) and minimal nitrate production

(7%). This was reflected in an AerAOB/AnAOB activity ratio of 0.6 (Table 7.1, Phase I). The

total nitrogen removal rate was on average 470 mg N L-1

d-1

and the total nitrogen removal

efficiency was 54%.

Decreasing the temperature from 29 to 24°C and further to 22°C over the following 40 days,

did not result in any significant changes of the operational conditions (Table 7.1, Phases I-III),

performance of the reactor (Fig. 7.1) or abundance of the bacterial groups (qPCR, Fig. S7.1).

However at 17°C, a decrease in total nitrogen removal efficiency was observed (Table 7.1,

Phase IV). As the ammonium consumption rate went down (from 501 23 to 383 80 mg N

L-1

d-1

) and the effluent nitrite and nitrate levels remained stable (Fig. 7.1), a higher relative

nitrite and nitrate production indicated an imbalance between the AerAOB and the AnAOB.

Moreover, NOB activity was for the first time detected while no difference in free ammonia

(FA) or free nitrous acid (FNA) suppression on NOB was observed (Table 7.1, Phase IV).

Moreover, no significant differences in abundance of NOB, AerAOB and AnAOB could be

detected with qPCR (Fig. S7.1). However, DO concentrations started to increase during that

period from 1.4 to 1.7 mg O2 L-1

. To counteract the decrease in ammonium removal

efficiency the immersion level was lowered to 55% to increase the availability of oxygen.

Consequently the volumetric loading rate increased (factor 1.7) due to the decrease in reactor

volume (day 210, Fig. 7.1). This action allowed higher ammonium removal efficiencies due

Chapter 7

115

to higher AerAOB activities (factor 3). AnAOB activity increased with a similar factor as the

volumetric loading rate (1.8 compared to 1.7) consequently resulting in an increased

imbalance between these two groups of bacteria (Table 7.1, Phase V). Moreover, although the

FNA increased with a factor 2, the NOB activity increased with a factor 7, resulting in a

relative nitrate production of 30% (Table 7.1, Phase V). As NOB activity prevented good total

nitrogen removal efficiencies, the immersion level was increased again to 78% (day 263, Fig.

7.1). This resulted indeed in a lower NOB activity (Table 7.1, Phase VI). However, also the

AerAOB activity decreased with the same factor, due to the lower availability of atmospheric

oxygen. Therefore, the reactor was subsequently operated at this low immersion level (55%)

to allow sufficient aerobic ammonium conversion. The latter allowed a stable removal

efficiency of 42%. The AnAOB activity gradually increased to a stable anoxic ammonium

conversion rate of 529 mg N L-1

d-1

. During the synthetic phase, no changes in AerAOB,

AnAOB and NOB abundance were measured with qPCR (Fig. S7.1). The effluent quality was

however not optimal as still high nitrite (around 15 mg N L-1

) and nitrate (around 13 mg N

L-1

) levels were detected.


116

Figure 7.1: Phases I-VII: Effect of temperature decrease on the volumetric rates (top) and nitrogen

concentrations (bottom).

Chapter 7

117

Table 7.1: Effect of temperature decrease on the operational conditions and performance of OLAND RBC reactor. DO: dissolved oxygen; HRT: hydraulic

retention time; FA: free ammonia; FNA: free nitrous acid; cons: consumption; tot: total

Phase I II III IV V VI VII

Period (d) 1-21 22-35 36-61 62-210 210-263 263-274 275-306

Immersion level (%) 78 78 78 78 55 78 55

Temperature (°C) 29 ± 2 24 ± 1 22 ± 0.6 17 ± 1.2 16 ± 0.9 15 ± 0.8 14 ± 0.4

Operational conditions:

DO (mg O2 L-1

) 1.1 ± 0.2 1.3 ± 0.2 1.4 ± 0.1 1.7 ± 0.3 2.8 ± 0.4 2.4 ± 0.2 3.1 ± 0.2

pH (-) 7.5 ± 0.1 7.5 ± 0.1 7.5 ± 0.1 7.6 ± 0.1 7.7 ± 0.1 7.7 ± 0.1 7.8 ± 0.1

HRT (h) 1.85 ± 0.04 1.84 ± 0.09 1.73 ± 0.04 1.86 ± 0.11 1.09 ± 0.02 1.57 ± 0.02 1.09 ± 0.02

FA (mg N L-1

) 0.35 ± 0.18 0.36 ± 0.18 0.34 ± 0.14 0.36 ± 0.13 0.25 ± 0.16 0.33 ± 0.17 0.13 ± 0.04

FNA (μg N L-1

) 0.3 ± 0.1 0.3 ± 0.2 0.4 ± 0.2 0.4 ± 0.1 0.9 ± 0.4 0.6 ± 0.1 0.9 ± 0.2

Performance:

Total N removal efficiency (%) 54 ± 5 52 ± 5 49 ± 9 34 ± 9 36 ± 9 36 ± 9 42 ± 4

Relative NO3- prod (% of NH4

+ cons*) 7 ± 1 7 ± 1 7 ± 1 14 ± 6 18 ± 9 16 ± 3 21 ± 4


+ cons) 2 ± 4 3 ± 4 5 ± 5 15 ± 5 30 ± 8 26 ± 6 31 ± 5

AerAOB activity (mg NH4+-N L

-1 d

-1) 267 ± 38 267 ± 49 260 ± 52 260 ± 53 811 ± 229 460 ± 44 986 ± 71

NOB activity (mg NO2—

N L-1

d-1

) 0 ± 0 0 ± 0 0 ± 0 9 ± 12 60 ± 94 20 ± 5 85 ± 25

AnAOB activity (mg Ntot L-1

d-1

) 412 ± 38 403 ± 37 368 ± 76 248 ± 67 448 ± 117 305 ± 74 529 ± 75

*NH4+ consumption is corrected for nitrite accumulation


118

3.2 Effect of COD/N increase

The synthetic feed was gradually changed into pretreated sewage by diluting raw sewage and

adding additional nitrogen to obtain a certain COD/N ratio. During the first 3 weeks of this

period (Fig. 7.2), the COD/N ratio was gradually increased from 0.5 to 2. Due to the short

adaptation periods (1 week per COD/N regime), the performance was unstable (Fig. 7.2,

Table 7.2, phase VIII-XI). Compared to the end of the synthetic period (phase VII), operation

at a COD/N ratio of 2 (phase XI) resulted in a sharp decrease in nitrite accumulation (Fig. 7.2)

and an increase in the ammonium and nitrate levels. This indicated increased NOB activity

(factor 4), decreased AerAOB (factor 3) and decreased AnAOB (factor 2) activity (Table 7.1

and 7.2). To allow higher nitrogen removal rates, the HRT was increased from 0.94 to 1.1 h,

by decreasing the influent flow rate. Moreover, the feeding regime was changed from 2 pulses

of 10 minutes in 1 hour to 1 period of 20 minutes per hour. These actions did not significantly

decrease the effluent nitrogen concentration (Fig. 7.2) and did not influence the microbial

activities (Table 7.2, phase XII). Therefore the loading rate was again increased to the levels

before phase XII. However the single-pulse feeding was maintained. This resulted in high

ammonium removal efficiencies and therefore low ammonium effluent concentration around

dischargeable level (4 ± 1 mg NH4+-N L

-1; Fig. 7.2). Nitrate and nitrite accumulation were not

counteracted by denitrification as only 0.02 mg COD L-1

d-1

was removed. Therefore nitrite

and nitrate levels were still too high to allow effluent discharge. The total nitrogen removal

efficiency (42%) and rate (549 ± 83 mg N L-1

d-1

) at COD/N ratios of 2 was similar as during

the synthetic period (phase VII). Compared to the reference period at 29°C, 22% of the

removal efficiency was lost, while the total nitrogen removal rate did not changed

significantly (470 ± 43 versus 549 ± 83 mg N L-1

d-1

at high and low temperature,

respectively).

Chapter 7

119

Figure 7.2: Phases VIII-XIII: Effect of COD/N increase on the volumetric rates (top) and nitrogen

concentrations (bottom) at 15°C and an immersion level of 55%.


120

Table 7.2: Effect of COD/N increase on the operational conditions and performance of OLAND RBC reactor at 15°C and an immersion level of 55%.. DO:

dissolved oxygen; HRT: hydraulic retention time; FA: free ammonia; FNA: free nitrous acid; cons: consumption; tot: total

Phase VIII IX X XI XII XIII

Period (d) 355-361 362-369 370-374 375-406 407-421 422-

Immersion level (%) 55 55 55 55 55 55

COD/N 0.5 1 1.5 2 2 2

Feeding regime (pulses h-1

) 2 2 2 2 1 1

Operational conditions:

DO (mg O2 L-1

) 2.9 ± 0.3 2.5 ± 0.6 2.4 ± 0.3 3.0 ± 0.7 3.6 ± 0.3 3.2 ± 0.3

pH (-) 7.8 ± 0.02 7.7 ± 0.1 7.6 ± 0.02 7.6 ± 0.1 7.6 ± 0.2 7.6 ± 0.1

HRT (h) 1.06 ± 0.11 1.03 ± 0.02 0.92 ± 0.02 0.94 ± 0.05 1.10 ± 0.05 1.06 ± 0.2

FA (mg N L-1

) 0.10 ± 0.05 0.04 ± 0.05 0.15 ± 0.05 0.21 ± 0.10 0.23 ± 0.12 0.04 ± 0.02

FNA (μg N L-1

) 0.4 ± 0.1 0.2 ± 0.2 0.2 ± 0.01 0.3 ± 0.1 0.2 ± 0.1 0.6 ± 0.2

Performance:

Total N removal efficiency (%) 36 ± 5 45 ± 18 23 ± 3 28 ± 6 23 ± 13 42 ± 3


+ cons*) 42 ± 5 43 ± 12 63 ± 2 50 ± 6 62 ± 18 46 ± 6


+ cons) 20 ± 4 10 ± 10 5 ± 1 8 ± 3 7 ± 4 13 ± 6


-1 d

-1) 592 ± 15 446 ± 31 238 ± 28 352 ± 73 289 ± 138 600 ± 204

NOB activity (mg NO2--N L

-1 d

-1) 257 ± 19 294 ± 81 465 ± 60 352 ± 84 427 ± 115 394 ± 76


d-1

) 385 ± 86 452 ± 205 262 ± 39 355 ± 73 281 ± 159 481 ± 73

*NH4+ consumption is corrected for nitrite accumulation

Chapter 7

121

3.3 Nitratation and NO/N2O emissions

At the end of the synthetic phase (Phase VII) and the end of the experiment (Phase XIII) the

total nitrogen balance of the reactor was measured. A total nitrogen balance was obtained by

measuring all nitrogen species (NH4+, NO2

-, NO3

-, NH2OH, N2O) in the liquid phase and N2O

and NO in the gas phase. A constant air flow, diluting the emitted N2O and NO concentrations

was created over the reactor to measure gas fluxes over time. The effect of the loading rate,

feeding pattern and concentration of nitrite and ammonium on the total nitrogen balance in the

reactor were tested (Table 7.3). NH2OH measurement showed low concentrations (< 0.2 mg

N L-1

) in all tests, making it difficult to link the profiles with the N2O emission.

Lowering the loading rate by increasing the HRT (Test B, Table 7.3) increased the DO values

and allowed higher DO fluctuations over time at synthetic conditions. Moreover NOB activity

increased significantly resulting in lower total nitrogen removal efficiencies and high levels of

nitrate in the effluent (Table 7.3, Test B). The relative N2O emissions did not change and

were relatively high (6% of N load). However, the concentration of N2O in the liquid and in

the gas phase decreased with a factor 2 (Table 7.3).

When pretreated sewage was fed to the reactor, the OLAND RBC was operated at lower

nitrite concentration, while similar ammonium and nitrate concentrations were obtained

(Table 7.3, Test C). The latter however did not result in lower N2O emission rates. When the

feeding regime was changed to a more continuous-like operation (4 pulses h-1

), the N2O

emission increased significantly, while NO emission remained constant (Table 7.3, test D).

Due to the lower ammonium removal efficiency (65 compared to 81%), but similar relative

nitrite and nitrate accumulation rate, the total nitrogen removal efficiency decreased.

When a nitrite pulse was added just after feeding, about 20 mg NO2--N L

-1 was obtained in the

reactor. This did increase the NO and N2O emissions significantly (p<0.05) compared to the

same feeding pattern (Table 7.3, Test C-E). Although similar constant total nitrogen removal

efficiencies were obtained during this operation, a significant (p<0.05) decrease in the relative

nitrate production was observed. The latter was mainly caused by a global increase in

AnAOB activity. In the last test (F), the influent ammonium concentration was doubled,

leading to higher ammonium and also free ammonia concentrations (1 ± 0.4 mg N L-1

compared to 0.1 ± 0.4 mg N L-1

). Due to overloading of the system, the total nitrogen removal


122

efficiency decreased. However, at these conditions a lower relative nitrate production was

obtained, due to a decrease in NOB and increase in AnAOB activity (Table 7.3, Test F).

Together with this, increased NO and N2O emissions were observed. As the influence of the

nitrogen loading and DO concentration could be considered minor in this test range (Figure

S7.2), these tests show a relation between increased NO emissions and decreased relative

nitrate productions (Table 7.3).

When the activity during the feeding cycle was studied in more detail, it could be concluded

that the highest nitrogen conversion rates took place during the feeding period (Fig. 7.3). As

the HRT is only 1 hour, the reactor volume is exchanged in 20 minutes. During this phase,

ammonium increased, while nitrite and nitrate concentrations decreased due to dilution (Fig.

S7.3-S7.5). The NOB/AnAOB ratio was around 1, which means that NOB were able to take

twice as much nitrite than AnAOB did, as the latter also consumed ammonium (Fig. 7.3).

After the feeding period, a lag phase of the ammonium increase was observed, because the

reactor liquid was not homogenously mixed yet. After mixing (10 minutes after feeding) was

established, a N2O peak was reached during every test (Fig. S7.3-S7.5). At this point, during

the reference period with pretreated sewage (Test C) total activity decreased and a very low

NOB activity was observed (Fig. 7.3). Moreover, the NOB/AnAOB ratio decreased to 0.4

(Test C, Fig. 7.3), which means that during these conditions nitrite consumption by AnAOB

was higher than nitrite consumption by NOB. The increased relative AnAOB activity was

more pronounced when a higher NO and N2O peak were present (Test E). The latter was

caused by an increased nitrite concentration in the reactor. When N2O concentration started to

decrease again (last 20 minutes of feeding regime), nitrite consumption by NOB was again

higher than the nitrite consumption by AnAOB (Fig. 7.3).

Chapter 7

123

Table 7.3: Operational parameters and nitrogen conversion rates during the 6 different RBC operations which differ from feeding composition and feeding

regime (volume 2.5 L and 50 % immersion of the discs, day 307-309 for synthetic feed, days 424-431 for pretreated sewage).

Reactor phase VII: synthetic XIII: pretreated sewage

Test A° B C° D E- F

Additive - - - - NO2- NH4

+

Feeding regime (pulses/h) 2 2 1 4 1 1

Total N loading rate (mg N L-1

d-1

) 1169 585 1340 1554 1737 2718

Temperature water (°C) 15 ± 0.3 16 ± 0.2* 14 ± 0.4 15 ± 0.1* 16 ± 0.1* 15 ± 0.4

DO (mg O2 L-1

) 2.9 ± 0.1 3.7 ± 0.6* 4.0 ± 0.1 3.2 ± 0.1* 3.3 ± 0.1* 3.2 ± 0.1*

pH 7.6 ± 0.06 7.6 ± 0.05 7.6 ± 0.04 7.6 ± 0.01 7.6 ± 0.02 7.8 ± 0.02*

Ammonium out (mg N L-1

) 9 ± 1 1.4 ± 1* 11 ± 3 19 ± 3* 12 ± 1 58 ± 4*

Nitrite out (mg N L-1

) 14 ± 2 13 ± 1 6 ± 1 6 ± 0.4 18 ± 2* 9 ± 0.3*

Nitrate out (mg N L-1

) 17 ± 3 37 ± 6* 18 ± 2 16 ± 1* 18 ± 0.4 20 ± 0.4

NH4+ oxidation rate (mg N L

-1 d

-1) 895 ± 22 509 ± 2* 1051 ± 73 957 ± 89 1053 ± 16 1285 ± 93*

Relative nitrite accumulation (%) 25 ± 3 20 ± 1* 14 ± 3 15 ± 1 8 ± 4* 15 ± 1

Relative nitrate production (%) 36 ± 8 76 ±6* 48 ± 1 47 ± 3 42 ± 2* 34 ± 3*

Total efficiency (%) 38 ± 4 17 ± 4* 35 ± 3 28 ± 4* 32 ± 2 27 ± 4*


-1 d

-1) 658 ± 88 469 ± 17* 827 ± 44 781 ± 57 795 ± 30 938 ± 46*

NOB activity (mg NO2--N L

-1 d

-1) 174 ± 59 299 ± 28* 375 ± 38 342 ± 24* 362 ± 13 277 ± 18*


d-1

) 205 ± 38 49 ± 13* 234 ± 20 218 ± 29 263 ± 15* 354 ± 49*

N2O in liquid (μg N L-1

) 64 ± 46 30 ± 22* 78 ± 12 104 ± 29* 61 ± 13 74 ± 4

NO emission (mg N d-1

) 0.53 ± 0.03 n.d. 0.66 ± 0.06 0.74 ± 0.08 1.65 ± 0.18* 0.82 ± 0.1*

N2O emission (mg N d-1

) 151 ± 28 93 ± 23* 170 ± 19 179 ± 6* 274 ± 37* 202 ± 18*

% N2O emission on loading 5.1 ± 1.0 6.4 ± 1.6* 5.0 ± 0.6 4.5 ± 0.2* 6.2 ± 0.8* 3.0 ± 0.3*

°Reference period for synthetic and pretreated sewage

*Significant differences (p < 0.05) compared to reference period


124

Figure 7.3: Detailed NO/N2O monitoring during the reference test (Test C, Table 7.3) and when

nitrite was pulsed (Test E, Table 7.3) and effect on AerAOB, AnAOB and NOB activity during the

different phases of the feeding cycle. Significant differences in AerAOB, AnAOB, NOB and NO/N2O

concentration compared to the reference period are indicated with *, °, “ and +, respectively.

4 Discussion

4.1 Effect of temperature decrease

Average temperatures of sewage in west European region are around 17°C, with a minimum

of 8°C and a maximum of 29°C (Mollen, personal communication). Therefore, the

temperature of the OLAND RBC was decreased from 29°C to 15°C. In contrast to the optimal

microbial balance at temperatures > 20°C, excess nitrite and nitrate formation was observed at

lower temperatures. Improved operational conditions (O2 availability) resulted in similar

nitrogen conversion rates for AerAOB and AnAOB at lower temperature (< 20°C) compared

to the reference period at 29°C. The gradual adaptation of the nitrogen converting community

to low temperatures probably attributed to the lower temperature dependence of AerAOB and

AnAOB activity compared to the temperature shocks described in literature (Dosta et al.,

2008; Guo et al., 2010). Similar long-term effect of temperature on AerAOB activity (Guo et

Chapter 7

125

al., 2010) and AnAOB activity (Hu et al., 2011; Hendrickx et al., 2012) were observed before.

Due to the higher DO concentration at lower temperatures, the oxygen penetration depth

possibly increased causing a decrease in AnAOB activity. On the other hand, higher oxygen

inputs were needed at lower temperatures to obtain the same AerAOB activity as at high

temperature. The combination of these two factors could have been responsible for the

increased nitrite accumulation from phase IV onwards. Therefore, at lower temperature the

OLAND performance will be limited by AerAOB activity as their activity guarantees anoxic

zones in the biofilm (Vazquez-Padin et al., 2011).

Increased AerAOB activities were obtained at high DO levels (3 times higher than at 29°C).

This was on one hand caused by a better solubility of oxygen at lower temperatures and on

the other hand by a decrease of the immersion level from 78 to 55%. Although the changes in

immersion level did not always resulted in a significant DO change (phase IV to V; phase V

to VI), the oxygen availability through contact with the atmosphere was changed drastically.

This suggests that oxygen transfer through atmospheric oxygen is more important in this

system compared to transfer from dissolved oxygen. Although oxygen concentrations in

OLAND systems at high temperature conditions are controlled at levels below 1 mg O2 L-1

to

avoid nitrate oxidation by NOB, at low temperatures 2-4 mg O2 L-1

is needed to allow

sufficient AerAOB activity (Vazquez-Padin et al., 2011). As nitrite accumulated in the

OLAND RBC, oxygen input was probably too high to allow a balanced performance between

nitrite production and consumption. Therefore, in practice a bulk DO control system is

advisable to obtain a better removal efficiency.

4.2 Effect of COD/N increase

COD addition did not result in a better nitrogen removal efficiency or lower AnAOB

activities (Lackner et al., 2008) as almost no COD removal was observed. Therefore, OLAND

performance was not affected by COD/N ratios of 2 and stable nitrogen removal rates were

maintained. It has been already successfully demonstrated that anammox can co-exist with

heterotrophic denitrifiers at COD/N ratios of 2.2 (Desloover et al., 2011). Therefore, it should

be possible to obtain high nitrogen removal efficiencies without the loss of AnAOB activities

at mainstream conditions.


126

4.3 NOB-AnAOB competition at mainstream conditions

Although Nitrospira sp. were present from day 0-375 (phase I-X) at a stable level of around

40 copies ng-1

DNA, at temperatures below 20°C (day 61, phase IV) NOB activity increased

significantly (Fig. S7.1). Together with the increased NOB activity, nitrite accumulated in the

system. Moreover, when COD was added and lower nitrogen concentrations (55 mg N L-1

instead of 60 mg N L-1

) were fed (phase X to XIII), relative nitrate productions up to 62%

were observed (Table 7.2, phase XII). Free ammonia (FA) and free nitrous acid (FNA)

concentrations (Table 7.1 and 7.2) were in all phases too low to suppress nitratation

(Anthonisen et al., 1976). Moreover, oxygen inputs (mainly through atmospheric contact)

were rather high to allow sufficient nitritation, which could also have stimulated NOB growth

and activity. Therefore, for mainstream treatment other strategies beside FA, FNA and

oxygen limitation should be applied to suppress nitratation. Detailed nitrogen balance tests

showed that the different feeding strategies did not affect the microbial balance in the system

(Table 7.3). However, pulse feeding allowed higher AerAOB activities increasing the

ammonium removal efficiency (phase XIII) and continuous-like feeding resulted in a

decreased ammonium and as a consequence total nitrogen removal efficiency (Table 7.3).

Moreover, a sufficient loading rate was needed to allow a good microbial balance and thus

increased AnAOB activity and/or decreased NOB activity (Table 7.2 Phase XIII; Table 7.3

test F). Overloading of the system and therefore obtaining higher FA levels, could inhibit the

NOB activity (Table 7.3 test F) in contrast to the long-term performance at lower FA

concentrations. Therefore, the latter could not be responsible for the better microbial balance

during reactor operation. However, nitrite accumulation, resulting in higher peaks in NO and

N2O production (Table 7.3) occurred in all well performing periods at low temperature. High

NO emissions, initiated by addition of nitrite could increase relative AnAOB activity and

decrease NOB activity (Fig. 7.3). It is well known that NO is toxic to most of the bacteria

(Mancinelli and McKay, 1983). It has been described before that NO2--dependent O2 uptake

by NOB could reversibly be inhibited by NO at concentrations of 7-448 μg NO-N L-1

(Starkenburg et al., 2008). In contrast, NO is an intermediate for the AnAOB metabolism and

high NO concentrations do not affect their activity (Kartal et al., 2010b). Therefore, at

conditions of high NO concentration, AnAOB can have a competitive advantage compared to

NOB. However, from the moment NO is depleted NOB activity can increase again (Fig. S7.3-

S7.5; Starkenburg et al. 2008). Although nitritation is stimulated by NO (Zart et al., 2000), it

seemed that also AerAOB activity was affected by NO at the NO/N2O peak (Fig. 7.3).

Chapter 7

127

Therefore, a balance should be found between stimulating AnAOB above NOB activity and

allowing sufficient nitrite production by AerAOB. The high volumetric loading rate applied,

together with the pulse feeding and the nitrite accumulation led to high NO/N2O emissions

compared to mesophilic OLAND applications (Kampschreur et al., 2009a; Weissenbacher et

al., 2010). This could however be a prerequisite for obtaining low nitratation levels at these

mainstream conditions.

4.4 OLAND application in the main line

At 15°C and a COD/N ratio of 2, high total nitrogen removal rates of 0.5 g N L-1

d-1

were

obtained. However, the total nitrogen removal efficiency was too low to obtain dischargeable

effluent (European Commision, 1991). As similar total nitrogen removal rates were obtained

at 15°C compared to 29°C, the performance was not limited by the mainstream conditions but

it was limited by the reactor configuration. Because the discs only had a spacing of 3 mm,

regular perforations of the biofilm were needed to allow sufficient diffusion. A better RBC

configuration (higher disc distance) or another reactor technology (suspended growth system)

could probably allow higher efficiencies due to more efficient diffusion. On the other hand,

by a combination of OLAND and conventional nitrification/denitrification in the B-step, a

better removal efficiency could be obtained. This can be achieved by only partly replacing the

activated sludge by OLAND biomass. To allow AnAOB retention in this system a selectively

higher SRT of the OLAND biomass compared to the activated sludge should be maintained.

For granules, the latter can be obtained by implementation of cyclones (Wett et al., 2010b),

but it should also be possible by inoculation of OLAND biomass on packing material which

can be kept in the system by a grid. In this way, OLAND should not be responsible for the

total nitrogen removal efficiency of the system and nitrite or nitrate formation can be

compensated by denitrification.

5 Conclusions

This study showed for the first time that total nitrogen removal rates of 0.5 g N L-1

d-1

can be

maintained when decreasing the temperature from 29°C at 15°C and when low nitrogen

concentration and moderate COD levels are treated. Nitrite accumulation together with

elevated NO/N2O emissions was needed to allow competition of AnAOB against NOB for

nitrite at low FA, low FNA and high DO levels. Further research should elucidate the

mechanism and the level of NO/N2O emission needed to obtain a balanced performance.


128

Moreover, it should be further evaluated if the increased NO/N2O emission can be

compensated with a decreased energy consumption to justify OLAND mainstream treatment.

6 Acknowledgements

H.D.C. received a PhD grant from the Institute for the Promotion of Innovation by Science

and Technology in Flanders (IWT-Vlaanderen, SB-81068) and S.E.V. was supported as a

postdoctoral fellow from the Research Foundation Flanders (FWO-Vlaanderen). The authors

gratefully thank Aquafin for providing the sewage, Eva Spieck for providing the qPCR

standards and Tom Hennebel, Joachim Desloover and Simon De Corte for inspiring scientific

discussions.

7 Supplementary data

Table S7.1: Overview of the presence of NOB and AnAOB species, confirmed with FISH

High temperature,

COD/N of 0 (day 1)

Low temperature,

COD/N of 2 (day 435)

Nitrobacter - -

Nitrospira + +

Cand. Scalindua - -

Cand. Kuenenia & Brocadia + +

Chapter 7

129

Figure S7.1: Abundance of functional Amo gene copies of AerAOB and 16SrRNA copies of AnAOB

and Nitrospira sp. measured by qPCRin the OLAND RBC.


130

Figure S7.2: Scatter plot showing the influence of the nitrogen load (A) and dissolved oxygen

concentration (B) on the AerAOB, AnAOB and NOB activity. The zone between the dashed lines

represents the nitrogen load and DO range studies in the N balance test (Table 7.3)

Chapter 7

131

Figure S7.3: Phase VII, Test B (Table 7.3): Profiles of ammonium, nitrite, nitrate and oxygen

concentration in the reactor water phase and N2O concentration in the defined air flow out of the

reactor. Gray boxes mark the feeding periods.

Figure S7.4: Phase XIII, Test C (Table 7.3): Profiles of ammonium, nitrite, nitrate and oxygen

concentration in the reactor water phase and N2O and NO concentration in the defined air flow out of

the reactor. Gray boxes mark the feeding periods.


132

Figure S7.5: Phase XIII, Test E (Table 7.3): Profiles of ammonium, nitrite, nitrate and oxygen

concentration in the reactor water phase and N2O and NO concentration in the defined air flow out of

the reactor when extra nitrite pulses were added just after the feeding period. Gray boxes mark the

feeding periods.

Chapter 8

133

Chapter 8:

Environmental assessment of one-

stage partial nitritation/anammox

implementation in sewage treatment

plants

Abstract

Implementation of one-stage nitritation/anammox (e.g. DEMON®, OLAND) for the

treatment of sludge digestates allowed energy autarky in the wastewater treatment plant

(WWTP) in Strass (Austria). To further increase the overall energy production of the plant a

first trial was performed to implement mainstream DEMON operation. To evaluate the

environmental impact of DEMON implementation, life cycle assessment (LCA) was carried

out based on on-site measurement campaigns for three scenarios: (1) without DEMON; (2)

with DEMON in the side line; (3) with DEMON in the side and main lines. The results of

these assessments showed that DEMON implementation in the side line, had a positive effect

on the eutrophication potential, abiotic depletion potential and global warming potential. For

the present situation with mainstream DEMON, 9% of the electrical needs could be saved at

the moment, but control optimization is needed to decrease N2O emissions on the long-term.

From the LCA, it could be concluded that the WWTP in Strass can be seen as one of the

benchmark plants, not only based on energy efficiency with 153 and 167% energy coverage,

but also based on overall environmental sustainability with a CO2 footprint of 7 and 36 kg

CO2 PE-1

year-1

for side and mainstream DEMON, respectively.

Chapter redrafted after: De Clippeleir H., Schaubroeck S., Weisssenbacher N., Dewulf J.,

Boeckx P., Boon N. and Wett B. Environmental assessment of one-stage nitritation/anammox

implementation in sewage treatment plants. Submitted.

Environmental assessment of the implementation of OLAND in sewage treatment plants

134

1 Introduction

Around 24000 sewage treatment plants (WWTP) are operational in Europe which together

treat about 580 million person equivalents (PE) (UWWTD, 2011). Although the main aim of

the WWTP is to decrease harmful emissions towards water bodies, recently more attention is

paid on energy efficiency and overall environmental sustainability (Verstraete and

Vlaeminck, 2011). The implementation of anaerobic digestion, which provides on-site

production of renewable energy increased significantly over the last 5 years (Chapter 4). The

latter has the potential to be implemented in around 85% of the WWTP in Europe as from a

size of around 10 000 PE anaerobic digestion starts to be economically feasible (UWWTD,

2011). Depending on the primary sludge production, performance of the anaerobic digester,

the efficiency of electrical energy production from biogas and the oxygen transfer efficiency

for aeration (Wett et al., 2007; Nowak et al., 2011), energy self-sufficiency can be reached.

Besides energy recovery through anaerobic digestion, energy minimization for nutrient

removal by implementation of one-stage partial nitritation/anammox, also known as DEMON

(Wett, 2006), OLAND (Kuai and Verstraete, 1998) and CANON (Third et al., 2001) could

positively contribute to the energy balance of the plant. At the moment around 30 one-stage

autotrophic nitrogen removal plants are operational for the treatment of sludge liquors from

anaerobic digestion, which account for 15-25% of the total nitrogen load of the WWTP

(Vlaeminck et al., 2012). High and stable nitrogen removal rates for the treatment of sludge

digestates are reported (Vlaeminck et al., 2012). Moreover, the implementation of DEMON in

the side line of the WWTP, can result in a total decrease in energy consumption of the WWTP

with more than 50% compared to conventional nitrification/denitrification (Siegrist et al.,

2008).

Further decreasing the energy consumption by up-grading the activated sludge step in the

main line by DEMON would offer two main advantages. The first advantage would be that

the aeration need for nitrogen removal in the main wastewater line can decrease by almost

60% as DEMON only requires 1.8 kg O2 kg-1

N removed and conventional

nitrification/denitrification requires 4.3 kg O2 kg-1

N removed. The second advantage of

mainstream DEMON is that this process allows higher COD removal through primary sludge

production, for example by a highly-loaded activated sludge step, as no carbon is needed in

the DEMON process to obtain high (89%) nitrogen removal in contrast to

Chapter 8

135

nitrification/denitrification which needs 3 kg COD kg-1

N. Theoretically, it was estimated that

mainstream DEMON could allow energy-positive wastewater treatment (Siegrist et al., 2008;

Verstraete and Vlaeminck, 2011; Chapter 4).

The challenges of mainstream DEMON (Vlaeminck et al., 2012) on the other hand, are the

retention of the anoxic ammonium-oxidizing bacteria (AnAOB) in the activated system

(sludge retention time (SRT) of around 10 days), as AnAOB have a doubling time of around

1-2 weeks. Moreover, nitratation suppression in the mainstream will be more challenging

because of the lower selection pressure on nitrite-oxidizing bacteria compared to side stream

conditions (Chapter 4). DEMON lowers the energy demand but, on the other hand emits on

average around 1% of the N load as N2O-N, a powerful greenhouse gas (Joss et al., 2009;

Kampschreur et al., 2009a; Weissenbacher et al., 2010). Therefore, greenhouse gas emission

at mainstream conditions should be evaluated as they can offset energy efficiency and

performance.

For a better environmental sustainability assessment, the environmental impact should not

only be considered on a process and plant level, but also from a more broader life cycle

perspective. Life cycle assessment (LCA) is an appropriate tool. It is a holistic tool

increasingly used to evaluate environmental impacts associated with a product, process or

activity (Iso, 2006a, b). LCA has been widely used to study WWTP configuration (Clauwaert

et al., 2010; Hospido et al., 2005; Hospido et al., 2008; Foley et al., 2010). These studies

showed that operational energy, direct greenhouse gas emissions and chemical consumption

generally increase with increasing nitrogen removal (Foley et al., 2010). However,

environmental impact assessment of the implementation of DEMON in WWTP has not been

evaluated before. Since a strong point of DEMON is its higher environmental sustainability

on process level due to its lower oxygen demand and its potential to allow higher energy

recovery in other steps of the WWTP, this study evaluated this on a plant and life cycle level

for different scenarios of one WWTP in Strass (Austria). The latter WWTP is based on a two-

stage activated sludge system (A/B system; Wett et al., 2007) as mainstream treatment and on

sludge digestion with electricity and heat production via a combined heat and power (CHP)

unit in the side line. Sludge digestate treatment (sidestream treatment) was in the reference

period (2003, scenario 1) based on nitritation/denitritation with A-stage sludge as external

carbon source, but has been replaced by DEMON (from 2004 onwards, scenario 2).

Preliminary results of DEMON implementation in the mainstream, as an up-grader for the B


136

stage, were also incorporated in this study (scenario 3) to elucidate the critical factors for final

implementation important to obtain a higher environmental sustainability.


2.1 Scope definition

In this research, three scenarios applied on the WWTP of Strass (Austria) were studied and

compared using the LCA framework according to the ISO 14040/14044 guidelines (Iso,

2006a, b). The three scenarios had a different degree of DEMON implementation level (from

none, to side stream and further to mainstream). The system boundaries were the same for all

scenarios. In general, the considered life cycle system included the WWTP itself, the part of

the human industrial system responsible for products (mainly chemicals and electricity)

needed in the WWTP and the part for the further processing of its waste products (composting

of the dewatered digestate). The foreground system, main system of interest, consisted of the

WWTP. The other parts of the life cycle were considered as the background system. The

transportation of chemicals from the supplier to the WWTP, the transportation of the

dewatered digestate to the composting facility and the infrastructure of the WWTP and

composting facility were excluded from the life cycle systems. The infrastructure is left out

since the focus was on the real-time operation of the WWTP. Not included in the life cycles

were the upstream collection and transportation of the municipal wastewater to the WWTP

and the usage/disposal on land of the compost processed out of the digestate.

The main system input was wastewater, which was considered as a waste product in LCA and

therefore no environmental impact of its generation was allocated to it (Iso, 2006b). Likewise,

the addition of co-substrate to the digester, which consisted out of kitchen waste and fat, was

considered in this study as a waste product. The organic carbon present in the wastewater was

assumed to be 100% biogenic, neglecting the amount of fossil carbon from detergents and

soaps (Griffith et al., 2009). CO2 emissions from the oxidation of this biogenic carbon were

by consequence considered as biogenic in compliance with the Intergovernmental Panel on

Climate Change (IPCC) accounting guidelines (Doorn et al., 2006). Two products were

formed in the studied life cycles namely electricity and compost. The conventional production

of these products was avoided and thus also the total impacts of their production processes.

Their impacts should therefore be subtracted from the total impact of the life cycle

(Finnveden et al., 2009). Electricity produced at the plant displaced electricity provision by

Chapter 8

137

the grid and thus electricity production in Austria. The substitutability of the compost, which

was intended for agricultural application, was assumed to be 50% for nitrogen and 70% for

phosphorus (Bengtsson et al., 1997). Concerning industrial products avoided, fertilizer mixes

with a similar nutrient composition were chosen (see data inventory).

The functional unit (FU) was the treatment of 1 m3 of sewage. The effluents of the different

scenarios (Table 8.1) were all in compliance with Austrian legislation regarding necessary

water quality (BGBL, 1996). Additionally, the waste sludge after dewatering complied with

the legal guidelines in terms of composition, especially the presence and quantity of heavy

metals for agricultural application (BGBL, 1996).

2.2 Plant description

The plant in Strass (Austria) is based on a two-stage activated sludge system in the main

water line, referred to as an A/B plant (A/B Verfahren, Wett et al., 2007). The first step is a

high rate activated sludge step with short hydraulic (30 minutes) and sludge (0.5 days)

retention time. About 50-60% of the COD is removed in the A-stage and due to the short

retention time the organics are adsorbed or incorporated in the sludge and not emitted as CO2.

Nitrogen and phosphorous removal in this step mainly occurs via the organic fraction and

only accounts for on average 23 and 26%, respectively. The second activated sludge step

(B-step) is a low loaded step with temperature dependent aerobic SRT of ca. 10 days. This

step consists of a predenitrification and nitrification step with recycle from the second to the

first step. The aeration is controlled by a combination of dissolved oxygen (DO) and NH4+

measurement to obtain optimal effluent quality without excess aeration. Sludge from the

mainline (A and B-stage) is send to the digester. Organic co-substrate, which mainly

consisted out of kitchen waste, was added to enhance the electrical energy recovery. Sludge

waste after dewatering of the digestate is composted in an external facility. The liquid fraction

from the digestate after filtration was initially treated in a separate nitritation denitritation

reactor (Fig. 8.1A), but was currently replaced by a DEMON system (scenario 2, Fig. 8.1B).

DEMON was further implemented in the B-stage by inoculation with DEMON granules and

implementation of cyclones in the sludge recycle to maintain the DEMON granules in the

system (scenario 3). The hydraulics and aeration control system was not changed (Fig. 8.1C).


138

Table 8.1: Monthly averages of the parameters of the inventory data for the different scenarios studied. All data are presented in function of 1 m3 sewage

treated. N/DN: nitrification/denitrification; N/DN*: nitritation/denitritation; DEMON: deammonification; DM: dry matter

Scenario 1a Scenario 1b Scenario 2 Scenario 3

Mainstream treatment N/DN N/DN DEMON

Sidestream treatment N/DN* DEMON DEMON

Inputs to foreground system

Waste

Water

COD (g) 666 643 526

NH4+-N (g) 27 28 23

Organic N (g) 19 17 13

PO43-

-P (g) 9 9 7

Co-substrate (g DM) 53 338 239

Products (External resources)

Electricity from the grid (Wh) 86.49 1.55 0.829

Sodium aluminate (g) 78.4 46.9 35.3

Flocculant (g) - 9.68 7.29

FeCl2 (g) 10.9 9.56 4.33

Polymer (g) 1.68 1.92 1.63

Avoided products (resources)

C fertilizer mix (g peat/g straw) 140/258 61/111 76/139

N&P fertilizer mix (g P)* 7.92 4.06 3.76

P fertilizer mix (g P)* 0 3.40 2.66

N fertilizer mix (g N)* 1.30 0 0

Chapter 8

139

Electricity into the grid (Wh) 0 179 209

Emissions to water

COD (g) 24 24 28

NH4+-N (g) 2 1 2

NO2--N (g) 0 0.13 1

NO3--N (g) 4 4 2

N org –N (g) 2 0.87 1

PO43-

-P (g) 0.710 0.27 0.39

Emissions to air

CH4 (g) 0.668 0.658 0.249

N2O (g) 0.325 1.59 0.520 2.45

NO (g) 0.00516 0.0318 0.0158 0.0134

NO2 (g) 1.59 1.269 1.275

CO (g) 0.659 0.799 0.803

CO2-biogenic (g) 357 547 518

SO2 (g) 0.122 0.0733 0.0737

* the specific composition of the fertilizer mix can be found in Table S8.1


140

Figure 8.1: Schematic overview of the 3 scenarios studied. Scenario 1 included a

nitritation/denitritation (N/DN*) in the side line (A), while scenario 2 and 3 had a DEMON reactor for

digestate treatment (B, C, respectively). Moreover, in scenario 3, the low loaded activated sludge step

(B-step) is upgraded to a DEMON step. AD: anaerobic digestion; CHP: combined heat and power;

AS: activated sludge

Chapter 8

141

2.3 Data inventory

Foreground data were collected directly from the WWTP itself. Operational data

(water/sludge flows, water/sludge composition, energy demands, chemical usage etc) for

scenarios 1, 2 and 3 were collected from daily logging results of April to July 2003, April

2011 and April 2012, respectively. Greenhouse gas emissions were measured on-site at all

biological treatment steps during April 2011, April-May 2012 for scenario 2 and 3,

respectively, following the method described for N2O emission measurements by Desloover

et al. (2011a) and for NO and NO2 measurements by Weissenbacher et al. (2010). Additional

gas measurement campaigns during July 2011 and November 2011, confirmed the emission

factors used in the different scenarios. For scenario 1, similar emissions for the A and B-step

in the mainline compared to the emissions measured for scenario 2 were considered. For the

side line treatment in scenario 1, two different emission estimations were proposed: (a) the

same emission as measured for a DEMON reactor, (b) an increased N2O and NO emission

(factor 5) because of the high nitrite concentrations (Desloover et al., 2011a). An overview of

all the in- and out-coming flows of the foreground system is given in Table 8.1.

Data for the processes of the background system were retrieved from the ecoinvent v2.2

database (Swiss Centre for Life Cycle, 2010), unless mentioned otherwise. The electricity

production mix of Austria (last update: June 2010) originated from burning of fossil fuels

(≈ 7%), nuclear energy (≈5%) and renewable energy (≈5 %) of which most part is

hydropower generated in Austria itself (≈9 %). Most chemical products added to the WWTP

were not present as such in the ecoinvent database. Their life cycle data and eventually their

impact calculated in the impact assessment phase were replaced by those of other products

available in the ecoinvent database, namely similar products or individual reagents needed for

their production. Sodium aluminate (NaAl(OH)4) added in high quantity was substituted by

stoichiometric quantities of its conventional reagents: sodium hydroxide (NaOH) and

aluminium hydroxide (Al(OH)3). The iron(II)chloride (FeCl2) (32%) solution was replaced by

iron(III)chloride (FeCl3) (40%). ‘S dflock K2’, a flocculent containing FeCl3 (3.00%) and

aluminiumchloride (AlCl3) (9.50%), was replaced by a solution of only FeCl3 (12.50%).

‘Zetag’ is a polyacrylamide and it is a polymer of acrylamide, which is in its turn formed by

hydratation of acrylonitril. ‘Zetag’ was substituted by acrylonitril. The amount of

‘Flockungsmittel M se K222L’ added was neglected. The composting process was custom

made. For 1 ton dry matter (DM) of biosolids input a fixed assumed amount of 75.73 Wh


142

electricity consumption and emissions of 163.3 g N2O, 63.90 g NH3, 1278 g CH4 were set for

all scenarios based on figures of the direct composting of biodegradable municipal waste (Van

Ewijk, 2008). It was assumed that 14% of carbon was removed during the composting process

(personal communication Steven De Meester, May 2012). Out of this carbon removal, the

amount of biogenic CO2 emissions per ton DM (scenario 1: 158 kg; scenario 2: 72 kg;

scenario 3: 123 kg) and residual carbon (86% of original C) left in the compost were

calculated. The final compost consisted of this residual carbon and the same amount as that of

the dewatered digestates for the other nutrients, neglecting the small amounts of these

removed by gaseous emission and via the leachate during the composting process. The

compost of each scenario was displaced by a fertilizer mix using the substitutability factor of

50 and 70% for N and P, respectively (Bengtsson et al., 1997). This was specifically done by

defining a mix with the representative C, N and P content as the displaced compost, based on

the methodology described by Hermann et al. (2011). For the amount of carbon present in the

compost, an amount of peat and straw in a ratio of 1:3 were calculated to replace the amount

of humus carbon (51% of its carbon content) present in the compost (Hermann et al., 2011).

The humus carbon present in peat and straw are 0.077 and 0.084 kg kg-1

fresh matter (FM),

respectively (Hermann et al., 2011). Out of these figures, FM quantities of peat and straw

were calculated. For the other nutrients, only nitrogen (N) and phosphorous (P) were taken

into account. The amounts of P and N in the carbon fertilizers (Phyllis-Database, 2012), peat

(P: 0.1%; N: 1%) and straw (P: 0.091%; N: 0.71%) were substracted from that in the compost,

leading to the amount needed to be displaced by other fertilizers. Different fertilizers are on

the market for these two nutrients. Out of the 2010 consumption figures of such fertilizers in

Western and Central Europe (International Fertilizer Industry, 2012). P&N, N and P fertilizer

mixes were constructed (Table S8.1). For the individual fertilizers, ecoinvent data was

available. To determine the fertilizer mix for each scenario, first, an amount of P&N fertilizer

mix was calculated. Thereafter, an extra quantity of N or P fertilizer mix was quantified for

the leftover P or N, respectively, not covered by the P&N fertilizer mix.

2.4 Impact assessment

The impact assessment was done using Simapro version 7.2 software (with ecoinvent

database version 2.2) and the selected method was the ‘CML 2001 method (all impact

categories)’ version 2.05, normalized for West-Europe 1995, hereafter referred to as the

‘CML 2001’ method. A CML method was selected since these are most commonly applied in

other LCAs of WWTPs (Hospido et al., 2008; Clauwaert et al., 2010; Foley et al., 2010). This

Chapter 8

143

method was developed by the Center of Environmental Science of Leiden University (CML),

the Netherlands. The version in the Simapro software was based on the spreadsheet version

3.2 (December 2007) published on the CML web site (http://www.cml.leiden.edu/). Important

to notice is that biogenic CO2 uptake and emissions were not accounted for in the global

warming potential (GWP) 100a (100 years) category. Resources are accounted for using

abiotic depletion. In the impact assessment, the emissions of the foreground system (the

WWTP) were selected to end up in a low populated area.

3 Results and discussion

3.1 Impact of nitrogen removal process on process level

The conventional process for nitrogen removal is nitrification/denitrification (N/DN) and was

applied in the main line of the WWTP of Strass (B-stage). COD/N ratios send to the B-stage

were 7.2, 8.2 and 9.2 for scenario 1, 2 and 3, respectively and therefore allowed full

denitrification. A cost-saving alternative for the conventional nitrification/denitrification is the

application of nitritation/denitritation, saving theoretically around 24% of aeration

requirement as during this process nitrite oxidation (nitratation) is avoided (Vlaeminck et al.,

2012). Moreover, the COD demand and sludge production can decrease with 50% and 40%,

respectively (Vlaeminck et al., 2012). This process was applied for digestate treatment in the

WWTP during scenario 1 (Fig. 8.1). Compared to the energy requirements for

nitrification/denitrification in the mainstream of the WWTP in Strass, assuming that 50% of

the available COD was denitrified and an oxygen transfer efficiency of 2 kg O2 kWh-1

was

applicable, the nitritation/denitritation applied in the side line decreased the energy

requirements from 4.30 to 2.65 kWh kg-1

N removed. So, this process has a potential to save

around 38% of the energy needed for aeration during nitrogen removal. DEMON

implementation in the side line of the WWTP could further decrease the energy requirement

for nitrogen removal to 1.52 kWh kg-1

N removed for scenario 2. This additional 43%

decrease in energy requirement for aeration is exactly what is theoretically expected for

DEMON implementation compared to nitritation/denitritation (Vlaeminck et al., 2012).

Besides energy savings for aeration, the choice of nitrogen removal process has an influence

on the energy recovery potential as nitritation/denitritation and DEMON save 52 and 100% of

COD source needed compared to conventional nitrification/denitrification (Vlaeminck et al.,

2012). Therefore, these processes can increase electricity production at the plant through


144

anaerobic digestion of primary sludge, which was otherwise used as a COD source (scenario

1).

The choice of nitrogen removal process also determines the sludge production. Conventional

nitrification/denitrification compared to nitritation/denitritation and DEMON produces around

1 compared to 0.6 and 0.1 kg sludge kg-1

N removed, respectively (Vlaeminck et al., 2012).

The latter is the result of the avoidance of heterotrophic growth in the system. DEMON

implementation can therefore lower production sludge and as a consequence emissions,

chemicals, electricity and waste products related to sludge handling.

At this moment insufficient comparative data are available for N2O emissions in nitrogen

removing processes, making it hard to evaluate the impact difference of these processes.

Reported N2O emission in activated sludge systems based on nitrification/denitrification

ranged from 0.001-25% of the N load and were mainly linked with nitritation activity

(Desloover et al., 2011b). Moreover, the N2O emission is mainly linked with the operational

conditions rather than the process itself (Kampschreur et al., 2009b; Chandran et al., 2011).

For the treatment plant of Strass, N2O-N emissions in the nitrification/denitrification step (B-

stage) were very low, e.g. 0.01% of the N load compared to 1.0-1.3% of the N load measured

during side stream DEMON (scenario 2 and 3). As nitrite, a precursor for N2O production

(Kampschreur et al., 2009b; Chandran et al., 2011), accumulated up to 96 mg N L-1

during

nitritation/denitritation compared to 0 mg N L-1

and 1 mg N L-1

in the nitrification/

denitrification and DEMON step, respectively, increased levels of N2O were expected for

nitritation/denitritation (scenario 1b). Due to possible nitrite accumulation in processes based

on nitritation, an increased level of N2O emission in the mainline was therefore expected

when DEMON was implemented in the B-stage initially based on nitrification/denitrification.

From the comparison of the different nitrogen removal processes, it could be concluded that

DEMON can lower the electricity needs, the sludge production and the COD demand, but has

the potential to increase N2O emission due the higher risk for accumulation of nitrite.

Chapter 8

145

3.2 From energy-negative to energy-positive WWTP on system level

On the WWTP level, scenario 2 including the implementation of DEMON in the side line

caused a 24% decrease in total energy consumption from 0.45 to 0.34 kWh m-3

sewage treated

(scenario 1 compared to 2). Although DEMON implementation could decrease the energy

needs for nitrogen removal with 43% in the side line (scenario 2), the relative energy

requirement of the DEMON reactor remained around 0.02 kWh m-3

raw sewage due to the

low contribution of side line treatment on the total energy consumption (4%). The latter was

also the result of the higher kitchen waste dosage, which increased the nitrogen load to the

DEMON reactor in comparison to scenario 1. Assuming relative on the incoming sewage, the

same nitrogen load to the DEMON reactor as in scenario 1, 0.01 kWh m-3

sewage could

potentially be directly saved by the implementation of DEMON due to lower aeration

requirements in the side line. Moreover, implementation of DEMON allowed a higher sludge

load to the digester (0.14 kg TS m-3

sewage) as this process had no need for a carbon source.

Therefore, DEMON implementation could directly increase the energy recovery and thus the

electrical energy production with 0.06 kWh m-3

sewage (13% of energy requirement in

scenario 1) without taking the co-substrate addition into account. DEMON implementation

could result in an electrical energy production on-site of 93% of the needs instead of 80% in

scenario 1. This increased electrical energy recovery due to DEMON implementation could

have been at least a factor 2 higher when compared to systems, which send the digestate

directly to the B-stage (Siegrist et al., 2008; Chapter 4). However, the observed energy

consumption decrease of 0.11 kWh m-3

sewage in scenario 2 compared to scenario 1 was also

attributed to further optimizations in the A-stage, B-stage and sludge handling which allowed

an energy consumption decrease of 0.02, 0.04 and 0.05 kWh m-3

sewage treated, respectively.

Due to an overall better energy efficiency both in the side line but as a consequence also in

the A- and B-stage, the electrical input of 0.087 kWh m-3

sewage of scenario 1 was avoided,

even without considering the higher biogas production by the increased sludge load and

addition of co-substrate to the digester during scenario 2. Therefore, DEMON implementation

together with energy optimizations in the other steps led to an energy self-sufficient system

(Wett et al., 2007). The addition of co-substrate increased the energy net production further to

153% of the energy demand of the plant (scenario 2), instead of 109% assuming that the

biogas production remained constant.


146

To further increase the energy efficiency of the plant, the potential of DEMON

implementation in the main water line (scenario 3) was tested at full-scale as the energy needs

for aeration in the B-stage accounts for 40% of the total energy needs of the WWTP (scenario

1 and 2). DEMON granules from the side stream reactor were inoculated in the B-stage and

the SRT of the DEMON granules was increased compared to the activated sludge flocs by the

installation of cyclones in the sludge recycle (Wett et al., 2012). Although the operation of the

B-stage remained the same in terms of hydraulics, oxygen profile and loading, a metabolic

shift was observed characterized by a significant decrease in nitrate concentration and

increase in nitrite concentration in the reactor. Especially in winter when the highest loading

rates were supplied, nitrite concentrations were higher than nitrate concentrations in the

effluent (Wett et al., 2012). A NO3--N over NO2

--N ratio of 2 was observed in the effluent of

scenario 3 (April 2012), compared to a ratio of 31 in scenario 2 (Table 8.1). The latter

indicated a decrease in the nitrite oxidation activity (nitratation), which was the first

prerequisite to allow anammox activity. Due to the higher COD/N ratio, the energy savings by

nitrogen removal through anammox were counteracted by the increased aerobic COD removal

in the B stage. It could be calculated that compared to nitrification/denitrification (scenario 2),

the energy demand for DEMON (0.9 kWh kg-1

N) increased with 1.4 kWh kg-1

N due to the

aerobic removal of the COD which was normally denitrified and with 0.5 kWh kg-1

N due to

the higher incoming COD/N ratio. As a consequence, the expected lower oxygen demand and

thus energy demand observed in the B-stage was only minor i.e. 0.12 instead of 0.14 kWh m-3

sewage. It is suggested that increasing the COD removal in the A-stage through primary

sludge production and thus decreasing the COD/N ratios of the B-stage influent could

increase the role of the anammox bacteria in the mainstream and would therefore allow higher

energy saving in the B-stage and higher energy recovery by the plant itself. During scenario 3,

electricity production remained constant, but the total energy consumption decreased from

0.34 to 0.31 kWh m-3

sewage, due to the poorly working A-stage (Table 8.2) and a decrease in

the energy demand in the B-stage. Therefore, the electricity production increased to 167% of

the electrical energy needs of the plant, showing the potential of scenario 3 to allow higher

energy recovery.

Chapter 8

147

Table 8.2: Overview of the performance for the different step of the WWTP. AD: anaerobic digestion;

TS: total solids; tot: total

Scenario 1 Scenario 2 Scenario 3

Mainstream treatment N/DN N/DN DEMON

Sidestream treatment N/DN* DEMON DEMON

High loaded AS

COD removal (%) 61 53 48

Ntot removal (%) 25 21 21

P removal (%) 25 25 30

Low loaded AS

COD removal (%) 91 92 90


P removal (%) 78 85 92

Sludge digestion

Co-substrate addition (% of AD feed) 15 44 39

TS to biogas (%) 39 66 64

Biogas yield (m3 kg

-1 TS input) 0.377 0.283 0.357

Reject water treatment

COD removal (%) 85° 48 45


°primary sludge was added as carbon source

3.3 Environmental impact of DEMON implementation on life cycle level

3.3.1 Eutrophication

The primary objective of a WWTP is to decrease COD, N and P concentrations in the water

phase and to obtain dischargeable effluent qualities. In all scenarios the eutrophication

potential (EUP) of the wastewater (0.05 kg PO43-

-eq m-3

sewage) decreased sharply with 91,

94 and 93% for scenario 1, 2 and 3, respectively, by implementation of the sewage treatment

system as expected (Fig. 8.2A). The best effluent quality was obtained during scenario 2. The

higher ammonium, phosphorus and COD effluent concentrations (Table 8.1) caused the

increased EUP of the WWTP during scenario 3. However, the lower effluent nitrate

concentration during scenario 3 resulted in a more equally distribution of the EUP over the

different effluent compounds (Fig. 8.3). A 60% decrease of the ammonium effluent

concentration to 1 mg N L-1

by a more stringent control of the aeration system would result in

the same EUP as during scenario 2. It is therefore expected that optimization of the

operational conditions (DO and ammonium set point) will limit the EUP.


148

Figure 8.2: Contribution of the different elements of the LCA system to (A) the eutrophication

potential, (B) the abiotic depletion potential and (C) the global warming potential for the different

scenarios studied. The wastewater itself had an eutrophication potential of 0.05 kg PO43-

-eq m-3

sewage. As the GWP of the WWTP was dominated by N2O (>90%), two different N2O emission

scenarios for scenario 1 were included. Negative values are related to impacts, which are avoided by

recovery of products on-site.

Chapter 8

149

Figure 8.3: Contribution of the different substances emitted by the WWTP to eutrophication potential

(kg PO43-

-eq m-3

sewage)

3.3.2 Abiotic depletion potential

The abiotic depletion potential (ADP) takes into account the use of abiotic resources such as

iron ore and crude oil (Guinee, 2001). Energy consumption and chemical addition are the

main factors of the described system, which relies on the use of abiotic resources (Fig. 8.2B).

The difference between the scenarios (1>2>3) in the use of sodium aluminate, needed for P

removal, dominated the ADP of chemical addition (Table 8.1. Fig. 8.2B). However, it should

be noted that the sodium aluminate used at this plant was considered as a product and not as a

waste, although sodium aluminate was retrieved from the alum industry nearby. Also, the

recovery of nitrogen and more importantly phosphorus through composting significantly

influenced the ADP of the WWTP. The resource intensive processes of the production of the

fertilizers are the main cause of this (Silva and Kulay, 2003). Therefore, this indicates that P

and N recovery from a life cycle perspective can be an important factor in decreasing the

resource needs and counteracting the need for chemicals and electricity. Besides nutrient

recovery, also electricity production (scenario 2 and 3) further decreased the ADP. So, it can

be concluded that for this impact category the implementation of DEMON in the side

(scenario 2) and mainstream (scenario 3) is advantageous, because it allows a higher net

electricity production. It should also be noted that depending of the electricity mix used, the

effect of electricity production on-site can increase with a factor 2 depending on the country

(Fig. S8.1). For example, energy recovery in countries that produce electricity from hard coal

or natural gas (i.e. Netherlands, Poland, USA) will have a bigger effect on the abiotic


150

depletion potential compared to countries based on hydropower or nuclear power (i.e.

Norway, Austria, France, Belgium) based on the ecoinvent v2.2 database.

3.3.3 Global warming potential

Global warming caused by an enhanced greenhouse effect is defined as the impact of human

radiative active gas emission on the radiative forcing of the atmosphere, causing the

temperature at the earth’s surface to rise (Guinee, 2001). It should be noted that biogenic

formation of CO2 is not incorporated in the global warming potential (GWP) in contrast to

fossil-based CO2, methane (CH4) and nitrous oxide (N2O) emission, which accounts for 1, 25

and 298 kg CO2-eq kg-1

emission, respectively. Figure 8.2C shows that the GWP is mainly

dominated by the greenhouse gases emission of the WWTP itself, although also electrical

energy consumption, composting and the production of chemicals contributed. On the other

hand, the recovery of C, N and P through composting and production of electricity on-site

saved greenhouse gas emissions produced during the production of the respective fertilizer

mixes and electricity at the Austrian grid.

Due to the high GWP of N2O, the CO2 footprint of the WWTP was for 91, 98, 98 and 99%

determined by the total N2O emissions measured for scenario 1a, 1b, 2 and 3, respectively. As

the latter was the only difference between scenario 1a and 1b, a 5-fold increase in N2O

emission in the side line, increased the total CO2 footprint of the plant from 0.16 to 0.53 kg

CO2-eq m-3

sewage. As the emissions between scenario 2 and 1a were similar, the lower total

GWP of the plant during scenario 2 was mainly caused by a net electricity production (Fig.

8.2C). The CO2 footprint of the plant with DEMON in the side line was 0.12 kg CO2 m-3

sewage or around 7 kg CO2-eq PE-1

year-1

, which is low compared to the average reported

operational CO2 footprints of WWTP ranging from 12-80 kg CO2-eq PE-1

year-1

(Hospido et

al., 2008; Clauwaert et al., 2010). Moreover, dependent of the electricity mix provided by the

grid, a CO2 neutral WWTP based on GWP is feasible as the GWP of electricity production

can significantly differ per country (Fig. S8.1).

During the implementation of DEMON in the mainline of the WWTP in Strass, nitrite

accumulation was observed increasing the B-stage N2O emission from negligible to 2.3% of

its N-load (Fig. 8.4). This increase caused the higher CO2 footprint of the total system during

scenario 3 compared to scenario 2: 0.66 compared to 0.12 kg CO2-eq m-3

sewage treated,

respectively (Fig. 8.2C). As the mainstream DEMON operation was not stable yet and

Chapter 8

151

adaptation and improved process control could probably lower the N2O emissions (Ahn et al.,

2011), the CO2 footprint can be further optimized. It could be estimated that one should aim

for a maximum N2O emission in the mainstream DEMON reactor of 0.5 % of the N-load to

maintain the same CO2 footprint as in scenario 2. However, it should also be noted that the

CO2 footprint of scenario 3 (36 kg CO2-eq PE-1

y-1

) still correlated well with the average CO2

footprints of WWTP (Hospido et al., 2008; Clauwaert et al., 2010). This indicated that the

WWTP of Strass can be seen as a benchmark WWTP, not only based on energy efficiency but

also based on GWP.

3.3.4 Impact categories of minor importance for DEMON implementation

The acidification potential (AC) was mainly counteracted by the recovery of nitrogen and

phosphorous (Table 8.3). For the plant itself, the SO2 emission of the CHP unit was the main

factor that contributed to the AC potential. Therefore, DEMON application as such had no

significant influence on this impact category. The same minor influence of DEMON

implementation could be observed for the ozone depletion potential, which was mainly

dominated by chemical usage, for the ecotoxicity, which was related with the amount of

nutrient recovery and for the photochemical oxidation potential which was mainly influenced

by the emissions (CO, NO2 and SO2) from the CHP unit (Table 8.3). It should however been

noted that the ecotoxicity impact in this study was relatively low in contrast to reported LCA

studies for WWTP (Hospido et al., 2008; Clauwaert et al., 2010; Foley et al., 2010), because

the usage phase of the compost was excluded and thus also the environmental impact of its

metal content.


152

Figure 8.4: N2O, NO and CO2 emission data selection from the B-stage during scenario 2 (top) and 3

(bottom) which had a N load of 780 and 826 kg N d-1

, respectively. Fluctuations in the CO2 emissions

were strongly correlated with the aeration regime.

Chapter 8

153

Table 8.3: Results of the impact assessment for the acidification potential (AC), ozone depletion potential (OD), photochemical oxidation potential (PO),

freshwater aquatic ecotoxicity (FET) and terrestrial ecotoxicity (TET). Negative values are related to impacts that are avoided for the production of C, N and P

fertilizers. Impact category Unit Scenario 1 Scenario 2 Scenario 3

Total WWTP Total WWTP Total WWTP

AC (kg SO2-eq m-3

sewage) 0.00033 0.00094 -5 10-5

0.00072 4.5 10-5

0.00073

OD (kg CFC-11-eq m-3

sewage) 2.1 10-8

0 1.4 10-8

0 3.0 10-9

0

PO (kg C2H4-eq m-3

sewage) 5.3 10-5

7.0 10-5

3.3 10-5

6.3 10-5

3.6 10-5

6.2 10-5

FET (kg 1.4-DB-eq m-3

sewage) -0.0034 0 -0.0030 0 -0.0027 0

TET (kg 1.4-DB-eq m-3

sewage) -7.9 0 -10 0 -8.7 0

MET (kg 1.4-DB-eq m-3

sewage) 0.00049 0 0.00010 0 -0.00018 0


154

4 Conclusions

On plant level, DEMON implementation, which excluded the need for a COD sources in the

side line, had the potential to save 13% of the electricity consumption through a higher

electrical energy recovery. Besides the saving in resources, side stream DEMON

implementation positively influenced the eutrophication and global warming potential, the

most important categories of the LCA of the WWTP in Strass. First results of DEMON

implementation in the mainstream of the WWTP showed the potential to further decrease the

energy consumption and therefore also the abiotic depletion potential. However, the first tests

also showed a higher risk for increased eutrophication potential and increased global warming

potential due to increased N2O emissions. Therefore, further optimization of the operational

conditions will be needed to obtain an environmental sustainable treatment plant with

DEMON in the mainstream.

5 Acknowledgements

H.D.C. is a supported by a PhD grant from the Institute for the Promotion of Innovation by

Science and Technology in Flanders (IWT-Vlaanderen, number SB-81068). T.S. is granted by

a research project (number 3G092310) of the Research Foundation - Flanders (FWO-

Vlaanderen). The investigations at the Strass treatment plant were also supported by the

Austrian Federal Ministry of Environment. The authors gratefully thank Tim Lacoere for

technical support, Martin Hell for providing operational data of the plant and Steven De

Meester, Rodrigo Alvarenga, Siegfried E. Vlaeminck and Chris Callewaert for inspiring

scientific discussions.

Chapter 8

155

6 Supplementary data

Table S8.1: P&N. N and P fertilizer mixes composition based on the consumption of the individual

fertilizers in 2010 (International Fertilizer Industry, 2012).

Compounds Consumption

(2010)

Amount per kg of

fertilizer mix

P&N fertilizer mix* ktonnes P2O5/yr % P

Monoammonium phosphate NH4H2PO4 313 26.16

Diammonium phosphate (NH4)2HPO4 884 73.84

N fertilizer mix ktonnes N/yr % N

Urea CO(NH2)2 4950 43.59

Ammonium nitrate NH4NO3 2714 23.90

Calcium ammonium nitrate 20-30% CaCO3

and 70-80%

NH4NO3

2928 25.79

Ammonium sulphate (NH4)2SO4 763 6.72

P fertilizer mix ktonnes P2O5/yr % P

Triple super phosphate Ca(H2PO4)2 251 100.00

*Amount of N per kg of P in P&N fertilizer mix is 0.7861


156

Figure S8.1: Country-dependent impact of the electricity mix used on the global warming potential and abiotic depletion potential.

157

158

Lab-scale OLAND rotating biological contactor (RBC junior, LabMET)

Chapter 9

159

Chapter 9:

General discussion and perspectives

1 Main outcome and positioning of this work

In this doctoral work, in a first phase, the output of the OLAND process for the treatment of

digestates was optimized and studied in detail. Low volumetric exchange ratios, which assure

stable hydraulic conditions, were needed to allow a fast start-up, granulation and high

performance in SBR systems (Chapter 2). The sustainability of the process in terms of

NO/N2O emissions, was mainly linked with accumulations of intermediates such as NO2- and

NH2OH and the frequency of transient conditions (Chapter 3). Better understanding of the

conditions which lead to the accumulation of intermediates and further optimization of the

feeding pattern which determines the degree of fluctuations, will allow a further decrease of

the N2O emission in these systems.

In a second part of this work, new application domains for the OLAND process, which could

improve the overall sustainability of the applied processes, were explored. Energy

calculations revealed that OLAND could significantly increase the energy index of agro-

industrial and OFMSW-based treatment system from 3-5 to 6-10 (Chapter 4). However, for

manure-based digestate treatment, OLAND application seemed more difficult and therefore

ammonia gas treatment by OLAND was suggested for this application domain. A pilot-scale

OLAND biofilter fed with a flow of ammonia gas, obtained a high performance (0.7 g N L-1

d-1

) and a high total nitrogen removal efficiency (75-80%). Although the filter was saturated

with oxygen, the low relative water flow rate ratio (≈1 L g-1

Nin) ensured high FA

concentration in the water phase, which resulted in a dominance of AnAOB compared to

NOB activity at the top of the biofilter (Chapter 5).

A specific application domain, which could particularly improve the energy efficiency of

sewage treatment plants was the implementation of OLAND in the mainstream of the system.

This would allow a net electrical energy production, due to a higher carbon recovery and

lower energy needs for aeration (Chapter 4). Four challenges to allow mainstream OLAND


160

were encountered. A first challenge, namely the performance of OLAND at low nitrogen

concentration and low hydraulic residence time was shown in an OLAND RBC (Chapter 6).

The reactor obtained high nitrogen removal rates (0.4 g N L-1

d-1

) treating nitrogen

concentration of 30-60 mg N L-1

at a HRT of 1-2 hours. A second challenge, operation at low

temperatures (15°C), was surmounted in the same RBC by gradually decreasing the

temperature starting from 29°C. During operation at 15°C with synthetic feed (60 mg N L-1

)

and a HRT of 1h, a similar nitrogen removal rate as at high temperatures was obtained i.e. 0.5

g N L-1

d-1

(Chapter 7). Compared to higher temperatures only a decrease of the total removal

efficiency of 22% was detected. The switch from synthetic feed to pretreated sewage with a

COD/N ratio of 2 (challenge 3) did not significantly affect the performance (Chapter 7).

However, during the low temperature performance of the RBC system, NOB activity started

to increase, as well as competition between AnAOB and NOB for nitrite (challenge 4). It was

shown that increased levels of NO selectively enhanced AnAOB over NOB activity (Chapter

7). Therefore, high peak loading rates together with nitrite accumulation, increasing the NO

production, enhanced the overall removal efficiency. To evaluate the mainstream OLAND

application in a broader context, a LCA was performed on full-scale data of the WWTP in

Strass, in which an OLAND-type of process, called DEMON was implemented. Three

scenarios were studied: (1) the WWTP without a DEMON system; (2) the WWTP with

DEMON in the side line; (3) the WWTP with DEMON in the main line. For the latter

scenario, data from a first full-scale trial were used. The LCA showed that implementation of

DEMON in the side line of the WWTP positively influenced all impact categories and

therefore resulted in a more sustainable WWTP. The first full-scale results ever of DEMON

implementation in the mainstream of the WWTP in Strass (Austria) showed that to obtain the

same degree of sustainability than the sidestream treatment, the N2O emission (around 2% of

N load) in the main line should be decreased. As N2O emission is mainly related with

operational conditions and not with the process itself, it should be possible to further optimize

the emission to around 0.5% of the N load allowing the same CO2 footprint of the plant in

comparison with sidestream DEMON implementation.

2 OLAND and sustainability

2.1 Balancing energy recovery with sustainability

The overall CO2 footprint of a WWTP is dominated by the amount of N2O emitted (Chapter

8, Foley et al. 2010). Generally, it is accepted that the OLAND process for the treatment of

Chapter 9

161

highly N-loaded streams such as digestates, emits around 1% of the N load as N2O

(Chapter 3, Kampschreur et al., 2009a; Weissenbacher et al., 2010). This is mainly steered by

the degree of transient oxygen and N-loading conditions and the accumulation of NH2OH and

NO2-. However, for mainstream OLAND it is not clear yet if the high N2O emissions (1-5%

of N load, Chapter 7-8) are essential to allow AnAOB activity at low nitrogen concentration

and low temperatures or if there is room for further optimization. Therefore, further research

in this field is needed to achieve full-scale mainstream OLAND treatment.

At this moment, insufficient comparative data are available about N2O emissions in

conventional activated sludge systems with nutrient removal, making it hard to set a critical

level of acceptable N2O emissions for mainstream OLAND. Reported N2O emission in

activated sludge systems ranged from 0.001-25% of the N load and were mainly related to

nitritation activity (Desloover et al., 2011b). To evaluate if a certain level of N2O emission is

acceptable for mainstream OLAND, it is proposed to evaluate the sustainability of the plant

before and after implementation of mainstream OLAND (Chapter 8). It should be possible to

counteract a certain increase in the N2O emission by the increase in energy recovery to

maintain a constant CO2 footprint. For the treatment plant in Strass (Austria), it was estimated

that a N2O emission in the mainstream of 0.5% of the N load would be acceptable compared

to the current performance (Chapter 8).

As the degree of N2O emissions are strongly related to the operational conditions rather than

the nitrification/denitrification or OLAND process, mitigation of N2O emission should be

possible (Chandran et al., 2011; Desloover et al., 2011b). In the next section, mitigation

strategies are proposed based on the control of the N2O production and control of the N2O

emission.

2.2 Mitigation strategies based on chemical markers

In Chapter 3, a detailed analysis of the relation between accumulation of chemical

intermediates and N2O emission was performed on a full-scale OLAND-type reactor treating

sludge digestate. This study showed that NH2OH and NO2- were both precursors for increased

N2O emission. These intermediates were mostly formed during transient conditions, regarding

the input of oxygen and ammonium (Chapter 3). The uncoupling of AerAOB with NOB,

AnAOB or heterotrophic denitrifiers can also occur in less transient conditions, for example

by inhibition of one of the above groups. Thus, the inhibitory effect of NO towards NOB


162

could have been responsible for the increased nitrite accumulation during mainstream

treatment and therefore also for the increased N2O emission (Chapter 7).

A first strategy to decrease N2O emission, as suggested in Chapter 1, could consist of

performing OLAND at stable operational conditions, which allow constant specific microbial

activities and therefore avoiding accumulation of NO2- and NH2OH. However, stable

operational conditions, for example by applying constant aeration instead of intermittent

aeration, do not always generate lower N2O emission (Joss et al., 2009). Moreover, it should

be noted that a certain accumulation of nitrite can be needed to channel nitrite into the

anammox or nitrite oxidation route as the corresponding microbial groups have affinity

constants of 0.05 and 5.5 mg N L-1

, respectively (Lackner et al., 2008). Therefore, monitoring

of the precursors, NO2- and NH2OH, of N2O emission could be necessary. However, NH2OH

concentrations detected in OLAND systems are always very low, which makes it difficult to

evaluate changes. Moreover, for mainstream conditions which work at lower nitrogen

concentrations, the monitoring of NH2OH will not be reliable enough. Compared to NH2OH,

NO2- concentration differences are easier to detect as they occur in higher concentrations.

However, the on-line NO2- probes on the market still need further optimization to allow

reliable long-term measurement.

During a first full-scale trial to implement OLAND in the mainstream of the WWTP of Strass

(Austria), an increased nitrite accumulation (up to 9 mg N L-1

) was observed. This was

especially the case in wintertime when the loading rate of the WWTP significantly increased

due to the tourist season (Wett et al., 2012). A shift in the effluent value of the nitrite over

nitrate levels was observed reaching values above 1, which however also lead to increased

levels of NO (0.0004-0.03%) and N2O (2-9%). N2O emission in this full-scale system was

mainly steered by the degree of nitrite accumulation and not by the transition from anoxic to

oxic conditions. This was indicated by the increased levels of N2O emissions at higher DO

concentrations from 1 to 3 mg O2 L-1

(Fig. 9.1) and concomitantly increased nitrite levels

from 0.5 at the lowest DO set point up to 4 mg N L-1

at the highest set point. Due to the

unreliable online measurement systems for nitrite in practice, control through an operational

parameter, which is strongly linked with nitrite is advisable. The latter is for example done in

the DEMON process, which is based on pH decrease caused by the oxidation of ammonium

to nitrite.

Chapter 9

163

Another option is to control the aeration system based on the on-line measurement of gaseous

NO2 as this parameter is directly determined by NO2- and is easy to measure in the gas phase.

A drawback could be that depending on the type of wastewater or type of sludge used, the

relation between the NO2 emitted and the NO2- concentration in the liquid phase could differ

(Weissenbacher et al., 2007). Moreover, NO2 levels tend to be very dynamic (Fig. 9.1) and

from the moment NO2 is emitted, already significant levels of NO2- are present in the system,

which could have already caused increased N2O emissions (Weissenbacher et al., 2010).

Therefore, a control strategy based on this parameter should be further explored.

NO can be seen as a universal N2O precursor as this compound is always formed before N2O

is emitted. Moreover due to the low solubility of NO, this compound could give a faster

indication of accumulation of intermediates, which are difficult to detect. As NO could

control the microbial balance under mainstream conditions (Chapter 7), online measurement

and aeration control through the measurement of NO could probably allow a good microbial

balance and avoid excessive NO2- levels and as a consequence excessive N2O emissions.

Further long-term measurement of NO at WWTP with mainstream OLAND is necessary to

find a relation between NO emission, performance and N2O emission and to select threshold

concentrations for proper control.

Figure 9.1: Emission of NO, N2O and NO2 in relation to the oxygen set point tested in the B-stage of

the WWTP of Strass (Austria) inoculated with DEMON granules (preliminary results). Detection limit

of the NO measurement was 1 ppm. Higher NO concentrations were therefore set at 1000 ppb.


164

2.3 Mitigation strategies which minimize emission

It is important to distinguish N2O formation from N2O emissions, which is a physical

mechanism governed by stripping in aerated parts of the system and by passive diffusion,

mixing and wind advection in non-aerated compartments. Because of the difference, limiting

the transfer of the formed N2O to the atmosphere can lower the overall emissions of nitrogen

removing plants. For systems with active aeration, minimization of the airflow rate could

lower N2O emissions (Kampschreur et al., 2008; Kampschreur et al., 2009a). Other factors

that could influence the physical transfer of N2O from the water to the gas phase are the

aeration system itself (size of bubbles) and the aeration control (avoidance of overaeration). In

Figure 3.3 (Chapter 3) an example has been given of the establishment of full stripping of

N2O. A lag phase between N2O stripping and CO2 and NO emission was observed due to the

difference in solubility and the stepwise formation of N2O from NO. The length of the

aeration periods (N2O formation) as well as the length of the anoxic periods (N2O

consumption) could therefore also play a role in the overall degree of N2O emission.

Moreover, it was shown that bubbleless aeration systems such as membrane-aerated systems

could lead to a 100-fold decrease in N2O emissions (Pellicer-Nacher et al., 2010). Similarly,

systems based on passive aeration such as RBC systems, are believed to emit less N2O than

systems with active aeration because of the lower kLa (Desloover et al., 2011b). However,

data on biofilm-based systems are limited and the high N2O emission measured in the RBC of

Chapter 7, did not confirm this assumption.

As NO and N2O will always be present at a certain level in systems based on nitritation, better

understanding of the parameters influencing the mass transfer of N2O from the liquid to the

gas phase will allow further optimization of the overall greenhouse gas emission of the

WWTP.

3 Energy positive WWTP: reality or fantasy?

3.1 Water-energy nexus

Water and energy are intertwined. Water is needed for energy production to power the

turbines in hydro-electric facilities, for cooling in thermal or nuclear energy plants, and to

extract oil from tar sands etc. Energy is needed to pump, treat and heat water, to generate

steam for urban, industrial and agricultural use and to deal with the resulting wastes (Table

9.1). Moreover, the water-energy nexus is deeply connected with the climate change. Burning

Chapter 9

165

fossil fuels, water transport through sewers and wastewater treatment systems all contribute to

the emission of greenhouse gases and therefore add their part to the global warming. As the

latter causes an increased rate of evaporation, variability in precipitation and a greater demand

for cooling, the climate change, which is mainly created by energy use, is strongly

experienced through the water cycle (Lazarova et al., 2012).

In a time of climate change and global warming, a need for a holistic approach, which also

integrates the growing urban development, is advisable to manage water and energy, along

with nutrients. Therefore, closing the water and energy cycles could be a step forward in

decreasing the use of resources. Advanced wastewater treatment methods, mainly based on

micro- and ultrafiltration are needed to reuse water, but the latter also requires more energy

(Table 9.1). Nevertheless, alternatives such as desalination are still not competitive enough in

terms of energy and costs (Table 9.1). The most energy efficient desalination plant (Ashkelon,

Israel) requires an energy consumption of 2.9 kWh m-3

water produced (Voutchkov, 2010),

which is still 6 times higher than the application of advanced water reuse for the same

treatment capacity (Mehul and Dunvin, 2010). Within the field of wastewater treatment, one

can try to maximize water reuse and therefore treat the water only towards a specific purpose.

It was already suggested to reuse for example grey water from the washing machines and

bathing, after a limited treatment in a decentralized system as toilet flushing water (Bieker et

al., 2010). On the other hand, production of potable water from conventional activated sludge

treatment effluent is economically and technically feasible through a multiple barrier

approach using microfiltration, reverse osmosis and UV disinfection methods. Several

examples of the latter are operational in Singapore (PUB, 2010), Belgium (Dewettinck et al.,

2001) and California (OCWD, 2009).

Table 9.1: Water footprint for energy production and energy footprint for water elements of the water

cycle (Lazarova et al., 2012).

Water for energy Energy for water

Energy source m3 MWh

-1 Elements of water cycle kWh m

-3

Gas 0.38 Potable water treatment 0.2 -1.5

Nuclear 0.38 Potable water distribution 0.05 – 0.24

Coal 0.72 Preliminary treatment of wastewater 0.16 – 0.3

Solar thermal 1.1 Activated sludge system (AS) 0.25 – 0.6

Crude oil 4.0 AS with nitrification 0.3 – 1.4

Hydropower 250 Water reuse 0.2 – 2.5

Biogas from crops 600 Brackish water desalination 1 – 1.5

Biodiesel from crops 1130 Seawater desalination 2.5 - 5


166

Besides water reuse from WWTP, also energy recovery can be obtained. As discussed in

Chapter 4, wastewater has a high energy content in the form of heat and organic carbon.

Enhanced reuse of the energy contained in wastewater is another possibility to improve the

water-energy nexus. Maximization of the energy recovery by anaerobic digestion and

minimization of the energy consumption by autotrophic nitrogen removal (Chapter 4), can

therefore lead to energy-neutral or even energy-positive wastewater treatment plants. How

one can reach energy self-sufficient systems is discussed in the following sections.

3.2 Is OLAND an essential treatment step?

3.2.1 Yes it is, to allow maximum energy recovery

Implementation of OLAND in the municipal wastewater treatment chain, allows high

nitrogen conversion rates without the need for organic carbon. Therefore, to allow maximum

recovery of organics from sewage and obtain a dischargeable effluent quality, implementation

of OLAND is needed. The application of OLAND in the sidestream for the treatment of the

digestate, has already shown to be reliable, robust and highly efficient (Table 1.4, Chapter 1).

Both by energy calculations (Chapter 4, Siegrist et al., 2008) and full-scale experiences (Wett

et al., 2007), it was shown that for sidestream OLAND, around 50% of the aeration

requirements of the plant could be saved compared to CAS. It should however be noted that

the net energy decrease was mainly caused by the higher recovery of organic carbon by

anaerobic digestion, due to the lower COD removal needed in the mainstream (Chapter 4).

Conditions for sidestream OLAND provide easy control of the microbial balance due to the

high temperatures and high nitrogen concentrations, which allow prevention of nitrate

production by FA inhibition. Therefore, it is possible to guarantee high nitrogen removal

efficiencies for the treatment of digestates (Table 1.4, Chapter 1). Also for industrial and

decentral treatment of digestates with OLAND, the energy index significantly increased

(Chapter 4). Moreover, stable and highly efficient performance is already demonstrated in

practice. Therefore, the implementation of OLAND for the treatment of digestates is

advisable to decrease the overall energy consumption and allow higher organic carbon

recovery.

For mainstream conditions, our lab-scale reactor tests (Chapter 6 and 7) showed that high

nitrogen removal rates could be obtained at mainstream conditions (low temperatures, low

nitrogen, COD/N of 2). However, competition between AnAOB and NOB in these conditions

is more difficult to control compared to the conventional techniques based on FA, FNA and

Chapter 9

167

oxygen used in sidestream treatment. It was suggested that NOB suppression by NO

concentration could be an alternative strategy as this compound stimulated AnAOB over

NOB activity (Chapter 7). However, NO also decreased the AerAOB activity, which

indicated that a balance between obtaining enough nitritation without substantial nitratation is

essential to allow high nitrogen removal efficiencies. A first full-scale approach (Strass,

Austria; Chapter 8) in which a combination of OLAND and nitrification/denitrification was

applied, has experimentally examined several operational conditions, mainly related to the

input of oxygen (data not shown). These first primary tests showed that NOB easily adapted

to lower DO conditions in the reactor, and therefore were strong competitors for nitrite

compared to AnAOB. However, when high loading rates (winter time) and/or DO set points

of 1-2 mg O2 L-1

were applied, high NO/N2O emission occurred together with nitrite

accumulation, showing a decrease in NOB activity. These primary results at full-scale

therefore correlate very well with the lab-scale reactor tests shown in this work (Chapter 7).

Further research is needed to elucidate the NO concentration needed to suppress NOB and

evaluate the sustainability of this application compared to conventional treatment. Moreover,

mainstream OLAND should first show reliability, sustainability and energy efficiency at

larger scale, before conclusions can be made about the need for implementation to allow

energy-positive treatment.

3.2.2 No it is not, other adjustments can help

Improving wastewater treatment performance is and should be the primary objective of a

WWTP. After obtaining stable discharge limits for the effluent, the best available practices

and technologies for enhanced energy efficiency and the best use of sludge for energy

production and recovery can be investigated and implemented. OLAND allows for lower

oxygen consumption and thus a lower aeration rate compared to conventional

nitrification/denitrification (Kuai and Verstraete, 1998). However, energy calculations

revealed that also oxygen transfer efficiency, digestibility of sludge and efficiency of

anaerobic digestion could significantly influence the energy balance (Chapter 4, BOX 5).

As aeration accounts for 60-70% of the energy demand of a WWTP (Zessner et al., 2010),

optimization in this area can save up to 20% of the energy consumption (Lazarova et al.,

2012). Aeration systems can achieve oxygen transfer efficiencies above 2 kg O2 kWh-1

(Nowak et al., 2011) and the use of premium efficiency motors and variable frequency drivers

for large pumps and aeration blowers can also limit the total energy consumption for the same


168

amount of oxygen input. Moreover, efficient aeration control systems, for example based on

the measurement of ammonium in the effluent, can further optimize energy input.

In addition to energy input for aeration, the efficiency of energy recovery from sludge can

influence the energy balance significantly (Chapter 4, Box 5). As primary sludge is easier to

digest than secondary sludge, increasing the primary sludge production can also improve the

energy recovery. This can for example explain the difference between the CAS and A60/B

system (Chapter 4, Table 4.2). Next to this, the primary and secondary sludge mixture can be

pretreated or codigestion can be applied to further increase the methane yield (Carrere et al.,

2011). The CHP-units operational at present have an electrical efficiency of 31-38% (Nowak

et al., 2011). Therefore, the choice of the CHP unit or the choice of the method to use the

biogas (through electricity, via gasification or through direct mechanical energy) can also

influence the energy gain obtained. Together, these adjustments can lead to 20-40% increased

energy recovery (Lazarova et al., 2012).

Furthermore, depending on the climate and transport distances, energy can also be gained

from the sewage flows themselves by hydro-turbines and heat pumps (Verstraete and

Vlaeminck, 2011). This can increase energy recovery up to 10% (Lazarova et al., 2012).

Another 10% increase in energy recovery can be obtained by the production of renewable

energy from external sources such as solar, wind or geothermal energy (Lazarova et al.,

2012).

Together, these adjustments can also lead to energy autarky, even for CAS systems. This was

for example shown in the Wolfgangsee-Ischl WWTP (Austria), which treats 40 000 IE and is

based on a singly stage AS-system with primary sedimentation and anaerobic sludge digestion

(Nowak et al., 2011). From 2009 onwards, energy self-sufficiency was reached, by

optimization of the aeration system and control (2.3 kg O2 kWh-1

), increasing primary

sedimentation (37%), optimization of the digesters (2 in series with a SRT of 80 days) and

implementation of a better CHP unit (electrical efficiency of 34%). Therefore, this full-scale

example shows that OLAND is not essential in all cases to obtain energy self-sufficient

WWTP.

Chapter 9

169

Figure 9.2: Overview of the degree of OLAND implementation and oxygen transfer efficiency needed

(A: 1 kg O2 kWh-1

; B: 2 kg O2 kWh-1

) to allow energy-positive WWTP in function of the primary

sludge production efficiency and COD/N of the incoming sewage. Grey: energy-negative;

yellow: energy-positive without OLAND implementation; orange: energy-positive if OLAND is

implemented in the side stream; red: energy-positive if OLAND is implemented in the meanstream.

Primary sludge production higher than 75% is considered as technically not feasible at the moment

(light grey boxes). Other parameters such as digestibility of the sludge, growth yield etc were kept at

default values (Chapter 4, BOX 1).


170

3.3 Decision making for the wastewater engineer

It is clear that OLAND is not essential in all cases to obtain energy neutral or even energy-

positive wastewater treatment. Several application domains have other needs and some

general guidelines are subsequently proposed to help decide if OLAND implementation can

be necessary to achieve an energy-positive WWTP.

For the treatment of highly-loaded organic streams, which can be immediately subjected to

biogas digestion, OLAND treatment of the resulting digestate can have a high impact on the

energy and cost balance. Examples were shown in Chapter 4 for the treatment of the

OFMSW, agro-industrial waste and manure-based organics. Due the high digestibility of the

first two streams, high energy recoveries could be obtained and OLAND could increase the

energy index with a factor 2. However, energy-positive treatment was also obtained when

conventional nitrification/denitrification was applied. However, the latter had higher needs for

external carbon addition to meet discharge limits. Therefore, in treatment schemes with

anaerobic digestion of agro-industrial waste and OFMSW, OLAND implementation is

advisable. OLAND implementation for the treatment of manure-based organics seemed more

difficult and the effect on the energy balance was therefore minor. In this field of application,

treatment of the gaseous ammonia streams by OLAND showed a better potential for

application (Chapter 5).

For a municipal WWTP due to its complexity, it is more difficult to estimate if energy-

positive wastewater treatment is possible and which prerequisites will determine the energy

balance of the WWTP. It was suggested by Nowak and colleagues (2011) that energy autarky

should be achievable for WWTP removing at least 70% nitrogen and treating sewage with

COD/N ratios > 10 (Nowak et al., 2011). To test this proposal, some additional calculations

were made based on the assumptions used in Chapter 4 (Fig. 9.2). It was shown that indeed

the COD/N ratio of the sewage together with the degree of primary sludge production could

affect the degree of OLAND implementation needed to obtain energy autarky. The higher the

COD/N ratio of the sewage, the more easily energy-positive treatment is obtained, because a

higher proportion of primary sludge can be separated without influencing the efficiency of the

conventional nitrification/denitrification (Fig. 9.2). OLAND implementation in the sidestream

becomes more important at sewage COD/N ratios between above 8-14 (Fig. 9.2). At lower

COD/N ratios, mainstream OLAND together with substantial primary sludge production is

Chapter 9

171

needed to decrease energy consumption and allow optimal recovery (Fig. 9.2B). Besides the

sewage COD/N ratio, the oxygen transfer efficicieny for aeration in both side and mainstream

treatment plays a crucial role. In case of inefficient aeration (1 kg O2 kWh-1

), energy autarky

is very difficult and can only be achieved with high primary sludge productions (>65%) and

high COD/N ratios (>10-13) in the sewage (Fig. 9.2). The aeration systems with higher

oxygen transfer efficiencies (2 kg O2 kWh-1

), create a higher possibility to attain energy

autarky (Fig. 9.2B). For municipal WWTP, energy-positive treatment is only possible when

during primary settling more than 50% of incoming COD is removed with the primary sludge.

It should be however noted that the latter assumes a default anaerobic digestion efficiency

while further improvements in this step are still conceivable (Chapter 4). A higher methane

yield can further shift the pattern towards energy-positive treatments even at lower primary

sludge productions, which was for example the case in the Wolfgangsee-Ischl WWTP

(Austria) (Nowak et al., 2011).

4 Nitrogen removal versus nitrogen recovery

Nowadays, the fertilizer industry is based on the Haber Bosh process, which catalytically

combines hydrogen and nitrogen gas to ammonia (N2 + 3 H2 2 NH3) under high pressure

(15-25 MPa) and high temperatures (300-550°C)(Chagas, 2007). The specific operational

conditions make this process energy intensive (Table 9.2). In total, this process accounts for a

nitrogen fixation rate of 120 106 tons N year

-1 and is expected to increase up to 165 10

6 tons N

year-1

by 2050 (Galloway et al., 2004). As a consequence, the occurrence in the environment

of reactive nitrogen compounds such as ammonium, nitrite and nitrate is sharply increasing. It

was estimated that the global population discharges around 20 106 tons N year

-1 in wastewater

of which 99% of this reactive nitrogen is not treated and released as such in the environment

(Galloway et al., 2008). The increasing amount of regulations and the increasing number of

WWTP with nutrient removal is a first step towards decreasing the environmental problems

of this excess of reactive nitrogen compounds. During our research, the focus was always on

nitrogen removal and the aim to remove nitrogen in a cost effective and energy friendly way.

However as resources are decreasing, recovery of nutrients will become necessary.

Nitrogen recovery can be obtained by several physico-chemical methods of which ammonia

stripping and struvite precipitation are the most common ones (Siegrist, 1996). When

ammonia stripping is applied an acidic salt NH4(SO4)2 is formed which can be concentrated


172

after drying. The nitrogen can also be recovered as (NH4)2CO3. To obtain high recovery

efficiencies, high nitrogen concentration levels, high pH values and high temperatures are

advisable as these factors shift the ammonium balance to ammonia. During struvite

precipitation an insoluble magnesium ammonium phosphate (MgNH4PO4.6H2O) is formed

under alkaline conditions (pH 8.5 – 10). A N/P ratio above 1 is needed and enough Mg-ions

have to present in the water to avoid external addition of phosphate or magnesiums salts,

respectively. Both NH4(SO4)2 and struvite can be used as fertilizers and therefore have the

potential to decrease the need to perform the energy intensive Haber-Bosh process. However,

at this moment the price of fertilizers made through the Haber-Bosh process are too low to

give enough driving force towards nitrogen recovery (Table 9.2). As energy prices are rising,

the price of the fertilizers will follow this trend, and the economical discrepancy between

nitrogen removal and nitrogen recovery will decrease in the future. At this moment nitrogen

recovery is only a cost-efficient option if streams contain more than 5 g N L-1

(Mulder, 2003).

The latter is for example the case for urine (Larsen and Gujer, 1996).

Table 9.2: Comparison of the maximum fertilizers market prices (Apodaca, 2007), converted at

1.4 USD EUR−1

, and the costs for nitrogen recovery (Siegrist, 1996; personal communication

Colsen nv) and removal (Fux and Siegrist, 2004) from sludge digestates.

Cost

(€ kg-1

N)

Energy

(kWh kg-1

N)

Nitrogen removal Nitrification/denitrification 1.79 2

OLAND 0.29 1

Nitrogen recovery Ammonia stripping 2-10 6-25

Struvite precipitation 13 28

Fertilizer production Haber-Bosh process 0.37 9-12

Anhydrous ammonia 0.59

Ammonium nitrate 0.87

Ammonium sulphate 0.93

Urea 0.83

A combination of nitrogen recovery and nitrogen removal can also be attractive. In this way

the nitrogen recovery efficiency only goes to the point where it is still economically attractive.

The last part of the nitrogen, which is the most difficult to recover could then be removed by

OLAND or nitrification/denitrification, depending on the COD/N ratio. Especially, after a

thermophilic anaerobic digestion step, nitrogen recovery through ammonia stripping could be

attractive as in this way the heat is more efficiently used compared to direct biological

treatment of thermophilic digestates which have to be cooled first. Full-scale tests revealed

that the treatment cost for ammonia stripping of thermophilic digestates with nitrogen

Chapter 9

173

concentrations around 2 g N L-1

, could already be decreased to 2 EUR kg-1

N recovered

(personal communication Colsen nv). This is only a factor 2 difference compared to a final

treatment scheme based on nitrogen removal, which can meet discharge limits. Therefore,

further optimizations of the operational conditions for ammonia stripping, especially

regarding the stripping mechanism itself, will allow a cost-effective combination of nitrogen

recovery and removal in the future.

Besides the economical aspect, from a LCA point of view, nutrient recovery can positively

affect impact categories such as the eutrophication potential, abiotic depletion potential,

global warming potential and ecotoxicity potential (Chapter 8). Therefore, nutrient recovery

can increase the overall sustainability of the WWTP. The latter can also be obtained by for

example composting of waste sludge (Chapter 8). However, the latter does not allow the

production of high purity products and therefore tends to decrease the value of the product.

5 Future challenges and opportunities

5.1 Future challenges for mainstream OLAND

Decreasing the nitratation activity at mainstream conditions and thus allowing sufficient

AnAOB activity will be the main challenge for the future large-scale implementation of

OLAND. One of the research lines that could be followed, as already highlighted before

(Chapter 9, section 2.2) is the optimization of a control system based on NO measurement or

a parameter closely related with NO. This strategy will therefore profit from the higher

sensitivity of NOB for NO, compared to AerAOB and AnAOB (Chapter 7). Another

possibility could be to install a small breeding reactor, which treats a highly-loaded nitrogen

stream (digestate), but at low temperature (for example max 20°C). By recirculating the

OLAND biomass from the mainstream to the breeding reactor, inhibition of NOB can be

established in this breeding reactor without complicated control strategies needed in the

mainstream. It was for example shown that the NOB in the mainstream biomass, which were

grown at low FA concentration, were easily inhibited by this compound (Chapter 7, Table

7.3). The main challenge in the latter case is to find an elegant way to easily circulate

OLAND biomass without the activated sludge and to determine the SRT of the sludge needed

in both steps to obtain an optimal mainstream performance without long-term adaptation to

the inhibitory factors in the breeding reactor. As during mainstream treatement a separation

between the SRT of AnAOB containing particles and aerobic flocs is advisable to maintain


174

AnAOB activity in the system (Wett et al., 2010), these separation systems could be

connected to the breeding reactor. Until now, one system based on cyclones was proposed

(Wett et al., 2012). However, the inoculation of the B-stage with OLAND biomass on e.g. a

floating carrier, which can easily be retained in the system by grids, could be another

technical solution. Besides, optimization of control mechanisms and suppression of NOB

externally, the performance of mainstream OLAND could also be improved by a better

combination of OLAND with nitrification/denitrification (Chapter 7). Implementation of

predenitrification – OLAND or OLAND – postdenitrification are some options to guarantee

stable discharge limits and to better counteract fluctuations. The latter would for example

allow retrofitting of existing activated sludge systems.

5.2 OLAND biofilter application

In Chapter 5, it was shown that nitrogen removal directly from the gas phase was possible and

that high nitrogen volumetric removal rates and efficiencies were obtained in an OLAND

biofilter. As in practice ammonia containing gaseous wastestreams will rarely contain only

ammonia, in most cases a combination of ammonia with sulfide (H2S) (Malhautier et al.,

2003) and/or ammonia with volatile organic compounds (VOC) (Cabrol et al., 2009) will

have to be treated. Therefore, the combination of ammonia removal through OLAND with

sulfide and VOC removal is the challenge for successful implementation in practice.

Sulfide can be inhibitory for AnAOB at levels around 10 mg S L-1

(Dapena-Mora et al.,

2007). However, in an autotrophic denitrifying reactor based on H2S oxidation AnAOB were

detected (Mulder et al., 1995) and other batch tests also showed that AnAOB could resist H2S

concentrations of at least 64 mg S L-1

(van de Graaf et al., 1996). Therefore, it should be

possible to cultivate an AnAOB culture that can adapt to higher H2S concentrations. A shift of

the nitrifying activity to lower sections of the biofilters was observed in a biofilter treating

H2S and NH3 (Malhautier et al., 2003), probably making the interference of H2S with the

OLAND process minor. Moreover, in these filters low nitratation activity was observed

leading to nitrite accumulation, which could give an opportunity to the AnAOB to survive in

the system (Malhautier et al., 2003). Besides the shift in activity, also the limitation of NO

emission will be challenging as H2S can chemically react with HNO2 to NO and oxidized

sulfur compounds (Vermeiren et al., 2012).

Chapter 9

175

Emission of VOCs such as volatile fatty acids, ketones, aldehydes and alcohols are frequently

associated with composting (Cabrol and Malhautier, 2011). The VOC concentration observed

in these systems is mostly below 1 mg m-3

, compared to NH3-N and H2S-S concentrations of

on average 30 mg m-3

(Cabrol et al., 2009). Moreover, the solubility of these compounds

differs and this will determine the contact with the OLAND biomass. Stable simultaneous

VOC removal and nitrification was already shown in biofilters (Sakano and Kerkhof, 1998;

von Keitz et al., 1999; Friedrich et al., 2003; Cabrol et al., 2009). To increase the total

nitrogen removal in these systems, and therefore to apply OLAND, the effect of the VOCs

composition and concentration on the AnAOB activity should be examined carefully.

5.3 What are the temperature limits of the OLAND process

This doctoral research showed that OLAND can be performed at low temperature (15°C) in

contrast to the main mesophilic (30-35°C) research domain (Chapter 7, Chapter 1, Table 1.4).

As AnAOB species were found in thermophilic conditions in nature, such as hot springs

(Byrne et al., 2009) and high temperature petroleum reservoirs (Li et al., 2011), AnAOB

activity should also be possible at thermophilic conditions. Thermophilic environmental

biotechnology is established for carbon treatment (Wiegel and Ljungdahl, 1986). However,

thermophilic nitrogen removal processes are not developed yet, although several types of

nitrogenous wastewaters have temperatures above the mesophilic range. Most progress has

been obtained in thermophilic denitrification (Laurino and Sineriz, 1991). However, for

nitritation, nitratation and anammox no successful reports are found for thermophilic

conditions (>40°C).

From more fundamental work however, thermophilic ammonium-oxidizing Archaea (AOA)

were isolated and cultured (Hatzenpichler et al., 2008). Moreover, enrichments of

thermophilic AerAOB and NOB were obtained (Lebedeva et al., 2005; Lebedeva et al., 2011;

Shimaya and Hashimoto, 2011). It should therefore be possible to cultivate a nitrifying culture

(AerAOB-NOB or AOA-NOB) and couple this to the existing thermophilic denitrification

process (Laurino and Sineriz, 1991). Moreover, also in this case combining nitritation by

AerAOB or AOA with thermophilic adapted AnAOB would allow a new application domain

for OLAND.

For this application domain, the same strategy as for low temperature application (Chapter 7)

can be applied e.g. gradual temperature adaptation. However, it could be needed to cultivate


176

some adapted species, to mix and inoculate afterwards, as high temperatures will probably

require thermostable enzymes (Haki and Rakshit, 2003). The first reactor tests performed on

nitrification showed that a mesophilc consortium could only survive for 1 week at 45°C,

although stable performance at 40°C was feasible (Shore et al., 2012). Therefore, this could

suggest that specialist groups of AerAOB, AOA and/or NOB are needed.

Until now, at temperatures above 40°C, nitrogen loss is mainly caused by ammonia stripping

and not by ammonium conversions (Abeynayaka and Visvanathan, 2011), which is not

sustainable. Thermophilic nitrogen removal could however offer several advantages. First of

all energy costs for cooling can be saved in contrast to treatment of thermophilic effluents

from several industries such as pulp and paper manufacturing (Suvilampi et al., 2001) or from

effluents of thermophilic activated sludge or anaerobic digestion, which at this moment first

needs a cooling step before treatment is applied. Moreover, because of the high FA

concentration and lower oxygen solubility at these high temperatures, suppression of

nitratation and therefore control of the OLAND process will be easier. The challenges for this

application domain will mainly be related with the search for a good consortium and with

limiting the ammonia stripping.

6 Conclusions

Autotrophic nitrogen removal based on partial nitritation/anammox, as performed in a one-

stage treatment during the OLAND process, can significantly decrease the energy

consumption, CO2 emission, sludge production and the needs for an external organic carbon

source compared to conventional nitrification/denitrification. Due to these main advantages,

the application of OLAND for the treatment of digestates can be seen as an established

technology. Recently, some 44 full-scale plants using this process are reported. The results of

this work for the mesophilic application domain showed that for the rapid start-up and high

performance of OLAND SBR systems, stable hydraulic conditions (low volumetric exchange

ratio) were needed. Moreover, to allow a sustainable process and thus low N2O emission,

accumulation of NH2OH and NO2- should be avoided. In addition, energy calculations

showed that new potential domains for OLAND were located (1) in agricultural application

requiring ammonia removal and (2) in municipal WWTP using mainstream treatment.

Chapter 9

177

This work constitutes an attempt to contribute to the application of OLAND by the following

key accomplishments:

For gas treatment, containing ammonia:

o High removal rates (0.7 kg N m-3

d-1

) and high removal efficiencies (75-80%)

were obtained at a pilot scale (height 1.6 m) OLAND biofilter.

o AnAOB activity and presence was obtained in the OLAND biofilter

demonstrating the contribution of AnAOB to the ammonia removal process.

For mainstream water treatment, containing ammonium:

o High total nitrogen removal rates (0.5 kg N m-3

d-1

) were obtained at 15°C,

nitrogen concentrations of 55 mg N L-1

and COD/N ratios of 2.

o An alternative strategy of nitratation suppression at mainstream conditions

based on NO was proposed.


178

179

180

Sludge from B-stage (WWTP Strass, Austria)

Abstract

181

Abstract

Several new biological nitrogen removal processes, which are based on partial

nitritation/anammox, have been developed to treat nitrogen-rich wastewaters devoid in carbon

such as digestates. Around 40 full-scale realizations of one-stage partial nitritation/anammox,

in this work referred to as the oxygen-limited autotrophic nitrification/denitrification

(OLAND) process, are operational at this moment for high strength nitrogen streams.

OLAND is based on partial nitritation, performed by aerobic ammonium-oxidizing bacteria

(AerAOB) and anammox, performed by anoxic ammonium-oxidizing bacteria (AnAOB). The

AerAOB, mainly belonging to Nitrosomonas europaea eutropha and halophila, are set so that

they oxidize half of the influent ammonium to nitrite in oxygen-limited conditions. The

AnAOB, mainly members of the Candidatus genera Kuenenia and Brocadia, oxidize the

residual ammonium with nitrite to dinitrogen gas under anoxic conditions. Consequently, in

the OLAND process ammonium is converted mainly into nitrogen gas without the use of

organic carbon in one reactor. Overall OLAND can save 84% of the operational costs, by a

100, 89 and 57% decrease in methanol requirement, sludge production and aeration,

respectively.

The close interaction between the different microbial groups during the OLAND process is

comparable with human beings working together in firms for a shared profit. In this sense, the

concept of human resource management (HRM) was translated to the microbial

biotechnology as Microbial Resource Management (MRM) and therefore strives after

maintaining the best performing microbial community for a certain application. A MRM

OLAND framework was elaborated (Chapter 1), showing how the OLAND

engineer/operator (1: input) can design/steer the microbial community (2: biocatalyst) to

obtain optimal functionality (3: output), depending on the application domain (0: wastewater).

Taken this MRM framework into account, the OLAND engineer can steer the OLAND

process to obtain maximum efficiency and higher sustainability or to increase the impact of

OLAND on the energy balance of wastewater treatment plants (WWTP).

Although the first OLAND applications have shown that this technology works in a stable and

efficient way, the implementation rate of this technology remains dependent on a few

Abstract

182

companies. Many potential users hold back because it seems that due to the long start-up

periods for the first reactors and the reported sensitivities, a lot of experience is needed to

keep this process running. To overcome this problem, the output box of the MRM framework

was further studied in detail for high-strength nitrogen containing wastewaters (known

application). Firstly, the effect of the hydraulic conditions on the start-up of the OLAND

sequencing batch reactor (SBR) was examined. Low volumetric exchange ratios, which

assure stable hydraulic conditions, were needed to allow a fast start-up, granulation and high

performance in SBR systems (Chapter 2). Furthermore, strategies to obtain a well-balanced

OLAND system, were proposed based on wash-out of nitrite oxidizing bacteria (NOB)

through selection on settling velocity or by stimulation of AnAOB through the

implementation of an anoxic phase. As not only the effluent quality, but also the sustainability

can be a competitive factor to choose an environmental technology, the N2O and NO

emissions were studied in a full-scale OLAND-type reactor (Chapter 3). The sustainability of

the process in terms of NO/N2O emissions was mainly linked with accumulations of

intermediates such as NO2- and NH2OH and the frequency of transient conditions. Better

understanding of the conditions which lead to the accumulation of intermediates and further

optimization of the feeding pattern which determines the degree of fluctuations, will allow a

further decrease of the N2O emission in these systems.

In a next part of this work, new opportunities for OLAND, which could improve the overall

sustainability of the applied processes, were explored (box 0 of MRM framework). Energy

calculations revealed that OLAND treatment of digestates could significantly increase the

energy index of agro-industrial and organic fraction of municipal solid waste-based treatment

system from 3-5 to 6-10 (Chapter 4). However, for manure-based digestate treatment,

OLAND application seemed more difficult and therefore ammonia gas treatment by OLAND

was suggested for this application domain. A pilot-scale OLAND biofilter fed with a flow of

ammonia gas, obtained a high performance (0.7 g N L-1

d-1

) and a high total nitrogen removal

efficiency (75-80%; Chapter 5). Although the filter was saturated with oxygen, the low

relative water flow rate ratio (≈1 L g-1

Nin) ensured high free ammonia concentration in the

water phase, which resulted in a dominance of AnAOB compared to NOB activity at the top

of the biofilter.

A specific application domain, which could particularly improve the energy efficiency of

sewage treatment plants was the implementation of OLAND in the mainstream of the system.

Abstract

183

This would allow a net electrical energy production, due to a higher carbon recovery and

lower energy needs for aeration (Chapter 4). Four challenges to allow mainstream OLAND

were encountered. A first challenge, namely the performance of OLAND at low nitrogen

concentration and low hydraulic residence time (HRT) was shown in an OLAND rotating

biological contactor (RBC; Chapter 6). The reactor obtained high nitrogen removal rates (0.4

g N L-1

d-1

) treating nitrogen concentration of 30-60 mg N L-1

at a HRT of 1-2 hours. A

second challenge, operation at low temperatures (15°C), was surmounted in the same RBC by

gradually decreasing the temperature starting from 29°C. During operation at 15°C with

synthetic feed (60 mg N L-1

) and a HRT of 1h, a similar nitrogen removal rate as at high

temperatures was obtained i.e. 0.5 g N L-1

d-1

(Chapter 7). Compared to higher temperatures

only a decrease of the total removal efficiency of 22% was detected. The switch from

synthetic feed to pretreated sewage with a COD/N ratio of 2 (challenge 3) did not

significantly affect the performance. However, during the low temperature performance of the

RBC system, NOB activity started to increase, as well as competition between AnAOB and

NOB for nitrite (challenge 4). It was shown that increased levels of NO selectively enhanced

AnAOB over NOB activity (Chapter 7). Therefore, high peak loading rates together with

nitrite accumulation, increasing the NO production, enhanced the overall removal efficiency.

To evaluate the mainstream OLAND application in a broader context, a life cycle assessment

(LCA) was performed on full-scale data of the WWTP in Strass, which applied an OLAND-

type of system, referred to as DEMON. Three scenarios were studied: (1) the WWTP without

a DEMON system; (2) the WWTP with DEMON in the side line; (3) the WWTP with

DEMON in the side and main lines (Chapter 8). For the latter scenario, data from a first full-

scale trial were used. The LCA showed that implementation of DEMON in the side line of the

WWTP positively influenced the eutrophication potential, abiotic depletion potential and

global warming potential and therefore resulted in a more sustainable WWTP. The first full-

scale results of DEMON implementation in the mainstream of the WWTP in Strass (Austria)

showed that to obtain the same degree of sustainability compared to the WWTP with

sidestream treatment, the N2O emission (around 2% of N load) in the main line should be

decreased as this compound dominated the global warming potential of the plant with 99%.

N2O emission is mainly related with operational conditions and not with the process itself, it

should therefore be possible to further optimize the emission to around 0.5% of the N load

allowing the same CO2 footprint of the plant in comparison with sidestream OLAND

implementation.

Abstract

184

Generally, this work showed that new potential domains for OLAND were located in

agricultural applications requiring ammonia gas removal and in municipal WWTP using

mainstream treatment. Future tests in these domains will need to evaluate the performance

and overall environmental sustainability at larger scale

Samenvatting

185

Samenvatting

Er werden reeds verschillende nieuwe biologische processen, gebaseerd op partiële

nitritratie/anmmox ontwikkeld voor de behandeling van stikstofrijk afvalwater zoals

bijvoorbeeld digestaten. Op dit moment zijn er ongeveer 44 volle-schaal éénstaps partiële

nitritatie/anammox reactoren, in dit werk ook wel oxygen-limited nitrification/denitrification

(OLAND) genoemd, operationeel voor de behandeling van hoogbelaste stikstofstromen.

OLAND is gebaseerd op partiële nitritatie, uitgevoerd door aerobe ammonium-oxiderende

bacteriën (AerAOB) en anammox, uitgevoerd door anoxische ammonium-oxiderende

bacteriën (AnAOB) in één reactor. De AerAOB, die meestal tot de groep van Nitrosomonas

europaea eutropha en halophila behoren, oxideren de helft van de ammonium tot nitriet

onder zuurstofgelimiteerde omstandigheden. De AnAOB, meestal behorende tot de

Candidatus genera Kuenenia en Brocadia, oxideren de overblijvende ammonium met het

gevormde nitriet tot stikstofgas onder anoxische omstandigheden. Dus, tijdens het OLAND

proces wordt ammonium in één reactor omgezet naar stikstofgas zonder gebruik te maken van

een organische koolstofbron. Hierdoor kan het OLAND proces 84% van de operationele

kosten besparen aangezien de behoefte aan externe methanol toediening, de slibproductie en

de energiekost voor beluchting, met respectivelijk 100, 89 en 57% dalen.

De nauwe interactie tussen de verschillende microbiële groepen in het OLAND proces kan

men vergelijken met werknemers, elk met hun specifieke taken, die werken voor de algemene

winst van een bedrijf. In dit perspectief, kan het concept van human resource management

(HRM) ook doorgetrokken worden naar microbiële biotechnologie. Microbial resouce

mangagement (MRM) zal daarom streven naar het onderhouden van de best presterende

microbiële gemeenschap voor een bepaalde toepassing. Een OLAND MRM kader werd

uitgewerkt (Hoofdstuk 1), waarbij getoond werd hoe de OLAND ingenieur/operator (1:

input) een microbiële gemeenschap (2: biokatalysator) kan aansturen om zo een optimale

functionaliteit (3: output) te bekomen, afhankelijk van het toepassingsgebied (0: afvalwater).

Met dit MRM kader in het achterhoofd kan de OLAND ingenieur het OLAND proces

aansturen zodat een maximale efficiëntie en hogere duurzaamheid of een grote impact van

OLAND op de energiebalans van afvalwaterzuiveringssystemen bekomen kan worden.

Samenvatting

186

Hoewel de eerst volle-schaal OLAND toepassingen een stabiele en efficiënte performantie

vertonen, blijft de snelheid waarmee dit proces wordt geïmplementeerd eerder beperkt en is

afhankelijk van een handvol bedrijven. Potentiële gebruikers wachten af omdat het lijkt dat dit

proces door de lange opstarttijden bij de eerste toepassingen en de gepubliceerde

sensitiviteiten veel ervaring vergt. Om dit probleem gedeeltelijk te overbruggen werd de

output box van het MRM kader verder in detail bestudeerd voor hoogbelaste stikstofstromen

(gekende toepassing). Ten eerste werd het effect van de hydraulische condities op de opstart

van OLAND sequencing batch reactoren (SBR) bestudeerd. Een kleine volumetrische

uitwisselingsverhouding, welke stabiele hydraulische condities verzekerde, was essentieel om

een snelle opstart, granulatie en een hoge performantie in SBR systemen te verkrijgen

(Hoofdstuk 2). Verder werden er ook strategieën voorgesteld om een goede microbiële

balans te behouden in het OLAND systeem die enerzijds gebaseerd waren op de uitwassing

van nitriet-oxiderende bacteriën (NOB) door selectie op bezinkingssnelheid en anderzijds

gebaseerd waren op de stimulatie van AnAOB door het invoeren van een anoxische fase.

Naast het behalen van een goede effluent kwaliteit, kan ook de algemene duurzaamheid een

competitieve factor worden tussen verschillende milieutechnologieën. In deze context werden

N2O en NO emissies bestudeerd in een volle-schaal OLAND-type reactor (Hoofdstuk 3). De

broeikasgasemissies waren vooral gelinkt aan accumulaties van intermediairen zoals nitriet en

hydroxylamine en de frequentie van het opleggen van transiënte condities. Het verder

bestuderen van de factoren die leiden tot de accumulatie van intermediairen en het verder

optimaliseren van het voedingspatroon, wat de graad van fluctuaties in de reactor bepaalt, zal

in de toekomst toelaten om de N2O emissies verder te onderdrukken.

In een tweede deel van dit werk werden nieuwe toepassingsmogelijkheden voor OLAND

onderzocht die een positieve invloed zouden hebben op de algemene duurzaamheid van

systemen (box 0 van MRM kader). Energieberekeningen toonden aan dat de energie-index

voor de behandeling van agro-industriële afvalstromen en groente-fruit en tuinafval verhoogd

kan worden van 3-5 tot 6-10 door de implementatie van OLAND voor de behandeling van

digestaten (Hoofdstuk 4). Echter, OLAND implementatie in de verwerking van mest-

gebaseerde afvalstromen lijkt een stuk moeilijker, waardoor voor dit toepassingsdomein de

OLAND behandeling van ammoniakgasstromen werd voorgesteld. Een pilootschaal OLAND

biofilter, gevoed met een ammoniakstroom, behaalde een hoge performantie (0.7 g N L-1

d-1

)

en een hoge stikstofverwijderingsefficiëntie (75-80%, Hoofdstuk 5). Hoewel de filter

gesatureerd was met zuurstof kon de lage relatieve waterstroom (≈1 L g-1

Nin), hoge vrije

Samenvatting

187

ammoniak concentraties in de waterfase verzekeren, welke een dominante activiteit van de

AnAOB over de NOB stimuleerden aan de top van de filter.

Uit verdere energieberekeningen bleek dat door de toepassing van OLAND in de hoofstroom

van rioolwaterzuiveringsinstallaties (RWZI), een netto energiewinst kon bekomen worden

(Hoofdstuk 4). De hoger koolstofrecuperatie en de lagere energiekosten waren hiervoor de

belangrijkste oorzaken. Vier uitdagingen die gepaard gingen met de toepassing van OLAND

in the hoofdstroom van RWZI werden in dit werk overwonnen. De eerste uitdaging, namelijk

OLAND performantie bij lage stikstofconcentraties en lage hydraulische verblijftijd (HRT)

werd aangetoond in een OLAND rotating biological contactor (RBC; Hoofdstuk 6). Er

werden in deze reactor voor de behandeling van lage stikstofconcentraties (30-60 mg N L-1

),

hoge stikstofverwijderingssnelheden (0.4 g N L-1

d-1

) behaald bij een HRT van 1-2 uur. Een

tweede uitdaging, namelijk performantie bij lage temperaturen (15°C) werd overwonnen in

dezelfde RBC door een geleidelijke daling van de temperatuur startende van 29°C. Bij

operatie op 15°C, een HRT van 1 uur en voeding met synthetisch afvalwater (60 mg N L-1

)

werden gelijkaardige stikstofverwijderings-snelheden behaald, namelijk 0.5 g N L-1

d-1

, ten

opzichte van operatie bij hoge temperatuur (Hoofdstuk 7). In vergelijking met de hogere

temperaturen, werd een daling in de stikstofverwijderingsefficiëntie van 22% gedetecteerd.

Overschakeling van synthetisch naar voorbehandeld rioolwater met een COD/N verhouding

van 2 (uitdaging 3) gaf geen significant verschil in de performantie. Echter, bij de lagere

temperaturen werd een stijging van de NOB activiteit waargenomen en hierdoor dus ook een

grotere competitie tussen AnAOB en NOB voor nitriet (uitdaging 4). Er werd aangetoond dat

verhoogde NO concentraties selectief de AnAOB konden stimuleren over de NOB activiteit

(Hoofdstuk 7). Piekbelastingen samen met nitrietaccumulatie, welke de NO productie deden

stijgen, zorgden dan ook voor een verhoogde verwijderingsefficiëntie. Om de toepassing van

hoofdstroom OLAND in een bredere context te beoordelen, werd een levenscyclusanalyse

(LCA) uitgevoerd op volle-schaal data van de RWZI in Strass (Oostenrijk). In deze RWZI

wordt een OLAND-type reactor, in dit specifieke geval DEMON genoemd, toegepast. Drie

scenarios werden onderzocht: (1) de RWZI zonder DEMON; (2) de RWZI met DEMON in de

zijstroom; (3) de RWZI met DEMON in zowel zij- als hoofdstroom (Hoofdstuk 8). Voor dit

laatste scenario werden volle-schaal data gebruikt van een allereerste poging tot hoofdstroom

DEMON in de RWZI in Strass. De LCA toonde aan dat DEMON in de zijstroom van de

RWZI een positief effect had op de eutrophication potential, abiotic depletion potential en

global warming potential en hierdoor kon zorgen voor een meer duurzame waterzuivering.

Samenvatting

188

Uit de eerste volle-schaal data van hoofstroom DEMON operatie in Strass kon worden

geconcludeerd dat om een gelijkaardige graad van duurzaamheid te bekomen, de N2O

emissies (nu ongeveer 2% van de N belasting) in de hoofdstoom verminderd dienen te worden

tot 0.5% van de stikstofbelasting. Dit aangezien de N2O emissies de dominante factor was in

de global warming potential en dus ook de CO2 footprint van de plant. Aangezien N2O

emissies vooral gestuurd worden door de operationele condities en niet door het specifieke

proces zelf, zou het mogelijk moeten zijn om de emissies verder te optimaliseren en te

verminderen zodat eenzelfde CO2 footprint bekomen wordt als bij zijstroom toepassing.

In het algemeen toonde dit werk aan dat nieuwe potentiële toepassingsdomeinen voor

OLAND te vinden zijn in landbouw, waar ammoniakbehandeling nodig is en in

huishoudelijke waterzuiveringssystemen, door een nieuwe hoofdstroombehandeling. Verdere

testen in deze toepassingsdomeinen zijn echter nodig in de toekomst om de performantie en

algemene duurzaamheid te evalueren op grotere schaal.

Bibliography

189

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Curriculum vitae

207

Curriculum vitae

Personal information

Full name: Haydée De Clippeleir

Date of birth: 6th May 1985

Place of birth: Sint-Niklaas, Belgium

Nationality: Belgian

Adres: Schaubeke 19a, 9220 Hamme, Belgium

Phone: +32 473 68 65 88

Email: [email protected]

[email protected]

Education

2008-now: Ph.D. in Applied Biological Sciences (option environmental technologies)

(LabMET, Ghent University)

Doctoral schools of engineering – Ghent University

Funding: Institute for the Promotion of Innovation through Science and

Technology in Flanders (IWT-Vlaanderen)

Ph.D. thesis: Microbial resource management of OLAND focused on

sustainability

Promotors: Prof. dr. ir. Willy Verstraete and Prof. dr. ir. Nico Boon

2003-2008: Bioscience engineer in Environmental technology (Master)

Faculty of Bioscience engineering – Ghent University

Graduated with great distinction

Master thesis: Technological and microbial aspects of the OLAND process

Promotor: Prof. dr. ir. Willy Verstraete

Training period: Svartsjöprojektet (Hultsfred, Zweden) for DEC nv.

1997-2003: Science – Mathematics (8h)

Sint-Vincentius instituut, Dendermonde

mailto:[email protected]

mailto:[email protected]

Curriculum vitae

208

Professional activities

2008-now: Scientific collaborator at Laboratory for microbial ecology and technology

(LabMET)

Contact: Coupure Links 653, 9000 Gent;

phone +32(0)92645976

Coordinator of PC exercises for the courses: ‘Environmental biotechnology’, ‘Microbial ecological

processes’ and ‘Re-use Technologies’

Tutor of 7 master students

Collaborations and contributions to:

CO project: aerobic granular sludge technology (Paul Ockier, TNAV)

Project: ‘Analyse des Einflusses der auptstrom -Deammonifikation auf die flüssigen und

gasförmigen Emissionen kommunaler Kläranlagen in Österreich’, in collaboration with

Norbert Wiessenbacher (BOKU, Vienna, Austria) and Bernhard Wett (ARAconsult,

Innsbruck, Austria)

Scientific contributions

A1 publications

De Clippeleir H., Schaubroeck S., Weisssenbacher N., Dewulf J., Boeckx P., Boon N. and Wett B.

Environmental assessment of one-stage nitritation/anammox implementation in sewage treatment

plants. Submitted.

De Clippeleir H., Vlaeminck S.E., De Wilde F., Daeninck K., Mosquera M., Boeckx P., Verstraete

W. and Boon N.. Cold one-stage partial nitritation/anammox on pretreated sewage: feasibility

demonstration at lab-scale. Submitted.

De Clippeleir H., Courtens E., Vlaeminck S.E., Boon N. and Verstraete W. Efficient total nitrogen

removal in an ammonia gas biofilter through high-rate OLAND. Environmental Science and

Technology, 46(16), 8826-8833.

Curriculum vitae

209

Vlaeminck, S.E., De Clippeleir, H. and Verstraete, W. Microbial resource management of one-stage

partial nitritation/anammox. Microbial Biotechnology, 5(3), 433-488.

Schaubroeck, T., Bagchi, S., De Clippeleir, H., Carballa, M., Boon, N., Verstraete, W. & Vlaeminck

S.E. Successful hydraulic strategies to start up OLAND sequencing batch reactors at lab scale.

Microbial Biotechnology, 5(3), 403-414.

De Clippeleir, H., Yan, X., Verstraete, W., Vlaeminck, S.E., 2011. OLAND is feasible to treat

sewage-like nitrogen concentrations at low hydraulic residence times. Applied Microbiology and

Biotechnology, 90, 1537-1545.

De Clippeleir, H., Defoirdt, T., Vanhaecke, L., Vlaeminck, S.E., Carballa, M., Verstraete, W., Boon,

N., 2011. Long-chain acylhomoserine lactones increase the anoxic ammonium oxidation rate in an

OLAND biofilm. Applied Microbiology and Biotechnology, 90, 1511-1519.

Desloover, J., De Clippeleir, H., Boeckx, P., Du Laing, G., Colsen, J., Verstraete, W., Vlaeminck,

S.E., 2011. Floc-based sequential partial nitritation and anammox at full scale with contrasting N2O

emissions. Water Research, 45, 2811-2821.

Vlaeminck, S.E., Terada, A., Smets, B.F., De Clippeleir, H., Schaubroeck, T., Bolca, S., Demeestere,

L., Mast, J., Boon, N., Carballa, M., Verstraete, W., 2010. Aggregate size and architecture determine

microbial activity balance for one-stage partial nitritation and anammox. Applied and Environmental

Microbiology, 76, 900-909.

De Clippeleir, H., Vlaeminck, S.E., Carballa, M. & Verstraete, W. (2009). A low volumetric

exchange ratio allows high autotrophic nitrogen removal in a sequencing batch reactor. Bioresource

Technology, 100, 5010-5015.

Vlaeminck, S.E., Cloetens, L.F.F., De Clippeleir, H., Carballa, M. & Verstraete, W. (2009). Granular

biomass capable of partial nitritation and anammox. Water Science and Technology, 59(3), 609-617.

Curriculum vitae

210

Other publications

De Clippeleir H., Vlaeminck S.E., Courtens E., Verstraete W. and Boon N. (2012) Behandeling van

anaerobe digestaten met OLAND maximaliseert de elektrische netto-energiewinst. WT-afvalwater, 2,

137-153.

Weissenbacher N., De Clippeleir H., Hell M. and Wett B. Lachgasemissionen bei der behandlung von

prozesswässern im deammonificationsverfahren. Österreichische Wasser-und Abfallwirtschaft, 64(1-

2), 247-252.

De Clippeleir H., Vlaeminck S.E., Courtens E., Verstraete W. and Boon N. (2012) Oxygen-limited

autotrophic nitrification/denitrification maximizes net energy gain in technology schemes with

anaerobic digestion, In Renewable Energy Sources, Academy Publish: Wyoming, U.S.A., accepted.

Contributions to conferences, symposia, workshops and seminars

De Clippeleir H., Courtens E., Vlaeminck S.E., Boon N. and Verstraete W. Efficient total nitrogen

removal in an ammonia gas biofilter through high rate OLAND. Ecotechnolgies for wastewater

treatment, Santiago de Compostela, Spain, 25-27 June 2012. Oral presenation

De Clippeleir H., Weissenbacher N., Boeckx P., Chandran K., Boon N. and Wett B. 2012. Interplay

of intermediates in the formation of NO and N2O during full-scale partial nitritation/anammox.

Ecotechnologies for wastewater treatment, Santiago de Compostela, Spain, 25-27 June 2012. Oral

Presentation

De Clippeleir H., Vlaeminck S.E., Courtens E., Verstraete W. and Boon N. Oxygen-limited

autotrophic nitrification/denitrification maximizes net energy gain in technology schemes with

anaerobic digestion Leading Edge Technology 2012, Brisbane, Australia, 4-7th June 2012. Poster

presentation

De Clippeleir H., Courtens E., Vlaeminck S.E., Boon N. and Verstraete W. A high-rate ammonia gas

biofilter based on partial nitritation/anammox removes total nitrogen at high efficiency. 17th PhD

symposium, 10 February 2012. Oral presenation

Curriculum vitae

211

Weissenbacher, N., De Clippeleir, H., Boeckx, P., Hell, M., Chandran, K., Murthy, S. and Wett, B.

Control of N2O-emissions from Sidestream Treatment. WEFTEC, Los Angeles, 15-19 October 2011.

Oral presentation (co-author)

De Clippeleir, H. Vlaeminck, S.E., Van Acker, J., Boon, N. and Verstraete, W. An oxygen-limited

batch test as experimental model for OLAND application screening and scenario analysis. IWA young

water professionals, 20-22 September 2011. Oral presentation

De Clippeleir, H., Yan, Y., Verstraete, W. and Vlaeminck, S.E. OLAND is feasible to treat sewage-

like nitrogen concentrations at low hydraulic residence time. MRM symposium, Gent, 30the June

2011. Poster presentation

De Clippeleir, H., Yan, X., Verstraete, W. and Vlaeminck, S.E. Approaching energy-positive sewage

treatment: OLAND removes nitrogen from low-strength wastewater. Nutrient Recovery and

Management Conference, Inside and Outside the Fence, Miami, 9-12 January 2011. Oral presentation

Desloover, J., De Clippeleir, H., Boeckx, P., Du Laing, G., Colsen, J., Verstraete, W., Vlaeminck,

S.E., 2011. Floc-based sequential partial nitritation and anammox at full scale with contrasting N2O

emissions. Nutrient Recovery and Management Conference, Inside and Outside the Fence, Miami, 9-

12 January 2011. Oral presentation (co-author)

De Clippeleir, H., Yan, Y., Verstraete, W. and Vlaeminck, S.E. OLAND is feasible to treat sewage-

like nitrogen concentrations at low hydraulic residence time, 16th PhD symposium Applied biological

sciences, 20 December 2010. Oral presentation

De Clippeleir, H., Defoirdt, T., Vanhaecke, L., Vlaeminck, S.E., Carballa, M., Verstraete, W., Boon,

N. Long-chain acylhomoserine lactones increase the anoxic ammonium oxidation rate in an OLAND

biofilm. 16th PhD symposium Applied biological sciences, 20 December 2010. Poster presentation

De Clippeleir, H., Defoirdt, T., Vanhaecke, L., Vlaeminck, S.E., Carballa, M., Verstraete, W. and

Boon, N. Long-chain acylhomoserine lactones increase the anoxic ammonium oxidation rate in an

OLAND biofilm. Conference ISME13: Microbe – stewards of a changing planet, Seatlle, 22the

August 2010, poster presentation

Curriculum vitae

212

De Clippeleir H., Defoirdt T., Vanhaecke L., Vlaeminck S.E., Carballa M., Verstraete W. and Nico

Boon (2010). Long-chain acylhomoserine lactones increase the anoxic ammonium oxidation rate in an

OLAND biofilm. Workshop on bacterial and fungal biofilms, Leuven, Belgium, 25 May 2010. Oral

presenatation

De Clippeleir, H., Vlaeminck, S.E., Carballa, M. and Verstraete W. High and stable autotrophic

nitrogen removal in a sequencing batch reactor by applying a low volumetric exchange ratio. IWA 2nd

Specialized Conference on Nutrient Management in Wastewater Treatment Processes, Krakow, 6-9

September 2009. Oral presentation

Vlaeminck, S.E., Carballa, M., De Clippeleir, H. and Verstraete, W. Biofilm and granule applications

for one-stage autotrophic nitrogen removal. Seminar Nederlandse Biotechnologische Vereniging and

UNESCO-ISHE on nitrogen removal and recovery from water and wastewater. Delft, 26 March 2009.

Oral presentation (co-author)

Vlaeminck, S.E., Terada, A., Carballa, M., De Clippeleir, H., Boon, N., Smets, B.F. and Verstraete,

W. Fluorescence in situ hybridization (FISH) to elucidate structure and diversity in granular biomass

for the treatment of nitrogenous wastewater. 14the Symposium on Applied Biological Sciences,

Ghent, 15 September 2008. Oral presentation (co-author)

Vlaeminck, S.E., De Clippeleir, H., Carballa, M., Terada, A., Smets, B.F. and Verstraete, W.

Granular biomass capable of partial nitritation and anammox. IWA World Water Congress and

Exhibition. Vienna, 7-12 September 2008. Oral presentation (co-author)

Vlaeminck, S.E., Cloetens, L.F.F., De Clippeleir, H., Carballa, M., Smets, B.F. and Verstraete, W.

Granular biomass capable of combined aerobic and anoxic ammonium oxidation. Fall symposium of

the ‘Nationale Vereniging voor Microbiologie (NVvM) - Microbiële Ecologie’. Amsterdam, 23

November 2007. Oral presentation (co-author)

Awards

Best platform presentation with “Interplay of intermediates in the formation of NO and N2O during

full-scale partial nitritation/anammox.” Ecotechnologies for wastewater treatment, Santiago de

Compostela, Spain, June 27th 2012.

Curriculum vitae

213

Best short presentation with “Efficient total nitrogen removal in an ammonia gas biofilter through

high rate OLAND.” Ecotechnolgies for wastewater treatment, Santiago de Compostela, Spain, June

27th 2012.

Curriculum vitae

214

Dankwoord

215

Dankwoord

“Doctoren is zoals een olympische discipline” werd me bij de start van mijn onderzoek

medegedeeld. Nu ik terugblik, kan ik toch menig gelijkenis erkennen. Zoals atleten die

toewerken naar de olympische spelen, werd ook tijdens dit onderzoek met vallen en opstaan

gezocht naar de beste techniek en tactiek om uiteindelijk dit ene einddoel, het afleggen van

een doctoraat, te behalen. Er werd gezweet (letterlijk en figuurlijk), gejuicht maar er werden

ook tegenslagen verwerkt. Echter, niets kon tot stand komen zonder een fantastisch team en

een enthousiaste achterban.

First of all I would like to thank the jury members, it is really an honor to defend my work in

front of such prominent group of scientists. Your thorough examination of this work and the

doubtlessly critical questions are greatly appreciated.

Mijn coaches, promotoren, Willy Verstraete en Nico Boon ben ik dankbaar omdat ze me de

kans gaven bij LabMET te werken. Prof. Verstraete, jouw energie en enthousiasme werkten

steeds aanstekelijk. Ik wil je bedanken om me steeds verder te pushen, maar ook voor de

vrijheid dit u me gegeven hebt om achter ideeën aan te gaan. Nico, ik wil je bedanken voor

het vertrouwen dat je altijd in me had. De korte, maar efficiënte discussies die we hadden

waren steeds leerrijk en brachten me steeds weer op het goede pad.

Het OLAND-team dat later werd uitgebreid naar het N-team heeft steeds een belangrijke rol

gespeeld in mijn onderzoek en hoewel dit team jaarlijks wisselde was er steeds één vaste

speler. Siegfried, a.k.a. doctor OLAND, jij was de dirigent van dit team. Ik had de

ongelooflijke luxe om met jou als OLAND expert samen te werken. Ik heb genoten van onze

brainstormsessies, discussies, uitstappen, bbq’s enz. Ik kon me geen betere begeleider

wensen. Dank je. One of the people who inspired and motivated me to do research was Marta,

which I admire for her nononsence approach and infectious motivation and energy. Marta,

thanks for the good talks, advices and great time in Santiago de Compostela. Ons team werd

elk jaar versterkt door thesisstudenten: Yan, Tijs, Katrien, Emilie, Jeroen, Fabian en Mariela.

Jullie hebben door jullie briljante werk een grote hand gehad in dit werk. Emilie, met jou is de

OLAND opvolging verzekerd. Jouw enthousiasme en gedrevenheid zal je nog ver brengen.

Dankwoord

216

I would also like to thank Bernhard, Norbert and Martin for the great cooperation during the

measurement campaigns in Strass. Bernhard, I could never believe that you agreed in a

cooperation on the mainstream treatment subject when I asked you in Miami (with some

pushing of Sudhir, I have to admit). You gave me the chance to get a feeling with the full-

scale application. I would like to thank you for the open discussions we had and for the good

contacts, which opened doors for new challenges. Norbert, I would like to thank you for the

enjoyable time in Strass, we formed a good team. Keep on practicing your lab-skills though!

Martin, if every operator was that dedicated to his work as you are, DEMON or OLAND

reactors were already operational in every wastewater treatment plant. Thank you for your

help, enthusiasm and great atmosphere during work.

Ook alle collega’s van LabMET wil ik bedanken voor de aangename werksfeer. De Rotonde,

ook wel in de volksmond beter bekend als ‘het centrum van de kennis’, werd bevolkt door één

voor één flamboyante figuren die zorgden voor een levendige sfeer en een goede afwisseling

tussen het labowerk door. Joachim, medebewoner van het eiland, bedankt voor het

tegengewicht aan al dat wielergeweld, de lachgasdiscussies en veel succes met het afwerken

van je sprookjesboek. Simon, voorzitter van de frietcluster, hou de traditie hoog en sprokkel

nog wat energie voor je laatste jaar. Willem, weetjesman van de rotonde, je flitsbezoeken aan

de rotonde zijn legendarisch. Tom, voorzitter van de rotonde, veel succes daar aan de

overkant van de grote plas. Als je in de buurt van DC komt, spring gerust even binnen. Aan

het nieuw jonge geweld (Sam, Emilie, Joeri en Stephen): bedankt voor de nieuwe frisse wind.

Ook de oud-rotondenaars (Ilse, Selin, Bart, Peter, Lois) waren één voor één kleppers die het

aangenaam werken maakten.

Verder waren er nog mensen van binnen en buiten LabMET waar ik veel aan te danken heb.

Greet, bedankt voor het helpen met de IC en bestellingen. Tim, je figuren maken dit werk af!

Bedankt voor je tijd en precisie. Siska, bedankt voor je moleculaire ondersteuning. Ook een

dikke merci aan het secretariaat (Kris, Regine). Samuel en Katja, bedankt voor de hulp bij de

NO testen en metingen. Jan bedankt voor de hulp met de N2O metingen. Thomas bedankt

voor je inzet bij het LCA werk.

Uiteraard zijn het niet alleen de werk-gerelateerde mensen maar des te meer de achterban van

vrienden en familie die maken dat je je zinnen kan verzetten als het iets minder. Een heel

groot deel van mijn vrije tijd ging uit naar volleybal. Mijn geweldige ploegmaats groeiden uit

Dankwoord

217

tot vrienden. De super sfeer, cava-momenten, verkleedtrainingen, maar ook het samen afzien

en strijden tijdens de wedstrijden waren een belangrijke bron van ontspanning. Het was dan

ook een eer jullie kapitein te zijn. Sophie, An, Marijke, Lore, Assunta, Tessa, Kimberly,

Anke, Ruth, Joëlle, Sylvie en Philip doe dit nog goed dit seizoen en hou met op de hoogte. Ik

hou me alvast klaar voor het kerstfeestje. Ook onze verlaters Evelien (mede orvalliefhebber),

Kim (altijd klaar voor een frietkotstop) en Steven (onze reddende engel) hebben me een leuke

tijd bezorgd. Naast dit fantastisch team was er ook nog het jeugdig geweld van onze

miniemenploeg. Hun enthousiasme en speelsheid deden me terug beseffen waarom we dit

spelletje zo leuk vinden. Naast het volleybalgeweld brachten de leuke babbels en etentjes en

drinks een goed tegengewicht. Annabel en endrik, hoewel onze agenda’s niet altijd even

gemakkelijk bij elkaar te leggen vielen, waren de momenten dat we samen waren altijd zeer

tof en ontspannend. Dennis en Jessica, met jullie heb ik mijn eerste ski-ervaring opgedaan,

mountainbike tochten georganiseerd, maar vooral leuke momenten beleefd. Zeker nu Yente

ertussen loopt, brengt dit steeds leven in huis.

Als laatste en belangrijkste steun wil ik nog mijn ouders en broers bedanken. Mama en papa,

bedankt voor jullie onvoorwaardelijke steun en vertrouwen. Door jullie ben ik kunnen

uitgroeien tot wat ik nu ben. Johan en Maarten, bedankt voor de leuke ontspannende

momenten, de toffe reizen en zoveel meer. Ik verwacht jullie dan ook in de zomer voor een

Amerikaans avontuur!

Haydée, Oktober 2012

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