microbial resource management of oland focused on sustainability
DESCRIPTION
Dr. ir. Haydée De Clippeleir, 2012, PhD thesisTRANSCRIPT
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
Engineering, Ghent University, Gent, Belgium
Members of the examination committee:
Prof. dr. ir. Pascal Boeckx
Department of Applied analytical and physical chemistry, Faculty of Bioscience
Engineering, Ghent University, Gent, Belgium
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
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 Materials and methods ................................................................................................................. 43
2.1 Reactor operation ..................................................................................................................................... 43
2.2 Emission measurments .......................................................................................................................... 43
3 Results and discussion .................................................................................................................. 45
4 Conclusions ....................................................................................................................................... 51
5 Acknowledgements ........................................................................................................................ 52
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
4 Acknowledgements ........................................................................................................................ 77
Chapter 5: Efficient total nitrogen removal in an ammonia gas biofilter through high-
rate OLAND ............................................................................................................................ 79
1 Introduction ..................................................................................................................................... 80
2 Materials and methods ................................................................................................................. 83
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
6 Acknowledgements ........................................................................................................................ 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
6 Acknowledgements ...................................................................................................................... 128
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
5 Acknowledgements ...................................................................................................................... 154
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
China1 Airlift Glutamate
factory
<500 - 2.0 >80 4500 - - - - Granules
China1 Airlift Glutamate
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
A low volumetric exchange ratio allows high autotrophic N removal in a SBR
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.
A low volumetric exchange ratio allows high autotrophic N removal in a SBR
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 *
A low volumetric exchange ratio allows high autotrophic N removal in a SBR
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
A low volumetric exchange ratio allows high autotrophic N removal in a SBR
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
A low volumetric exchange ratio allows high autotrophic N removal in a SBR
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
Denitrification (HDN)
Nitritation (AerAOB)
NH4+
NO2− NO3
−
O2
N2
NON2O
Nitratation (NOB)
O2
Anammox (AnAOB)
NH2OH
Denitrification (HDN)
Nitritation (AerAOB)
Figure 3.1: Nitrogen conversion in relation to NO and N2O formation.
2 Materials and methods
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
Interplay of intermediates in formation of N2O/NO during OLAND
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
Interplay of intermediates in formation of N2O/NO during OLAND
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).
Interplay of intermediates in formation of N2O/NO during OLAND
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
Interplay of intermediates in formation of N2O/NO during OLAND
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
Interplay of intermediates in formation of N2O/NO during OLAND
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;
OLAND maximizes net energy gain in systems with anaerobic digestion
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
OLAND maximizes net energy gain in systems with anaerobic digestion
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
1.2.1 State of the art
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.
1.2.2 Implication of OLAND application on the energy balance
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.
OLAND maximizes net energy gain in systems 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
1.3.1 State of the art
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.
1.3.2 Implication of OLAND application on the energy balance
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
OLAND maximizes net energy gain in systems with anaerobic digestion
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.
OLAND maximizes net energy gain in systems with anaerobic digestion
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
OLAND maximizes net energy gain in systems with anaerobic digestion
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.
1.4.3 Implication of OLAND application on the energy balance
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
OLAND maximizes net energy gain in systems with anaerobic digestion
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
OLAND maximizes net energy gain in systems with anaerobic digestion
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
OLAND maximizes net energy gain in systems with anaerobic digestion
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).
OLAND maximizes net energy gain in systems with anaerobic digestion
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).
OLAND maximizes net energy gain in systems with anaerobic digestion
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.
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
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 Materials and methods
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
-2 biofilter section d
-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).
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
84
2.4 Chemical analyses
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
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
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
-2 biofilter section d
-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
-2 biofilter section d
-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
-2 biofilter section d
-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).
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
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
-2 biofilter section d
-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.
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
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
-2 biofilter section d
-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).
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
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
-2 biofilter section d
-1) and the nitrogen loading rate (kg N m
-2 biofilter section d
-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
Efficient total nitrogen removal in an ammonia gas biofilter through high-rate OLAND
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.
OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT
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
OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT
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.
2.3 Chemical 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.
OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT
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).
OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT
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
OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT
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.
OLAND is feasible to treat sewage-like nitrogen concentrations at low HRT
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 Materials and methods
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
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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).
2.5 Chemical 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). 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
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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.
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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
Relative NO2- prod (% of NH4
+ 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
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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%.
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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
Relative NO3- prod (% of NH4
+ cons*) 42 ± 5 43 ± 12 63 ± 2 50 ± 6 62 ± 18 46 ± 6
Relative NO2- prod (% of NH4
+ cons) 20 ± 4 10 ± 10 5 ± 1 8 ± 3 7 ± 4 13 ± 6
AerAOB activity (mg NH4+-N L
-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
AnAOB activity (mg Ntot L-1
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
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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*
AerAOB activity (mg NH4+-N L
-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*
AnAOB activity (mg Ntot L-1
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
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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.
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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.
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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.
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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.
Cold OLAND on preteated sewage: feasibility demonstration on lab-scale
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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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 Materials and methods
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).
Environmental assessment of the implementation of OLAND in sewage treatment plants
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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
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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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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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.
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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.
Environmental assessment of the implementation of OLAND in sewage treatment plants
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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.
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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
Ntot removal (%) 89 84 79
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
Ntot removal (%) 83 88 91
°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.
Environmental assessment of the implementation of OLAND in sewage treatment plants
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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.
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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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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.
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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.
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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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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
Environmental assessment of the implementation of OLAND in sewage treatment plants
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
General discussion and perspectives
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
General discussion and perspectives
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.
General discussion and perspectives
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
General discussion and perspectives
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
General discussion and perspectives
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).
General discussion and perspectives
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
General discussion and perspectives
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
General discussion and perspectives
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
General discussion and perspectives
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.
General discussion and perspectives
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.
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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]
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
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.
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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
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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.
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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