Faculty of Bioscience Engineering
Centre for Environmental Sanitation
Academic Year: 2010 - 2011
Effect of heterotrophic growth on autotrophic nitrogen removal in a
granular sludge reactor
Md. Salatul Islam Mozumder
Promotor:
Prof. dr. ir. Eveline I. P. Volcke
Master’s dissertation submitted in partial fulfillment of the requirements for the
degree of Master of Environmental Sanitation
Acknowledgment
This thesis is the result of one year of work during which I have been accompanied and
supported by many people. It is a pleasant aspect that I hereby have the opportunity to
express my gratitude for all of them.
First of all I would like to thank my promoter, prof. dr. ir. Eveline Volcke. I have been in her
laboratory in the department of Biosystems Engineering since August 2010. During this
period I have known prof. Eveline as a motivated, patient and principle centered person. Her
extreme enthusiasm and integral view on research and her mission for providing only high
quality work have made a deep impression on me. I owe her a great debt of gratitude for
showing me the way of research. I will never forget that she helped me to build confidence
and collect experience. These will affect me in my future life.
My sincere gratitude also goes to prof. dr. ir. Mark van Loosdrecht and dr. ir. Cristian
Picioreanu, Department of Biotechnology, Delft University of Technology, The Netherlands
for their valuable advice and constructive comments that inspired me to work more
effectively. I also send my cordial thanks to Matthijs Daelman for his support during my
research stay at Delft University of Technology.
I am also grateful to the fellow students I met while studying in Belgium during the past two
years. They are too numerous to mention but I want to thank all of them for being such good
colleagues and friends.
Finally, I am forever indebted to my parents and wife who supported me in a lot of things that
really matter in life. Without their wholehearted help and understanding, I could not have
accomplished my master study.
Md. Salatul Islam Mozumder
Ghent, 2011
Notation index
List of abbreviations
AOB
Ammonium oxidizing bacteria
NOB
Nitrite oxidizing bacteria
Anammox
Anaerobic ammonium oxidation
COD
Chemical oxygen demand
List of symbols
Soluble compounds (S)
Ss
Concentration of organic substrate
g.m-3
SNH
Ammonium concentration
g.m-3
SNO2
Nitrite concentration
g.m-3
SNO3
Nitrate concentration
g.m-3
SN2
Nitrogen concentration
g.m-3
SN2A
Nitrogen concentration produced by autotrophs
g.m-3
SN2H
Nitrogen concentration produced by heterotrophs
g.m-3
Particulate compounds (X)
XAOB
Ammonium oxidizing bacteria
gCOD.m-3
XNOB
Nitrite oxidizing bacteria
gCOD.m-3
XAN
Anammox bacteria
gCOD.m-3
XH
Heterotrophic bacteria
gCOD.m-3
XH,A
Aerobic heterotrophs
gCOD.m-3
XH,NO2
Anoxic heterotrophs on nitrite
gCOD.m-3
XH,NO3
Anoxic heterotrophs on nitrate
gCOD.m-3
XI
Inert biomass
gCOD.m-3
Process
ρG,AOB
Growth rate of XAOB
gCOD.m-3.d-1
ρG,NOB
Growth rate of XNOB
gCOD.m-3.d-1
ρG,AN
Growth rate of XAN
gCOD.m-3.d-1
i
Notation index
ρG,H
Growth rate of XH,A
gCOD.m-3.d-1
ρAG,HNO2
Growth rate of XH,NO2
gCOD.m-3.d-1
ρAG,HNO3
Growth rate of XH,NO3
gCOD.m-3.d-1
ρD,AOB
Decay rate of XAOB
gCOD.m-3.d-1
ρD,NOB
Decay rate of XNOB
gCOD.m-3.d-1
ρD,AN
Decay rate of XAN
gCOD.m-3.d-1
ρD,H
Decay rate of XH
gCOD.m-3.d-1
Stoichiometric parameters
YAOB
Yield of ammonium oxidizers on ammonia
gCOD.g-1N
YNOB
Yield of nitrite oxidizers on nitrite
gCOD.g-1N
YAN
Yield of Anammox bacteria on ammomium
gCOD.g-1N
YH
Yield of aerobic heterotrophic bacteria
gCOD.g-1N
YH,NO2
Yield of anoxic heterotrophic bacteria on nitrite
gCOD.g-1N
YH,NO3
Yield of anoxic heterotrophic bacteria on nitrate
gCOD.g-1N
iNXB
Nitrogen content in active biomass
gN.g-1COD
iNXI
Nitrogen content in XI
gN.g-1COD
iNSS
Nitrogen content in organic substrate
gN.g-1COD
fI
Inert content in biomass
gCOD.g-1COD
Kinetic parameters
AOB
 max
Maximum growth rate of XAOB
d-1
NOB
 max
Maximum growth rate of XNOB
d-1
AN
 max
Maximum growth rate of XAN
d-1
H
 max
Maximum growth rate of XH
d-1
K AOB
NH
Affinity constant of XAOB for ammonium
gN.m-3
NOB
K NO
2
Affinity constant of XNOB for nitrite
gN.m-3
K AN
NH
Affinity constant of XAN for ammonium
gN.m-3
AN
K NO
2
Affinity constant of XAN for nitrite
gN.m-3
KH
NO 2
Affinity constant of XH for nitrite
gN.m-3
KH
NO 3
Affinity constant of XH for nitrate
gN.m-3
KH
S
Affinity constant of XH for organic substrate
gCOD.m-3
ii
Notation index
K AOB
O2
Affinity constant of XAOB for oxygen
gO2.m-3
K NOB
O2
Affinity constant of XNOB for oxygen
gO2.m-3
K AN
O2
Affinity constant of XAN for oxygen
gO2.m-3
KH
O2
Affinity constant of XH for oxygen
gO2.m-3
bAOB
Decay constant of XAOB
d-1
bNOB
Decay constant of XNOB
d-1
bAN
Decay constant of XAN
d-1
bH
Decay constant of XH
d-1
ηNO2
Anoxic reduction factor for XHNO2
-
ηNO3
Anoxic reduction factor for XHNO3
-
Physical parameters
DNH4
Ammonium diffusion coefficient in water
m2.d-1
DNO2
Nitrite diffusion coefficient in water
m2.d-1
DNO3
Nitrate diffusion coefficient in water
m2.d-1
DO2
Oxygen diffusion coefficient in water
m2.d-1
DN2
Nitrogen diffusion coefficient in water
m2.d-1
DS
Organic substrate diffusion coefficient in water
m2.d-1
ρA
Density of autotrophic and particulate inert biomass
gCOD.m-3
ρH
Density of heterotrophic biomass
gCOD.m-3
rp
Granule radius
mm
T
Temperature
K
Tref
Reference temperature
K
R
Gas constant
J.mole-1.K-1
EaAOB
Activation energy of XAOB
kJ.mole-1
EaNOB
Activation energy of XNOB
kJ.mole-1
EaAN
Activation energy of XAN
kJ.mole-1
iii
Summary
This study deals with the influence of heterotrophic growth on autotrophic nitrogen removal
in a granular sludge reactor. Autotrophic nitrogen removal is an innovative technique for
biological nitrogen removal from wastewater during which ammonium is nitrified to nitrite
by ammonium oxidizing bacteria followed by subsequent oxidation of ammonium and
reduction of nitrite to nitrogen gas by anammox bacteria. In this process, nitrification of
nitrite to nitrate needs to be prevented by outcompeting nitrite oxidizing bacteria, which can
be achieved at relatively low oxygen level. The abovementioned biomass groups concern
autotrophic species1. Heterotrophic organism, are a priori not expected when treating influent
wastewater stream which contain only nitrogen and no organic carbon. However,
heterotrophic bacteria can grow on microbial decay products and their presence will affect
the process performance. The relation between autotrophic and heterotrophic species is the
subject of this dissertation.
In this research work, a mathematical model was constructed to describe the effect of
heterotrophic growth on completely autotrophic nitrogen removal. The developed model
considered both autotrophic and heterotrophic growth, besides decay of all. Subsequently,
simulation studies were performed in which autotrophic nitrogen removal with and without
heterotrophic growth would be compared. With respect to the modeling assumption, the
sensitivity of the density of heterotrophs was evaluated and found to be insensitive to the
simulation results.
The biomass profile in a granule revealed that heterotrophs were present at the outer layer
just below the ammonium oxidizing and nitrite oxidizing bacteria that consume oxygen.
Anammox bacteria grew in the inner anoxic parts of the granules where they consume
ammonia and nitrite and produce nitrogen gas. The nitrogen removal was significantly higher
when heterotrophic growths were considered in the model. The optimum bulk oxygen
concentration levels corresponding with maximum nitrogen removal were related to the
process variables such as granular size, possible presence of organic substrate in influent,
ammonium surface load and temperature. Higher granular size, organic substrate load and
ammonium surface load needed higher bulk oxygen concentration for maximum nitrogen
1
Use CO2 as a carbon source.
iv
Summary
removal. Similar maximum nitrogen removal efficiency was found in a range of
temperatures.
This study considers a wastewater influent stream containing only nitrogen, as well as a
wastewater stream that contains organic substrate. Heterotrophic growth that increased
nitrogen removal, was facilitated by the presence of organic substrate in the influent.
This study clearly demonstrates the influence of heterotrophs on the performance of
autotrophic nitrogen removal in a granular sludge reactor even if little or no organics is
present in the wastewater stream. The insight gained on the interaction between heterotrophic
and autotrophic bacteria becomes ever more important at lower temperature and will thus
gain importance for the operation of granular sludge reactor in future energy-positive
wastewater treatment plants.
v
Table of Contents
Chapter I: Introduction ............................................................................................................... 1
Chapter II: Literature review ..................................................................................................... 3
1. Introduction ..................................................................................................................... 3
2. Nitrogen removal pathways ............................................................................................ 4
2.1.
Conventional nitrification-denitrification over nitrate ............................................ 4
2.2.
Nitrification-denitrification over nitrite ................................................................... 5
2.3.
Anaerobic ammonium oxidation (Anammox) ......................................................... 5
2.4.
Partial nitritation combined with anaerobic ammonium oxidation ......................... 6
3. Reactor conditions affecting biological nitrogen removal .............................................. 7
3.1.
Oxygen concentration .............................................................................................. 7
3.2.
Temperature ............................................................................................................. 8
3.3.
pH ............................................................................................................................ 9
4. Relation between influent organic carbon (COD/N ratio) and biological nitrogen
removal ................................................................................................................................ 10
4.1.
Nitrification-denitrification over nitrate ................................................................ 12
4.2.
Anammox process ................................................................................................. 14
4.3.
Partial nitritation-anammox ................................................................................... 14
5. Conclusions ................................................................................................................... 15
Chapter III: Model development .............................................................................................. 16
1. Process stoichiometry and kinetics ............................................................................... 16
2. Reactor configuration, simulation parameters and initial conditions ........................... 23
Chapter IV: Results and discussion ......................................................................................... 25
1. Role of heterotrophic growth on nitrogen removal....................................................... 25
2.
1.1.
Active biomass composition .................................................................................. 25
1.2.
Competition among active biomass ....................................................................... 27
1.3.
Comparison of nitrogen removal performance ...................................................... 28
Biomass dynamics and steady state .............................................................................. 30
2.1.
Biomass dynamics in a granule ............................................................................. 30
2.2.
Influence of initial conditions on the time needed to reach steady state ............... 32
3. Influence of operational parameters on the reactor performance ................................. 34
3.1.
Influence of the oxygen concentration .................................................................. 34
3.1.1.
Dynamics of nitrogen removal and steady state biomass profile ................... 34
vi
Table of contents
3.1.2.
Steady state performance and biomass composition ...................................... 36
3.1.3.
Sensitivity analysis for the density of heterotrophs ........................................ 38
3.2.
Influence of the granule size .................................................................................. 39
3.2.1.
Dynamics of nitrogen compounds and steady state biomass profile .............. 39
3.2.2.
Steady state reactor performance and biomass composition .......................... 41
3.2.3.
Interaction between granule size and oxygen concentration .......................... 42
3.3.
Role of temperature ............................................................................................... 43
3.3.1.
Effect of temperature at fixed oxygen level ................................................... 43
3.3.2.
Interaction of bulk oxygen with temperature ................................................. 44
3.4.
Effect of ammonium surface load ......................................................................... 45
4. Influence of influent organic substrate on reactor performance ................................... 47
4.1.
Effect of organic substrate at fixed oxygen level .................................................. 47
4.2.
Effect of oxygen concentration at fixed influent organic substrate ....................... 48
4.3.
Interaction between organic substrate and oxygen concentration ......................... 49
4.4.
Effect of organic substrate on dynamics of nitrogen removal ............................... 50
Chapter V: Conclusions and perspectives................................................................................ 52
1. Steady state and dynamic model behaviour .................................................................. 52
2. Influence of operational parameters and influent organic substrate ............................. 53
3. Future works ................................................................................................................. 54
References ................................................................................................................................ 55
vii
Chapter I: Introduction
Nowadays, nitrogen removal is very important in terms of water pollution control. The
nitrogen pollutants in wastewater are either ammonium (NH4+) or organic nitrogen
compounds which are ultimately converted to ammonium through hydrolysis. Traditional
biological nitrogen removal from wastewater is performed through nitrificationdenitrification over nitrate. This pathway requires a significant amount of aeration energy for
biological nitrification and an external carbon source for denitrification. Partial nitritationanammox process is a promising alternative biological nitrogen removal pathway. During
partial nitritation, about 50% of the ammonium present in the wastewater is converted to
nitrite. In the subsequent anammox reaction, ammonium and nitrite are combined to form
nitrogen gas. The resulting process requires up to 63% less oxygen (low aeration cost) and
causes less carbon-dioxide emissions and a lower sludge production compared to
conventional nitrification-denitrification over nitrate. A key issue in the partial nitritation anammox process is partial nitrite production and prevention of further oxidation of nitrite to
nitrate. One process option to achieve partially nitrite formation is by limiting oxygen in a
biofilm reactor such as the granular sludge reactor.
The effectiveness of nitrogen removal process does not only depend on the applied treatment
technology but also the process conditions. The principles of biological nitrogen removal
processes and the parameters affecting their operation are reviewed in Chapter II.
In most previous studies on partial nitritation on anammox processes, only autotrophic
organism were considered to be present (Koch et al., 2000; Matsumoto et al., 2010; Okabe et
al., 2005; van de Graaf et al., 1996). However, some authors showed that in autotrophic
biofilms, heterotrophic biomass can grow on microbial decay products (Kindaichi et al.,
2004; Lackner et al., 2008; Okabe et al., 2005). However, until now this has not been studied
for autotrophic nitrogen removal in granular sludge reactors. In Chapter IV (section 1), the
performance of autotrophic nitrogen removal in a granular sludge reactor together with
biomass profile in a granule with and without heterotrophic growth are compared. The
nitrogen compounds in bulk are also examined while identifying the effects of heterotrophs.
A biofilm granule is a complex microbial system containing different types of bacteria. In a
granular sludge reactor, ammonium oxidizers are active in the outer layer of the granules and
1
Chapter I: Introduction
produce sufficient amount of nitrite for anammox bacteria active in the inner part of the
granules. The dynamics of the microbial profile in a granule and the time required to reach
steady state are evaluated in Chapter IV (section 2).
One of the most important process variables for establishing nitrite formation in a biofilm
reactor is dissolved oxygen. The dissolved oxygen concentration has a large influence on
both ammonium oxidizers and nitrite oxidizers. On the other hand, the anammox bacteria are
strictly anaerobic and inhibited by the dissolved oxygen. The partial nitritation - anammox
process can be well controlled by regulating dissolved oxygen concentration. The limited
oxygen concentration allows for the partial oxidation of ammonium to nitrite while
preventing nitrite oxidation. Under anoxic conditions the unconverted ammonium and nitrite
are utilized by anammox bacteria to remove the nitrogen from aqueous system as nitrogen
gas. Chapter IV (section 3.1) analyzes the effect of bulk oxygen concentration on the
nitrogen removal performance through partial nitritation-anammox process in a granular
sludge reactor.
The granule size in a granular sludge reactor determines the surface to volume ratio and
affects the nitrogen removal performance. Chapter IV (section 3, sub-section 3.2) addresses
changing granular size and evaluates its effects on the performance of granular sludge reactor
in terms of nitrogen removal.
The process temperature plays a very important role in the partial nitritation-anammox
process. It is not only responsible for the microbial growth rate but also for all kinds of
interactions within the system, such as decay rate, microbial activities, equilibrium relation
etc. A higher temperature increases the growth rate, decay rate and microbial activities but at
too high temperature the microbial community is destroyed. Chapter IV (section 3.3)
discusses the temperature effects on partial nitritation - anammox process.
The presence of organic substrates in the influent significantly influences the microbial
community composition in an autotrophic granule. When organic carbon is present
heterotrophic growth increases. Nitrate can be reduced by these denitrifiers to nitrite that can
be utilized by anammox for the oxidation of ammonium (Kumar and Lin, 2010). To
investigate the effect of heterotrophic bacteria, growth on influent organic substrate is
addressed in Chapter IV (section 4).
2
Chapter II: Literature review
1. Introduction
Generally, nitrogen is present in wastewater in the form of ammonium (NH4+). Several
human activities such as agriculture, industrial processes, and household activities produce
nitrogen containing wastewater. High strength nitrogen containing wastewater originates
from manure (Luo et al., 2002; Qiao et al., 2010), landfill leachate (Cema et al., 2007),
several organic chemicals; plastics and synthetic fibers industries (Love et al., 1999) and
sludge digester supernatants (Fux et al., 2002; van Loosdrecht and Salem, 2006).
Uncontrolled disposal of wastewater containing high ammonium concentration causes a huge
damage to the environment. It is prime factor for the eutrofication of the receiving aquatic
system. Besides, dissolved ammonium is considered as a harmful agent for the aquatic life
(Effler et al., 1990). For this reason, nitrogen removal from wastewater has become an
important issue. Due to increasingly stringent environmental regulations, advanced and cost
effective techniques for the nitrogen removal from wastewater are required.
In conventional wastewater treatment plants (WWTPs), ammonium is removed by biological
nitrification-denitrification over nitrate. New approaches that are based on (partial)nitritation
and/or the anaerobic ammonium oxidation (Anammox) process (Mulder et al., 1995) to
remove the nitrogen are more cost-effective, environmentally friendly, efficient and
sustainable. The combined partial nitritation-anammox process is a completely autotrophic
process which can be performed either in one stage or in two stages (reactors). A number of
research groups worked on autotrophic nitrogen removal processes resulting in a variety of
process configurations and various names such as Oxygen-Limited Autotrophic
Nitrification/Denitrification (OLAND) (Kuai et al., 1998), Completely Autotrophic Nitrogen
removal Over Nitrite (CANON) (Third et al., 2005), Sustainable High rate Ammonium
Removal Over Nitrite (SHARON) (Hellinga et al., 1998), Deammonification (Hippen et al.,
1997), pH controlled Deammonification (DEMON) (Wett, 2006) etc.
This literature review gives a short introduction on the principle pathways involved in
biological nitrogen removal followed by an overview of the reactor conditions affecting the
process. This knowledge is important to increase the effectiveness of biological nitrogen
processes.
3
Chapter II: Literature review
To establish autotrophic nitrogen removal, oxygen is the most important process variable. In
first place oxygen is needed to establish partial nitritation but the oxygen level should be low
enough in order not to inhibit anaerobic denitrification. Another critical parameter that has a
large effect on nitrogen removal is the organic (carbon) load which can be expressed as COD
to nitrogen ratio.
The partial nitritation-anammox process is an autotrophic process. Nevertheless, upto half of
the biomass in autotrophic biofilms can be heterotrophic, growing on the microbial decay
products(Kindaichi et al., 2004; Okabe et al., 2005). Heterotrophic growth reduces the nitrate
production. It utilizes the organic substrate in both aerobic and anoxic conditions and
produces carbon dioxide that ultimately decreases the pH of the system. Therefore
heterotrophic activity is an important factor that affects autotrophic nitrogen removal.
2. Nitrogen removal pathways
2.1. Conventional nitrification-denitrification over nitrate
Biologically nitrification and denitrification are two individual processes that are carried out
by distinct groups of bacteria. During nitrification, ammonium is oxidized to nitrate (Eq. 1).
The ammonium oxidizing bacteria (XAOB) convert ammonium to nitrite (NO2-)(Eq. 1a),
which can be further oxidized to nitrate (NO3-) (Eq. 1b) by nitrite oxidizing bacteria (XNOB).
During denitrification, nitrate is transformed to nitrogen gas (Eq. 2). C (in Eq. 2) denotes the
carbon source; for autotrophic denitrification it is carbon dioxide (CO2) and for heterotrophic
it is organic carbon.
Nitrification:
Overall:
Denitrification:
NH4+ + 1.5 O2 → NO2- + H2O + H+
(1a)
NO2- + 0.5 O2 → NO3-
(1b)
NH4+ + 2O2 → NO3- + H2O + 2H+
(1)
NO3- + C + 2H+ → CO2 +0.5N2+ H2O
(2)
Heterotrophic denitrification is a four step reduction processes in which nitrogen gas (N2) is
formed from nitrate (NO3-) over nitrite (NO2-), nitric oxide (NO) and nitrous oxide (N2O).
Each reduction step is catalyzed by different enzymes (Baumann et al. 1996). If for any
4
Chapter II: Literature review
reason, one or more individual reduction steps become slower, the intermediate products may
accumulate in the system, ultimately reducing nitrogen removal (Udert et al., 2008).
Biological nitrification-denitrification over nitrate is considered as an efficient process
characterized by a relatively easy operation and moderate costs (Metcalf and Eddy, 2003). It
is generally used for the treatment of wastewater containing low nitrogen concentration
(<100mgNL-1). This conventional biological nitrification and denitrification process is
considered as more favorable than the chemical nitrogen removal by magnesium-ammoniumphosphate (MAP) precipitation or by air stripping (Siegrist, 1996) for the removal of
ammonium nitrogen from the wastewater.
2.2. Nitrification-denitrification over nitrite
Ammonium in a concentrated stream is oxidized to nitrite only (Eq. 1a) by controlling the
aeration, saving up to 25% aeration cost. The denitrification of nitrite to nitrogen gas is based
on external carbon source (Eq. 3).
NO2- + 0.5C + H+ → N2 + CO2 + H2O
(3)
Nitrification-denitrification over nitrite needs less external carbon source, saving 40% cost
for external carbon. Moreover the process emits less carbon dioxide (CO2) and produces less
sludge compared to conventional nitrification-denitrification over nitrate.
2.3. Anaerobic ammonium oxidation (Anammox)
Until the early 1990s, it was believed that the oxidation of ammonium could only proceed
under aerobic conditions. This thinking was changed by the discovery of the anaerobic
ammonium oxidation process by Mulder et al. (1995). At that time the scientific community
was greatly surprised by the proof of a biological process in which nitrite and ammonium are
directly converted into dinitrogen gas. The overall reaction is:
NH4+ + 1.32NO2- + 0.13H+ → 1.02N2 +0.26NO3- + 2.03 H2O
(4)
Hydrazine and hydroxylamine are produced as intermediates during the anammox process
(Sinninghe-Damste et al., 2002; van de Graaf et al., 1997). Strous et al. (2006) found NO as
an intermediate product of the anammox process.
As the process is anoxic, anammox
bacteria do not need oxygen which results in decreased aeration costs. Furthermore anammox
5
Chapter II: Literature review
bacteria use CO2 as a carbon source and hence they do not require the addition of organic
compounds. They grow relatively slowly, leading to a low sludge production.
2.4. Partial nitritation combined with anaerobic ammonium oxidation
The application of the anammox process for the removal of ammonium from wastewater
requires a proceeding step in which nitrite is produced. About half of the ammonium needs to
be converted to nitrite, a process that is known as partial nitritation (Eq. 1a) and is carried out
by ammonium oxidizing bacteria. The resulting nitrite and unconverted ammonium are
converted to nitrogen gas in the anammox process (Eq. 4). The overall process stoichiometry
becomes (Eq. 1a+Eq. 4).
NH4+ + 0.75 O2 + HCO3-→ 0.5N2 + CO2- + 2.5H2O
(5)
The combined partial nitritation-anammox process requires 63% less oxygen and no
additional organic carbon source compared to conventional nitrification-denitrification over
nitrate. The partial nitritation and anammox processes can take place in a single reactor in
which ammonium oxidizing bacteria and anammox coexist in a biofilm or form compact
granules.
Batch experiments and microbial analysis showed that nitrite is at the outer biofilm layer
under aerobic conditions. The remaining ammonium and nitrite diffuse into the deeper part of
the biofilm where anoxic conditions are maintained and nitrite acts as an electron acceptor,
reacting with the remaining ammonium to form nitrogen gas (Koch et al., 2000).
The success of partial nitritation-anammox process depends first of all on the continuous
suppression of nitrite oxidizers and, secondly on the produced nitrite to ammonium ratio,
which should be about 1.32 (stoichiometric ratio, see Eq.4). The increase of either the
ammonium or nitrite concentration has an adverse influence on the anammox activity.
Dapena-Mora et al. (2007) found that higher ammonium and nitrite concentration reduced the
performance of anammox bacteria. Jung et al. (2007) described a decrease of anammox
bacterial activities with increasing free ammonium concentration.
6
Chapter II: Literature review
3. Reactor conditions affecting biological nitrogen removal
This section consist effect of oxygen concentration (subsection 2.1), temperature (subsection
2.2) and pH (subsection 2.3) on biological nitrogen removal through various pathways for
biological nitrogen removal.
3.1. Oxygen concentration
The dissolved oxygen concentration is very important for both ammonium and nitrite
oxidation. It becomes a limiting factor for nitrification when it is lower than 2 mgO2L-1
(Beccari et al., 1992). Due to higher oxygen affinity, at low oxygen concentration level the
ammonium oxidizers are more vigorous then the nitrite oxidizers. In other words, oxygen
deficiency influences the performance of nitrite oxidizers more significantly than the
ammonium oxidizers (Philips et al., 2002). This is explained by the oxygen half saturation
constant. Hunik et al. (1994) found that the half saturation constant for dissolved oxygen was
0.16 mgO2L-1 for ammonium oxidizers and 0.54 mgO2L-1 for nitrite oxidizers. In the
activated sludge processes, oxygen half saturation constants of ammonium oxidizers and
nitrite oxidizers were 0.25 – 0.5 and 0.34 – 2.5 mgO2L-1 respectively (Barnes and Bliss,
1983). A reason for this variability is that the oxygen concentration inside the sludge matrix
and in the bulk liquid is not same. As a result, the half saturation constant depends on a
number of parameters such as biomass density, the size of sludge matrix, the mixing intensity
and the rate of diffusion of oxygen into the sludge matrix (Munch et al., 1996; Manser et al.,
2005).
It is possible to remove nitrogen through nitrification-denitrification over nitrite by
controlling the dissolved oxygen concentration. High oxygen levels favor nitrite oxidizers,
resulting in nitrate formation. In oxygen limiting conditions nitrite oxidizers are outcompeted
and nitrite accumulates. This is demonstrated by Peng et al. (2004) in a sequencing batch
reactor and by Jubany et al. (2009) in an activated sludge system. Nitrogen removal over
nitrite can be established by turning off aeration at the point where the ammonium oxidation
has completed. Hidaka et al. (2002) reported an aeration pattern to control ammonium to
nitrite. By frequently changing between aerobic and anoxic in an activated sludge system, the
nitrate formation can also effectively be prevented (Yoo et al., 1999). The aeration was
turned off before all the ammonium was consumed and nitrite started to be converted to
nitrate.
7
Chapter II: Literature review
The anammox process is a strictly anaerobic process and is inhibited by oxygen
concentration. The anammox metabolism is reversible at low oxygen concentration (0.25-2%
air saturation) but irreversible at high concentration (higher than 18% air saturation) (Egli et
al., 2001).
Bulk oxygen concentration is a very important controlling variable for partial nitritationanammox process. In partial nitritation, oxygen is needed for converting half of the
ammonium to nitrite but the conversion to nitrogen gas from unconverted ammonium and
nitrite through anammox process is completely anaerobic. At high oxygen concentration
nitrate formation prevails. Volcke et al. (2010) demonstrated that in partial nitritationanammox process nitrite was converted to nitrogen gas at low bulk oxygen concentration.
The same was observed by Hoa et al. (2002).
3.2. Temperature
Temperature affects the nitrification process directly as well as indirectly. A higher
temperature increases the microbial growth rate according to the Arrhenius law, which is
valid up to a certain critical temperature, above which biological activity starts to decrease.
Grunditz and Dalhammar (2001) found an optimum temperature of 35°C for ammonium
oxidizers and 38°C for nitrite oxidizers. Van Hulle et al. (2007) reported a maximum oxygen
uptake rate by ammonium oxidizing bacteria in the temperature range between 35 and 45°C.
Hellinga et al. (1998) mentioned that above 25°C the specific growth rate of ammonium
oxidizing bacteria become higher than that of nitrite oxidizing bacteria in a SHARON
process.
The optimal temperature for anammox bacteria is reported between 30 - 40°C (Strous et al.,
1999; Egli et al., 2001). Dosta et al., (2008) indicated that temperature of 45°C or higher
causes irreversible loss of efficiency of anammox bacteria. On the other hand the anammox
process can be successfully operated at temperature as low as 20°C (Cema et al., 2007; Isaka
et al., 2007). In this case slow adaptation of anammox bacteria to low temperature is very
important.
Temperature makes an indirect effect on biological nitrogen removal process by participating
in free ammonium and nitrous acid accumulation. Anthonisen et al. (1976) made
mathematical expressions (Eq. 6 and 7) for calculating the amount of free ammonia and
8
Chapter II: Literature review
nitrous acid based on total ammonium (TAN) and total nitrite (TNO2) and incorporating with
temperature (T) and pH:
(6)
(7)
According to these equations, the amount of free ammonia increases with increasing
temperature while the amount of nitrous acid decreases.
The effect of temperature on biological nitrogen removal from wastewater was examined by
Komorowska- Kaufman et al. (2006) in the temperature range from 7.8 to 21°C. They related
influence of temperature on a nitrification-denitrification to sludge age. Temperatures above
15°C are favorable for nitrification even when the sludge age was very short. For a
temperature below 15°C and sludge age lower than 20 days, the nitrification process became
unstable and the removal efficiency varied between 61.7 to 99.3%. They also found that the
unfavorable effect of low temperature (below 15°C) was reduced and stabilized nitrification
process was achieved when the sludge age was more than 20 days. Yamamoto et al. (2006)
performed partial nitritation in a ‘swim-bed’ reactor. In this study, a stable efficiency was
maintained between 15 to 30°C but the performance suddenly deteriorated below 15°C.
3.3. pH
During the conversion of one mole of ammonium to one mole of nitrogen through
nitrification-denitrification over nitrate one mole of H+ is produced. As a result, sufficient
alkalinity is required for buffering the produced protons in wastewater.
The optimum pH for both ammonium oxidizers and nitrite oxidizers lies between 7 and 8
(van Hulle et al., 2010). The ammonium oxidizers prefer a slightly alkaline environment as
these organisms use ammonia (NH3) as substrate (Suzuki et al., 2974). It maintains the
inorganic carbon (HCO3-) that is important for metabolism of nitrifying bacteria. Hellinga et
al. (1998) detected that the growth rate of nitrite oxidizers were decreased by a factor 8 for
the pH change from 8 to 7 whereas the change of the growth rate of the ammonium oxidizer
were negligible.
9
Chapter II: Literature review
Anammox bacteria can grow in a pH range from 6.7 to 8.3. Strous et al. (1999) mentioned an
optimum pH of 8.0. Jung et al. (2007) reported that it is important to keep free ammonia
below 2 mgN.L-1 and free nitrite nitrogen below 35 mgN.L-1 for continuous growth of
anammox bacteria. Below these levels the anammox activity increases gradually in an
anaerobic condition.
The pH also influences the concentration of free ammonia (NH3) and free nitrous acid
(HNO2), which are the actual substrates for ammonium oxidation and nitrite oxidation
respectively and also inhibit nitrification (Anthonisen et al., 1976).
In general nitrite
oxidizing bacteria are more sensitive to free ammonia and nitrous acid inhibition ammonium
oxidation. According to eq. 6 and 7, pH has influence on NH4+/NH3 and HNO2/NO2equilibrium. The amount of nitrite (NO2-) increases with increasing pH. At high pH (>8), free
ammonia becomes the main inhibitor for the nitrification process; at low pH (<7.5) nitrous
acid is the main inhibitor.
Nitrite plays a very critical role in biological nitrogen removal process as it may cause severe
substrate limitation for nitrite oxidizing bacteria at low concentration. High nitrite
concentration inhibits anammox activities. Inhibition starts at nitrite concentrations higher
than 100 mgN.L-1 (Strous et al., 1999) and microbial activities are completely lost at or above
185 mgN.L-1 (Egli et al., 2001).
The optimum pH for nitrification is 8 and the nitrification rate abruptly decreases below a pH
of 6.5 (Shammas 1986). The pH interval for anammox process is 6.7 – 8.3 whereas pH 8.0 is
considered as optimum.
4. Relation between influent organic carbon (COD/N ratio) and
biological nitrogen removal
In systems for biological nitrogen removal from wastewater, autotrophic and heterotrophic
bacteria coexist. In case of conventional nitrogen removal through nitrification-denitrification
over nitrate, nitrification is autotrophic but denitrification is heterotrophic and requires
external organic carbon. In case of completely autotrophic nitrogen removal through partial
nitritation-anammox, no organic carbon source is required. However, even if the influent
does not contain organic carbon, heterotrophic growth is possible on organic material
10
Chapter II: Literature review
generated from biomass decay (Lackner et al., 2008) and/or on excretion of the living cells
(Rittmann et al., 2002).
Matsumoto et al. (2010) studied an autotrophic biofilm process for ammonium oxidation to
nitrite. The process behavior was without any external carbon source and heterotrophic
growth was based on decay of nitrifying bacteria. The resulting biomass distribution profile
in a nitrifying granule (Figure 1) reveals that 22% of the microbial community is heterotrophs
and 68% nitrifying bacteria (ammonium oxidizing and nitrite oxidizing).
Figure 1. Microbial community composition for the nitrifying granule as determined by
quantitative FISH (Matsumoto et al., 2010).
Figure 2. Effect of influent COD concentration on the concentration of the heterotrophic
biomass in nitrification-denitrification over nitrate system (Moussa et al., 2005).
11
Chapter II: Literature review
Heterotrophic bacteria in the treatment system do not only consume COD but also generate
some COD by decay. Moussa et al. (2005) examined the simultaneous effect of influent COD
and sludge retention time (SRT) on the heterotrophic biomass fraction in a nitrifying SBR
system (Figure 2). They mention that the influent COD yields about 40% of the total
heterotrophic biomass and the remaining 60% results from decay for 10 mg.L-1 influent COD
and 30 days SRT. They also found that the heterotrophic biomass increased by 11% with
increasing SRT from 30 to 100 days resulting from increasing decay product with SRT.
4.1. Nitrification-denitrification over nitrate
The influent COD/N ratio is a very important factor for the biological nitrogen removal
through conventional nitrification-denitrification over nitrate. It affects both nitrifying and
denitrifying bacterial population growth in the system.
Yang et al. (2004) observed that in a granular sludge reactor the performance of both
ammonium oxidizing bacteria and nitrite oxidizing bacteria significantly increased with a
decreasing influent COD/N ratio from 20 to 3.3. They also found that the specific oxygen
utilization rate of nitrifying bacteria increased with decreasing COD/N ratio level whereas the
specific heterotrophic oxygen utilization rate tended to decrease. It implied that higher
COD/N ratio is favorable for heterotrophic population.
At high organic carbon, heterotrophic bacteria grew excessively and competed with
ammonium oxidizing bacteria for oxygen. This reduced the nitrification process and
ultimately increased the ammonium concentration in the effluent. Moreover high
concentration of organic compounds also stimulated the biofilm growth as well as increased
the diffusion resistance of ammonium into the biofilm. This also reduced the nitrification. At
high nitrogen levels the nitrifying bacteria were competitive with heterotrophs for oxygen and
the nitrifying bacteria became an important component of the aerobic granules.
Carrera et al., (2004) estimated the effect of COD/N ratio on the nitrification rate in a process
of nitrification-denitrification over nitrate. They found an exponential decrease of
nitrification rate with changing the COD/N ration from 0.71 to 3.4 and the relation defined by
an exponential mathematical expression (Eq. 8)
rnitrification = 0.0323 + 0.334e(- 1.660(COD/N))
(8)
12
Chapter II: Literature review
The influent COD/N ratio not only affect the nitrification rate but also the nitrification
capacity. The nitrifying biomass fraction in a biofilm was increase with decreasing COD/N
ratio (Rittmann et al., 1999). Harremoes et al. (1995) evaluated the autotrophic biomass
fraction for an activated sludge system with biological nitrogen removal and found the
autotrophic biomass fraction increase by 1.5 to 2% with decreased COD/N ratio from 3.4 to
2.6 gCOD.gN-1. There is also a relationship between fraction of nitrifying bacteria and the
relationship between biological oxygen demand (BOD5) and total Kjeldahl nitrogen (TKN) in
the influent (EPA., 1975). Carrera et al. (2004) developed a mathematical expression (Eq. 9)
based on obtained experimental data from a pilot scale biological nitrogen removal system,
relating the fraction of nitrifying bacteria with BOD5 and TKN as:
Nitrifiers’ fraction = 0.0265 + 0.508e(-2.39(BOD5/TKN)) + 0.21e(-0.43(BOD5/TKN))
(9)
According to the eq. 9, the nitrifying biomass decreases with increasing BOD5/TKN ratio.
Therefore low organic carbon is required for nitrification.
For heterotrophic denitrification, organic carbon is required. Most types of wastewater
contain some COD that may be used for the nitrogen production. Carrera et al. (2004)
reported that the nitrification rate remained constant (0.032 gN.gVSS-1 per day) at COD/N
ratio higher than 4 gCOD.gN-1 even though a ratio of at least 7.1 was required to achieve
complete denitrification. They also found that the denitrification percentage had a linear
relation with the COD/N ratio when it was below 7.1.
Hsieh et al. (2003) experimentally revealed that the nitrification and denitrification efficiency
decreased with increasing influent ammonium loading from 2.0 to 11.5 gNm-2d-1 in a biofilm
reactor which could have resulted from limited surface area of the biofilm causing
insufficient reaction site. But nitrification and denitrification rates increased to a peak value
and then decreased at the highest ammonium loading. At highest ammonium loading, some
part of it transformed into free ammonia, which is toxic to most microorganisms and
decreased the nitrification and denitrification rates. They also showed that the nitrification
efficiency also decrease with COD concentration while the denitrification efficiency
increased. Vrtovšek and Roš (2006) performed an experiment in which ground water was
treated in a biofilm reactor; they found minimal nitrite, nitrate and residual COD
concentrations in the effluent for an influent COD/N ratio 3.7. A higher influent COD/N ratio
13
Chapter II: Literature review
led to higher residual COD concentration in the effluent, while a lower influent COD/N ratio
caused incomplete denitrification.
For the simultaneous removal of organic compounds and nitrogen from wastewater, the
membrane aerated biofilm reactor (MABR) was considered as an advanced technology
(Lackner et al., 2008; Satoh et al., 2004; Semmens et al., 2003). In a MABR, the biofilm
grows on a membrane through which oxygen is supplied, while substrate diffuses from the
bulk liquid through the other side of the biofilm. The satisfactory removal of COD and
nitrogen largely depends on the oxygen concentration in the gas stream and the influent
COD/N ratio. Liu et al. (2010) described the effect of substrate COD/N ratio on
denitrification for membrane aerated biofilm reactor and found 96% ammonium removal for
the ratio 3. The effluent nitrate (NO3-) sharply decrease with increasing the COD/N ratio to 5
whereas other substances remained same as ratio 3 and COD removal, nitrification and
denitrification efficiency reached 85, 93 and 92% respectively. When the COD/N ratio was
further increased to 6, the effluent ammonium concentration increased very rapidly.
4.2. Anammox process
The anammox process does not require organic carbon source. A number of studies report
that the presence of organic matter has a negative effect on the anammox processes
(Chamchoi et al., 2008; Guvan et al., 2005; Jianlong and Jing, 2005; Sabumon, 2007; Tang et
al., 2010). If certain amounts of organic carbon are present the growth rate of denitrifiers is
higher than the one of anammox bacteria (Strous et al., 1999), such that anammox bacteria
cannot compete with denitrifiers. Lowering the influent COD/N ratio can control denitrifiers
and results in higher nitrogen removal through anammox process.
4.3. Partial nitritation-anammox
Lackner et al. (2008) performed a simulation study regarding the effect of heterotrophic
growth on autotrophic nitrogen removal through partial nitritation-anammox process. In their
simulations they found that by including the heterotrophic growth on decay biomass only the
nitrogen removal efficiency decreased for the counter diffusion biofilm model but no
significant difference was found for co-diffusion. In the counter diffusion model, anammox
denitrification dominates at COD/N ratio of 0 but at the COD/N ratio equal or higher than 2
14
Chapter II: Literature review
the autotrophic denitrification disappears completely. Under increasing COD load anammox
bacteria are outcompeted by denitrifying heterotrophic bacteria and nitrogen removal is due
to heterotrophic denitrification. In the co-diffusion system the anammox microbial fraction is
almost constant but the heterotrophic bacteria slightly increase and ammonium oxidizing
bacteria decrease with COD/N ratio.
5. Conclusions
Biological nitrogen removal techniques are widely applied to treat nitrogen containing
wastewaters. Among different treatment options, partial nitritation-anammox process is more
sustainable than conventional nitrification-denitrification over nitrate.
The success of operation of partial nitritation-anammox and also nitrification-denitrification
process depends on influent characteristics and operating parameters of the biological
nitrogen removal process such as COD/N ratio and organic carbon concentration, oxygen
concentration, temperature, pH etc.
The highest effectiveness of nitrogen removal for partial nitritation-anammox process is
achieved at lower COD/N ratio. The anammox bacteria are anaerobic bacteria and are
inhibited by dissolved oxygen. A lower bulk oxygen concentration is important for the
oxidation of half of the ammonium in partial nitritation-nammox process. For successful
biological nitrogen removal it is required to maintain the temperature within a certain range.
The nitrogen removal is relatively higher in the presence of heterotrophic bacteria.
Heterotrophic bacteria can grow on influent COD. But the success of heterotrophic nitrogen
removal is observed up to a certain value of COD/N ratio. There are also possibilities to
inhibit the ammonium oxidizing bacteria by increasing nitrite concentration by the
heterotrophic bacteria at low pH. So an optimum pH level must be maintained for getting the
best performance.
15
Chapter III: Model development
In this study, an existing model for autotrophic nitrogen removal in a granular sludge reactor
was extended to include the influence of heterotrophs on the reactor performance. The
granular sludge reactor model was based on a previous model (Volcke et al., 2010), in which
the heterotrophic growth was neglected.
1. Process stoichiometry and kinetics
The model of Volcke et al. (2010) was extended to evaluate the effect of heterotrophic
activities on autotrophic nitrogen removal through partial nitritation-anammox process in a
granular sludge reactor. Four different groups of bacteria were considered: ammonium
oxidizing bacteria (XAOB), nitrite oxidizing bacteria (XNOB), anammox bacteria (XAN) and
heterotrophic bacteria (XH). Nitrification is described as a two-step process: ammonium
oxidation to nitrite by XAOB followed by nitrite oxidation to nitrate by XNOB. Anammox
bacteria convert ammonium and nitrite to nitrogen gas (SN2A). Growth of heterotrophic
bacteria takes place under aerobic as well as anoxic (in presence of NO2- and/or NO3-)
conditions. Heterotrophic growth relies on organic carbon, which is either present in the
reactor influent or results from biomass decay. In case of no readily biodegradable suspended
or particulate organics or soluble organic substrate (Ss) in the influent, heterotrophic growth
only results from decay material (dead biomass). In absence of dissolved oxygen, nitrite
(NO2-) or nitrate (NO3-) is used as an electron acceptor in heterotrophic growth. Therefore
three types of heterotrophic bacteria were considered: aerobic heterotrophs (XH,A) on soluble
organic substrate (SS), anoxic heterotrophs (XH,NO2) on SS and NO2- and anoxic heterotrophs
(XH,NO3) on SS and NO3-. In heterotrophic processes XH,NO3 reduce nitrate to nitrite whereas
XH,NO2 nitrite to nitrogen gas (SN2H).
Biomass decay has been modeled according to the death-regeneration concept instead of the
endogenous respiration approach followed by Volcke et al. (2010). The death-regeneration
concept includes a transition of living cells to substrate together with a fraction of inert
material by decay of microorganism and/or hydrolysis (van Loosdrecht and Henze, 1999).
All decay processes follow first order kinetics and convert biomass to inert and particulate
organics. Hydrolysis of particulate organics makes soluble organic substrate that is utilized
16
Chapter III: Model description
by the heterotrophic bacteria. Within the steps of decay and hydrolysis, decay rather than
hydrolysis is the rate limiting step (personal communication with Mark van Loosdrecht on
November 2010). Moreover hydrolysis compile with decay in death-regeneration concept.
Therefore, in this model hydrolysis is not considered and soluble organic substrate is
generated directly from the decay of biomass.
The stoichiometric matrix format is outlined in Tables 1 and Table 2 gives the process rate
expressions. The values for kinetic and stoichiometric parameters were based on literature
and are summarized in Tables 3 and 4. Ten processes are included in the model. The
autotrophic process comprises growth of XAOB, XNOB and anammox and decay of them and
heterotrophic process includes growth and decay of XHA, XH,NO2 and XH,NO3. The growth of
XAOB, XNOB and anammox were based on Hao et al. (2002) and heterotrophs was based on
ASM1 (Gujer 1999). Like ASM1, decay of XAOB, XNOB, anammox and heterotrophs were
expressed as a death-regeneration concept (Henze et al. 2000). The model stoichiometry and
kinetics were based on the ones from Koch et al. (2000) and Hao et al. (2002).
17
Chapter III: Model description
Table 1. Stoichiometric matrix Aij
Aij
i component
→
j process
SS
[gCOD.
m-3]
SNH
[gN.m-3]
SNO2
[gN.m-3]
SNO3
[gN.m-3]
SO2
[gO2.m-3]
SN2A
[gN.m-3]
↓
growth
1. growth of
XAOB
2. growth of XNOB
3. growth of
anammox
4. aerobic growth
of heterotrophs
6. anoxic (on
NO2-) growth
of heterotrophs
7. anoxic (on
NO3-) growth
of heterotrophs
decay
8. decay of XAOB
9. decay of XNOB
10. decay of XAN
11. decay of XH
composition matrix
gCOD/unit comp
gN/unit comp
-1/YH
-1/YAOB iNXB
1/YAOB
-iNXB
-1/YNOB
1/YNOB
-1/YAN- iNXB
-(1/YAN )(1/1.14)
1/1.14

1
YH, NO 2
1
YH, NO3
1-fI
1-fI
1-fI
1-fI
1
iNSS
XNOB
[gCOD.
m-3]
XAN
[gCOD.
m-3]
11.14/YNOB
1  YH, NO3
XH,NO2
[gCOD.
m-3]
XH,NO3
[gCOD.
m-3]
1
1
1
1 YH, NO 2
1.71 YH, NO 2
2
1.14 YH, NO3
XH,A
[gCOD.
m-3]

1  YH, NO3
1
1.14 YH, NO3
-1
fI
-1
fI
-1
-3.43
1
XI
[gCOD.
m-3]
1
2/YAN
2
XH [gCOD.m-3]
1
iNXB - fI iNXI
– (1-fI) iNSS
iNXB - fI iNXI
– (1-fI) iNSS
iNXB - fI iNXI
– (1-fI) iNSS
iNXB - fI iNXI
– (1-fI) iNSS
0
1
XAOB
[gCOD.
m-3]
1-1/YH
H, NO
-iNXB+1/ YH. 
1.71 YH, NO
iNSS
-iNXB+1/ YH.
iNSS
SN2H
[gN m-3]
13.43/YAOB
-iNXB+1/ YH.
iNSS
1 Y

SN2 [gN.m-3]
-4.57
1
-1
0
-1.71
1
1
iNXB
1
iNXB
1
iNXB
fI
-1
fI
1
iNXB
1
iNXI
18
Chapter III: Model description
Table 2. Kinetic rate expressions
j process
↓
1. growth of
XAOB
AOB
G,AOB =  max

SO 2
K OAOB
2
 SO 2

S NH
AOB
K NH
 S NH
 X AOB
2. growth of
XNOB
NOB
G,NOB =  max

SO 2
S
S
 NOB NO2
 NOBHNH
. X NOB
K
 S O 2 K NO2  S NO2 K NH  S NH
3. growth of
anammox
AN
G,AN =  max

K OAN2
4. growth of
aerobic
heterotrophs
5. anoxic growth
(on NO2-) of
heterotrophs
6. anoxic growth
(on NO3-) of
heterotrophs
H
G,H =  max

NOB
O2
K OAN2
S
S
 AN NH
 AN NO2
 X AN
 S O 2 K NH
 S NH K NO
2  S NO 2
SS
S
S
 H O2
. NOBHNH
 XH 
K  S S K O 2  S O 2 K NH  S NH
H
S
AG,HNO2 =
H
 max
 η NO2
K OH2
S
S NO2
S
S
 H NO2

 H S
. NOBHNH
 XH
H
K O 2  S O 2 K NO2  S NO2 S NO2  S NO3 K S  S S K NH  S NH
AG,HNO3 =
H
 max
 η NO3
K OH2
S
S NO3
S
S
 H NO3

 H S
. NOBHNH
 XH 
H
K O 2  S O 2 K NO3  S NO3 S NO2  S NO3 K S  S S K NH  S NH
7. decay of XAOB D,AOB = b AOB X AOB 
8. decay of XNOB D,NOB = b NOB X NOB 
9. decay of
anammox
10. decay of
heterotrophs
D,AN = b AN X AN 
D,HA = bH X H 
19
Chapter III: Model description
Table 3. Stoichiometric and kinetics parameters values
parameter
value
Unit
Stoichiometric parameters
YAOB
0.20
g COD.g-1 N
Wiesmann, 1994 (1)
YNOB
0.057
g COD.g-1 N
Wiesmann, 1994 (1)
YAN
0.17
g COD.g-1 N
Strous et al (1998) (2)
YH
0.67
g COD.g-1 COD
Henze et al (2000) (ASM1)
YH,NO2
0.53
g COD.g-1 COD
Muller et al (2003)
YH,NO3
0.53
g COD.g-1 COD
Muller et al (2003)
iNXB
0.07
g N.g-1 COD
Assumed in this study
iNXI
0.07
g N.g-1 COD
Assumed ame as iNXB
iNSS
0.03
g N.g-1 COD
Henze et al (2000) (ASM3)
fI
-1
0.08
g COD.g COD
Henze et al (2000) (ASM1)
AOB
 max
1.36
d-1
Hellinga et al (1999) (3)
NOB
 max
0.79
d-1
Hellinga et al (1999) (3)
AN
 max
0.052
d-1
Strous et al (1998) (3)
H
 max
12
d-1
Henze et al (2000) (ASM1) (4)
K AOB
NH
1.1
g N.m-3
Wiesmann (1994) (5)
NOB
K NO
2
0.51
g N.m-3
Wiesmann (1994) (5)
0.03
-3
kinetic (at 30°C)
K AN
NH
AN
K NO
2
g N.m
AOB
AN
: K NH
Assumed, such that ratio K NH
is about the same as in Hao et al (2002)
0.005
-3
g N.m
NOB
AN
Assumed, such that ratio K NO
2 : K NO 2
is about the same as in Hao et al (2002)
KH
NO 2
0.3
g N.m-3
Alpkvist et al (2006)
KH
NO 3
0.3
g N.m-3
Alpkvist et al (2006)
KH
S
20
g COD.m-3
Henze et al (2000) (ASM1)
K AOB
O2
0.3
g O2.m-3
Wiesmann (1994)
K NOB
O2
1.1
g O2.m-3
Wiesmann (1994)
K AN
O2
0.05
g O2.m-3
Assumed in this study
KH
O2
0.2
g O2.m-3
Henze et al.(2000) (ASM1)
20
Chapter III: Model description
bAOB
0.068
d-1
H
AOB
Assumed, set such that bAOB:  max
= bH:  max
bNOB
0.04
d-1
NOB
H
Assumed, set such that bNOB:  max
= bH:  max
bAN
0.0026
d-1
AN
H
Assumed, set such that bAN:  max
= bH:  max
bH
0.6
d-1
H
Assumed  max
/ 20 for this study
ηNO2=ηNO3
0.8
-
Henze et al. (2000) (ASM1)
DNH4
1.5x10-4
m2.d-1
DNO2
1.4x10
-4
DNO3
mass transfer
Williamson and McCarty P.L. (1976)
m .d
-1
Williamson and McCarty P.L. (1976)
1.4x10-4
m2.d-1
Williamson and McCarty P.L. (1976)
DO2
2.2x10-4
m2.d-1
Picioreanu et al. (1997)
DN2
2.2x10-4
m2.d-1
Williamson and McCarty P.L. (1976)
DS
1x10-4
m2.d-1
Hao and van Loosdrecht (2004)
(1)
2
after unit conversion, using a typical biomass composition of CH1.8O0.5N0.2,
corresponding with 1.3659 g COD.g-1
(2)
after unit conversion, using a anammox biomass composition of CH2O0.5N0.15, (Strous
et al., 1998) corresponding with 36.4 g COD.mole-1 or 1.51 g COD.g-1
(3) Conversion of values given by Hellinga et al. (1999) at 35°C and by Strous et al.
(1998) at 32.5°C to 30°C using the relationship (written for XAOB, analogous for XNOB
and XAN)

AOB 1
max
(T )  
AOB 1
max
( Tref
 EaAOB  T  Tref  

)  exp 
 R T T

ref


with E aAOB =68 kJ.mole-1 ; EaNOB =44 kJ.mole-1; EaAN = 70 kJ.mole-1 (Strous et al.,
1999); R=8.31 J.mole-1.K-1.
(4)
Conversion of ASM1-values given by Henze et al. (2000) at 10°C and 20°C to 30°C
using temperature relationship proposed by these authors (ASM3).
AOB
-3
(5) Calculated value at T=30°C and pH=7 from K NH
3 = 0.028 g NH3-N.m and from
NOB
-5
-3
K HNO
2 = 3.2x10 g HNO2-N.m considering the T and pH dependency of the
chemical equilibrium NH 4  NH 3  H  and HNO2  NO2  H 
21
Chapter III: Model description
Table 4: Temperature dependent kinetic parameters
Temperature
10°C
15°C
20°C
25°C
35°C
40°C
Parameters
AOB
 max
(1)
0.201
0.333
0.541
0.865
2.11
3.22
NOB
 max
(1)
0.117
0.194
0.314
0.502
1.23
1.872
0.0077
0.0127
0.021
0.033
0.081
0.123
3
4.24
6
8.49
16.97
24
AN
 max
(1)
H
 max
(2)
bAOB
(3)
0.01
0.017
0.027
0.043
0.105
0.161
bNOB
(3)
0.0059
0.0097
0.016
0.025
0.061
0.0936
0.0004
0.00064
0.001
0.0017
0.00403
0.0062
0.15
0.212
0.30
0.424
0.848
1.2
bAN
(3)
bH (4)
(1) Conversion of values given by Hellinga et al. (1999) at 35°C and by Strous et al.
(1998) at 32.5°C to different temperature using the relationship (written for XAOB,
analogous for XNOB and XAN)

AOB 1
max
(T )  
AOB 1
max
( Tref
 EaAOB  T  Tref  

)  exp 
 R T T

ref


with E aAOB =68 kJ.mole-1 ; EaNOB =44 kJ.mole-1; EaAN = 70 kJ.mole-1 (Strous et al.,
1999); R=8.31 J.mole-1.K-1.
(2) Conversion of ASM1-values given by Henze et al. (2000) at 10°C and 20°C to
different using temperature relationship proposed by these authors (ASM3).
H
AOB
(3) Assumed, set such that bAOB:  max
= bH:  max
(written for XAOB, analogous for XNOB
and XAN).
H
(4) Assumed  max
/ 20 for this study.
22
Chapter III: Model description
2. Reactor configuration, simulation parameters and initial conditions
A one dimensional biofilm model, only considering radial gradients was set up to describe
the autotrophic and heterotrophic interaction in a granular sludge reactor. The model was
implemented in the Aquasim software (Reichert, 1994). The reactor had a fixed volume of
400 m3. Spherical biomass particles (granules) were grown from an initial radius of 0.10 mm
to a predefined steady state granule radius, rp (0.75mm < rp< 2.75 mm) such that the reactor
eventually contains 100 m3 of particulate material, comprising both active biomass as well as
inert material generated during growth and decay. Growth of the granules was associated
with a decrease in the bulk liquid volume to 300 m3. The oxygen level in the bulk liquid was
controlled at a fixed value (between 0.1 and 4.00 gO2.m-3). The bulk liquid was assumed to
be well-mixed, and external mass transfer limitation was neglected, which simplifies the
evaluation of the simulation results.
Biomass granules were typically quite dense with very small pores, in which no relevant
motion of suspended solids takes place. The granule structure was further assumed to be
rigid, meaning that particulate components were displaced only due to the expansion or
shrinking of the biofilm solid matrix. Besides, the biofilm porosity had been assumed
constant (εW=0.75); its value was determined by the initial fractions of particulate
components
(εXAOBini=0.1;
εXNOBini=εXANini=εXHini=0.05;
εXHAini=εXHNO2ini=εXHNO3ini=εXHini/3;
εXIini=0). The density of autotrophic biomass and particulate inerts (ρA) in the granules were
set to 60000 g VSS.m-3(van Benthum et al., 1995), corresponding to 80000 g COD.m-3 (for a
typical conversion factor of 0.75 g VSS.g-1 COD (Henze et al., 2000)). The density of the
heterotrophs (ρH) was 20000 gVSS.m-3 (van Benthum et al., 1995) which is equivalent to
26666 g COD.m-3.
The reactor behavior had been simulated for an influent containing mainly ammonium, with a
flow rate of 2500 m3.d-1. The ammonium concentration through the process was 300 g N.m-3,
except when the ammonium concentrations were varied from 200 gN.m-3 to 900 gN.m-3 to
find the effect of ammonium surface load on reactor performance (Chapter IV section 3.4). It
was assumed that no nitrite or nitrate was present in the influent. Influent was assumed not to
contain any readily degradable or particulate organic substrate except the part (Chapter IV
section 4) where effects of the influent organic substrate on reactor performance were
analyzed. To find out the effect of readily degradable organic substrate concentration on
reactor performance, the concentration of organic substrate varied from 0 to 1000 gCOD.m-3.
23
Chapter III: Model description
The initial concentrations of soluble compounds in the bulk liquid had been assumed equal to
influent concentrations. The processes were operated at 30°C temperature. The temperature
effect on nitrogen removal was analyzed (Chapter IV section 3, sub-section 3.3) where
temperature was changed from 10 to 40°C. Simulations have been performed for several
years of operation to assure steady state conditions.
24
Chapter IV: Results and discussion
1. Role of heterotrophic growth on nitrogen removal
In order to compare the model with and without heterotrophic growth, the simulations
without any influent organic substrate were run. For making the model without heterotrophic
growth, the processes 4, 5, 6 and 10 in Table 2 are inactivated in the simulation. To
investigate the effect of heterotrophic activities in the partial nitritation-anammox process, the
reactor performance is evaluated in terms of nitrogen removal. The active biomasses and
fraction of three types heterotrophic bacteria within heterotrophic biomass are shown in
Figure 3, 4 and 5 and the dynamic results of the nitrogen removal and nitrogen compounds in
bulk are summerized in Figure 6 and 7.
1.1. Active biomass composition
Without heterotrophic activities, relatively higher ammonium oxidizing bacteria (XAOB) and
the Anammox bacteria are present (Figure 3). At the period of 200 to 600 days the anammox
bacteria is higher in case of without heterotrophic growth that made higher nitrogen removal.
The ammonium oxidizing bacteria (XAOB) increase during first 200 days and then decrease;
first rapidly (until 350 days) and then slowly. During first 600 days anammox is growing very
fast but not enough to convert all the nitrite (NO2-) to nitrogen gas. At steady state higher
fraction of anammox and ammonium oxidizing bacteria are for without heterotrophic growth.
Figure 4 is extracted from figure 3 where y-axis is extended to visualize the fraction of nitrite
oxidizing bacteria (XNOB) for both cases and found higher XNOB for without heterotrophic
growth in the model.
In the model an artificial distinction is made between three types of heterotrophic bacteria
according to the substrate they grow on; aerobic heterotrophs (XH,A) growing on organic
substrate (SS), anoxic heterotrophs on nitrite (XH,NO2) and anoxic heterotrophs on nitrate
(XH,NO3). The evaluations of the fraction of these heterotrophic bacteria are shown in figure 5.
At the beginning of the process, a higher amount of XH,NO2 are present in a granule and it
decreases to zero at 1200 days. A very low fraction of XH,NO3 is present during the first 1000
days and then sharply increases. At steady state no XH,NO2 but XH,NO3 and XH,A present in the
granules.
25
Chapter IV: Results and discussion
Active Biomass (gCODX105)
12
10
8
AOB (without heterotrophs)
NOB (without Heterotrophs)
Anammox (without Heterotrophs)
AOB (with Heterotrophs)
NOB (with Heterotrophs)
Anammox (with Heterotrophs)
Heterotrophic biomass
6
4
2
0
0
400
800
1200
1600
2000
Time (Day)
Figure 3. Comparison of microbial community in a granule for the condition of considering
heterotrophic growth and without considering heterotrophic growth (rp=0.75mm, O2=0.5 gO2.m-3,
0.5
NOB (without Heterotrophs)
NOB (with Heterotrophs)
Heterotrophic biomass
0.4
0.3
Heterotrophic X, rp=0.75 mm; O2=0.5 g.m-3; NH4(in)=300
1
XXH,A
HA
0.2
0.1
XXH,NO2
HNO2
XXH,NO3
HNO3
0.8
fraction
Active Biomass (gCODx105)
SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
0.6
0.4
0.2
0
0
0
0
400
800
1200
1600
500
2000
1000
Time [day]
1500
2000
Time (Day)
Figure 5. The dynamics of heterotrophic
Figure 4. Comparison of nitrite oxidizing microbial community fraction (rp=0.75mm,
bacteria (NOB) in a granule in case of with and O2=0.5 gO2.m-3, SS= 0gCOD.m-3, NH4(in) =300
without heterotrophic growth and heterotrophic gN.m-3, T=30°C).
organism (extended from Figure 3)
26
Chapter IV: Results and discussion
1.2. Competition among active biomass
Different bacteria compete with each other for oxygen and substrate. This has an affect on
nitrogen removal performance. Table 5 shows the microorganisms acting in partial
nitritation-anammox process for both with and without heterotrophic growth in the model and
their competition for oxygen and substrate.
Table 5. Microorganism acting for both with and without heterotrophic growth in model and
their competition for oxygen and substrate.
Without heterotrophs
With heterotrophs
Substrate
XH
XAOB
XNOB
Anammox
XAOB
XNOB
Anammox
XH,A
O2
+
NH4+
+
NO2-
+
+
+
+
+
+
+
XH,NO2
XH,NO3
+
+
+
NO3-
+
+
+
When heterotrophic growth is not taken account (without heterotrophs), ammonium oxidizing
bacteria (XAOB) and nitrite oxidizing bacteria (XNOB) compete with each other for oxygen.
Anammox bacteria compete with XAOB for ammonium (NH4+) and with XNOB for nitrite
(NO2-). But when the heterotrophic growth is taken in the model (with heterotrophs), aerobic
heterotrophs (XH,A) compete with XAOB and XNOB for oxygen and heterotrophs consuming
nitrite (XH,NO2) compete with anammox and XNOB for nitrite. The heterotrophic bacteria have
a very low competition with anammox and XAOB for ammonium. The heterotrophic bacteria
based on nitrate (XH,NO3) do not compete strongly with any autotrophic bacteria, moreover it
produces nitrite from nitrate without consuming any oxygen.
27
Chapter IV: Results and discussion
1.3. Comparison of nitrogen removal performance
From Figure 6, it is found that first 200 days there is no significantly different on nitrogen
removal performance between with and without heterotrophic growth. But at between 200
and 600 days better nitrogen removal for without heterotrophic growth whereas after that
time the process with heterotrophs shows better nitrogen removal. The competitions of
heterotrophic bacteria with anammox for NO2- reduce the anammox growth when
heterotrophic growth is considered in the model. Therefore lower total nitrogen removal is
observed in the period between 200 and 600 days, even though heterotrophic nitrogen
removal is higher due to heterotrophic bacteria on nitrite (XH,NO2). But the amount of
heterotrophic bacteria is very low; 0.32 – 0.14x10-5 gCOD per granule between 200 and 600
days. On the other hand, during this period the difference between anammox bacteria in the
simulation with and without heterotrophic growth condition varies between 1.4 – 2.9 x10-5
gCOD per granule.
At steady state based on soluble compounds, that is after 1200 days, the differences of
anammox and ammonium oxidizing bacteria (XAOB) between two conditions are very low.
Despite the lower fraction of anammox and XAOB in the simulation with heterotrophic
growth, higher nitrogen removal is observed due to lower nitrite oxidizing bacteria (XNOB)
and presence of heterotrophic bacteria on nitrate (XH,NO3). These heterotrophic bacteria
convert the nitrate to nitrite and give advantages to anammox bacteria for higher nitrogen
removal at steady state.
In Figure 7 shows that first 1000 days there is nitrite accumulation for both conditions. These
nitrite accumulations are firstly increase upto 260 days and then decrease. The decreasing of
nitrogen removal after 200 days (Figure 6) is due to this high concentration of nitrite in bulk
that inhibits the anammox bacteria. Due to lower anammox bacteria, higher nitrite
accumulation is in case of with heterotrophic growth. At the steady state, there is nitrate
accumulation and lower accumulation in case of with heterotrophic growth. Lower nitrite
oxidizing bacteria (XNOB) and heterotrophic bacteria on nitrate (XH,NO3) in with heterotrophic
growth are responsible for comparatively lower nitrate accumulation in with heterotrophic
growth
28
Chapter IV: Results and discussion
250
gN.m-3
200
150
100
50
0
0
500
1000
1500
2000
Time (day)
Total N removal without heterotrophs
Total N removal with heterotrophs
Autotrophic N removal with heterotrophs
Only Heterotrophic N removal
Figure 6. Comparison of nitrogen removal performance for with and without heterotrophic
growth (rp=0.75mm, O2=0.5 gO2.m-3, SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
300
Without heterotrophs S_NH
Without heterotrophs S_NO2
Without heterotrophs S_NO3
Without heterotrophs S_S
With heterotrophs S_NH
With heterotrophs S_NO2
With heterotrophs S_NO3
With heterotrophs S_S
gN.m-3 or gCOD.m-3
250
200
150
100
50
0
0
500
1000
1500
2000
Time (day)
Figure 7. The nitrogen compounds in bulk for with and without considering heterotrophic
growth (rp=0.75mm, O2=0.5 gO2.m-3, SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
In this section, it is found that at initial time (100 – 600 days) the nitrogen removal is higher
for without considering heterotrophic growth compare to with considering heterotrophic
growth. Anammox bacteria is the main responsible for this nitrogen removal. If the
heterotrophic growth is considered, it reduce the initial anammox growth as well as nitrogen
29
Chapter IV: Results and discussion
removal due to compition with heterotrophic bacteria for nitrite. But at steadystate higher
nitrogen removal for the condition of heterotrophic growth. Lower amount of nitrite
oxidizing bacteria (XNOB) and heterotrophic bacteria on nitrite (XHNO2) give adventages to
anammox bacteria to make better performance in terms of nitrogen removal.
2. Biomass dynamics and steady state
In this biofilm model, granules are grown from an initial size to predefined steady state
granule size. During this growing period, the composition of the active biomass and
particulate inerts are changing together with their position in the granules. The duration of
growing period depends on initial size of the granule and biomass composition in granules.
2.1. Biomass dynamics in a granule
To evaluate the microbial community inside a granule in time, a simulation is performed in a
model with 0.10 mm initial granule size and a predefined steady state granule size of
0.75 mm, an influent ammonium concentration of 300 gN.m-3 and a bulk oxygen
concentration of 1 gO2.m-3. The results are displayed in Figure 8 to 10.
Figure 8 describes that after 50 to 100 days of starting the process there are very small
amounts of anammox present in the centre. More ammonium oxidizing bacteria (XAOB) are
observed from middle to surface of the granules. So within these 100 days, a large amount of
nitrite is accumulated in bulk (Figure 9a). Moreover within these days the predefined granule
size (0.75mm) is not formed. It takes around 400 days to reach the steady state granular size
(Figure 10). At the earlier time of steady state, anammox bacteria are formed in the centre of
granules but this active biomass is move slowly towards surface of the granule. The nitrite
oxidizing bacteria (XNOB) starts to grow at around 400 days and as a consequenc nitrate starts
to increase. The evaluation of the bulk nitrogen compounds and the biomass composition as
well as the biomass fraction in a granule over time is shown in Figure 9. The amount of
anammox bacteria and nitrogen removal increase up to 2000 days and then anammox bacteria
decrease at a very slow rate to reach the steady state level. The process takes almost 2300
days to reach substrate steady state and 6000 days to reach biomass steady state. At steady
state, it is found that the active parts of the biomass are situated within 0.3 mm depth from the
surface of the granule.
30
Chapter IV: Results and discussion
Time = 100 days
100
80
80
XAOB
60
XNOB
XAN
40
XI
XH
20
0
0
[kg COD.m-3]
[kg COD.m-3]
Time = 50 days
100
0.2
0.3
z [mm]
Time = 500 days
0.4
XAN
XI
XH
20
0.2
0.4
z [mm]
Time = 2000 days
0
0
0.6
80
80
[kg COD.m-3]
XAOB
XNOB
XAN
XI
XH
Xtot
0.2
0.4
z [mm]
Time = 6000 days
XAOB
XNOB
XAN
XI
XH
Xtot
0.2
0.4
z [mm]
Time = 4000 days
0.6
XNOB
XAN
XI
XH
Xtot
0.2
0.4
0.6
z [mm]
Time = 10000 days
100
80
XAOB
[kg COD.m-3]
[kg COD.m-3]
40
0
0
0.6
80
0
0
0.6
XAOB
60
20
100
20
0.4
z [mm]
Time = 1000 days
40
100
40
0.2
60
100
60
Xtot
20
Xtot
0
0
[kg COD.m-3]
[kg COD.m-3]
[kg COD.m-3]
XNOB
40
0
0
XH
80
XAOB
60
20
XI
100
80
40
XAN
40
0
0
0.5
100
60
XNOB
20
Xtot
0.1
XAOB
60
XNOB
XAN
XI
XH
40
20
Xtot
0.2
XAOB
60
0.4
z [mm]
0.6
0
0
XNOB
XAN
XI
XH
Xtot
0.2
0.4
z [mm]
0.6
Figure 8. The profile of biomass and particulate inerts in a granule over time (radius 0.75mm,
bulk oxygen concentration 1.0 gO2.m-3, T=30°C).
31
Chapter IV: Results and discussion
(a)
(b)
SO2=0.99912
[g N.m-3 or g COD.m -3]
300
250
SNH
200
SNO2
SNO3
150
SN2
SS
100
SNtot
50
0
0
2000
4000
6000
time [days]
8000
10000
Figure 9. Evaluation of (a) bulk nitrogen compound and (b) biomass and inert fractions in a
Biofilm thickness [mm]
granule over time (radius 0.75mm, bulk oxygen concentration 1.0 gO2.m-3, T=30°C).
0.8
0.6
0.4
0.2
0
0
200
400
600
800
1000
Time [day]
Figure 10. Evaluation of biofilm thickness over time (radius 0.75mm, bulk oxygen
concentration 1.0 gO2.m-3, T=30°C).
2.2. Influence of initial conditions on the time needed to reach steady state
From the dynamic behavior of the active biomass and the bulk concentration, it is found that
the process takes more than thousand days to reach steady state. But this time requirement
depends on initial conditions of biomass matrix and bulk concentrations. It is observed that
the change of initial biomass composition and initial granular size do not affect the steady
state performance but changes the dynamic behavior and time to reach steady state. Table 6
describes the different initial conditions and time required to reach steady state of bulk
32
Chapter IV: Results and discussion
substrate concentration. Steady states is reached earlier for higher initial granular size. Time
requirement to reach steady state is reduce from 445 days to 117 days with increasing initial
granule size from 0.10 mm to 2.00 mm when all other parameters are same (5% anammox,
bulk oxygen concentration 1.00 gO2.m-3, final granule size 2.00 mm). For 2.00 mm final
granular size with 5% anammox, 0.10 mm initial granule and 1.00 gO2.m-3 bulk oxygen level,
445 days are needed to reach the substrate steady state and the time requirement increase with
deviation of final granule size from 2.00 mm. 240 days are needed for 0.30 gO2.m-3 oxygen
level where the initial and final granule size are 0.10 and 0.75 mm respectively. The deviation
of bulk oxygen concentration from 0.30 gO2.m-3 also increases the time for steady state.
Table 6. Time require for reaching steady state at different initial conditions
Initial
biomass
composition
Bulk oxygen
Initial granule
Final granule size
Time to reach
(gO2.m-3)
size (mm-radius)
(mm-radius)
steady state (days)
0.10
406
0.30
0.75
0.50
XAN = 0.05
1150
0.10
XAOB = 0.10
XNOB = 0.05
XH = 0.05
1.00
2389
1.50
561
2.00
445
2.50
621
1.00
XAN = 0.15
XAOB = 0.02
XNOB = 0.02
XH = 0.02
0.30
1.00
240
2.00
250
2.00
117
12
12
0.75
0.75
12
More than 70% removal efficiency achieved after very first day
If cultured full size granules (initial and final granule size are equal) are used with higher
fraction of anammox (XAN) the steady state will be reached much earlier, even within only a
few days. The full size granules with 15% XAN needs only 12 days to reach 100% efficiency
of the process and 70% efficiency can be achieved just after a day from the starting of the
process.
33
Chapter IV: Results and discussion
3. Influence of operational parameters on the reactor performance
The partial nitritation-anammox process intensity depends on operating parameters such as
bulk oxygen concentration, granular size, temperature, ammonium surface load etc.
Optimization of these parameters is important for getting maximum performance in terms of
nitrogen removal in granular sludge reactor.
3.1. Influence of the oxygen concentration
In partial nitritation-anammox process, the bulk oxygen concentration plays a very important
role on nitrogen removal. Different biomass species are active in the granular sludge reactor.
Both ammonium oxidizing bacteria (XAOB) and nitrite oxidizing bacteria (XNOB) compete for
oxygen but XAOB has higher oxygen affinity than XNOB (substrate affinity constant is higher
for XAOB compared to XNOB). On the other hand the anammox bacteria (XAN) are inhibited
by oxygen that mean the anammox grows in anoxic condition. The distribution of biomass in
the granules and the reactor performance depends on all microbial interactions associated
with oxygen and substrate and on competition for the space in the granules. To find out the
effect of bulk oxygen concentration on nitrogen removal and on microbial community in a
granular sludge reactor, the simulation is run at different oxygen concentrations (0.10 – 4.00
gO2.m-3) with fixed initial granular size (0.10 mm), final granular size (0.75 mm), influent
ammonium concentration (300 gN.m-3) and without any influent organic substrate. The
dynamic and steady state behavior and sensitivity of heterotrophic density are analyzed with
oxygen concentration.
3.1.1. Dynamics of nitrogen removal and steady state biomass profile
The bulk liquid concentrations of nitrogen components with time and corresponding steady
state biomass profile in a granule for various bulk oxygen concentrations are shown in Figure
11. Most of the active biomass (XAOB, XNOB and XH) are present in the outer layer of the
granules (Figure 11b) due to the limitation of the oxygen mass transfer in the aerobic
granules for its large and compact structure.
34
Chapter IV: Results and discussion
(a)
(b)
SO2=0.10 gO2.m-3
SO2=0.099827
250
80
SNH
200
SNO2
150
SNO3
100
SS
SN2
SNtot
50
0
0
2000
4000
6000
time [days]
SO2=0.29959
8000
SNO2
-3
SO2=0.30 gO2.m
SN2
SS
100
SNtot
50
2000
4000
6000
time [days]
8000
150
SNO3
100
SS
[kg COD.m-3]
[g N.m-3 or g COD.m -3]
0.6
XAOB
XNOB
XAN
40
XI
XH
Xtot
0.2
0.4
z [mm]
0.6
SO2=0.49943; time =10000
80
SNO2
SN2
SNtot
50
2000
4000
6000
time [days]
XAOB
60
XNOB
XAN
40
XI
XH
20
8000
10000
0
0
SO2=1.25 gO2.m-3
SO2=1.249
300
Xtot
0.2
0.4
0.6
z [mm]
SO2=1.249; time =10000
100
80
SNH
SNO2
[kg COD.m-3]
[g N.m-3 or g COD.m -3]
0.4
z [mm]
SO2=0.29959; time =10000
60
SO2=0.50 gO2.m-3
SNH
200
0
0
0.2
100
250
50
Xtot
0
0
10000
300
100
XH
20
SO2=0.49943
150
XI
80
SNO3
150
200
XAN
40
0
0
10000
[kg COD.m-3]
[g N.m-3 or g COD.m -3]
SNH
200
250
XNOB
100
250
0
0
XAOB
60
20
300
0
0
SO2=0.099827; time =10000
100
[kg COD.m-3]
[g N.m-3 or g COD.m-3]
300
SNO3
SN2
SS
SNtot
2000
XAOB
60
XNOB
XAN
40
XI
XH
20
4000
6000
time [days]
8000
10000
0
0
Xtot
0.2
0.4
z [mm]
0.6
Figure 11. (a) Evolution of bulk concentration of ammonium (SNH), nitrite (SNO2), nitrate
(SNO3) and nitrogen gas (SN2) and (b) distribution of biomass in the granule at steady state (z
is the distance from the granule centre); ammonium oxidizer (XAOB), nitrite oxidizer (XNOB),
anammox bacteria (XAN), heterotrophs (XH) and particulate inerts (XI) for various bulk
oxygen concentration at fixed granular radius 0.75 mm, initial ammonium concentration 300
gN.m-3 and temperature 30°C.
35
Chapter IV: Results and discussion
The biomass distribution profiles show that ammonium oxidizers (XAOB) are at the outer
surface of the granules, where oxygen and ammonium are available. Nitrite oxidizers (XNOB)
and heterotrophs (XH) also need oxygen to survive and are just below the XAOB layer where
still some oxygen is present. Both XNOB and heterotrophic bacteria are almost in same
position in the granules. The anammox are present in the inner part of the active biomass
layer in granules, just behind the XNOB and heterotrophs. Its position is like to facilitate from
the diffusion of substrate, ammonium and nitrite. The deeper parts of the granules are
particulate inert matter (XI) that is produced by the decay of active biomass. As there is no
external organic carbon, the growth of heterotrophs depends on organic substrate that
produced from decay. The density of heterotrophs using in this model is much lower from the
other microorganism. Therefore total biomass density becomes lower at the position where
heterotrophic bacteria are grown in Figure 11b.
Due to the high oxygen affinity of ammonium oxidizing bacteria (XAOB) and its position in
the granules at the outer surface, ammonium is converted to nitrite (NO2-) very fast. Then the
nitrite is converted to either nitrogen gas (N2) or nitrate (NO3-) based on competition between
anammox bacteria (XAN) and nitrite oxidizing bacteria (XNOB). From the dynamics of the
nitrogen removal (Figure 11a), the nitrite accumulation within the system is found at the
initial stage of the process. This is due to faster growth rate of ammonium oxidizing bacteria
compare to the anammox and/or nitrite oxidizing bacteria. But the amount and period of
nitrite accumulation depend on bulk oxygen concentration. Nitrite accumulation is observed
at bulk oxygen concentration 0.20 gO2.m-3 to 2.00 gO2.m-3. Nitrite accumulation increases
but duration of the accumulation firstly increases and then decreases with oxygen
concentration. Nitrite accumulates during the first 240 days for oxygen level 0.30 gO2.m-3
whereas it lasts for 1528 days for 0.75 gO2.m-3 and 252 days for 1.25 gO2.m-3. At higher
oxygen concentration, the nitrite oxidizing bacteria start to grow earlier and convert the nitrite
to nitrate and reduce the time of nitrite accumulation. Lower oxygen levels facilitate
anammox bacteria to grow fast and reduce the growth of ammonium oxidizing bacteria
(XAOB) that reduce the nitrite accumulation.
3.1.2. Steady state performance and biomass composition
The steady state reactor behaviour in terms of nitrogen removal and biomass fraction are
summarized in Figure 12 for various bulk oxygen concentrations. Nitrogen gas is produced at
low bulk oxygen concentration, while nitrate accumulation is at high oxygen concentration.
36
Chapter IV: Results and discussion
At a very low oxygen concentration, due to lack of oxygen lower ammonium oxidizing
bacteria (XAOB) are formed that reduce the conversion of ammonium to nitrite and the nitrite
oxidizers (XNOB) are completely outcompeted by anammox bacteria. At high oxygen level,
there are high fractions of XAOB and XNOB but anammox are completely outcompeted.
In Figure 12a, there is a clear peak for nitrogen removal at oxygen level 0.30 gO2.m-3. At this
point the total nitrogen gas is 280 gN.m-3 which correspond to 93.4% removal. A small
deviation from this optimal oxygen level results a significant decrease of the nitrogen
removal efficiency. For the variation of 0.10 gO2.m-3 the removal decreases with about 10%
and for a deviation of 0.20 gO2.m-3 it is around 25%. Hao et al. (2002) found the decrease of
nitrogen removal to be about 20% for 0.20 g.m-3 bulk oxygen concentration variation.
(a)
(b)
1
250
200
SNO2
SNO3
150
SN2
100
2
O2 [gO2.m-3]
3
XAOB
0.4
XAN
XNOB
0.2
SS
1
0.6
XI
SNtot
50
0
0
0.8
SNH
fraction
[g N.m-3 or g COD.m -3]
300
4
0
0
XH
1
2
O2 [gO2.m-3]
3
4
Figure 12. Influence of bulk oxygen concentration on steady state reactor performance. (a)
bulk concentration of nitrogen components, and (b) biomass and particulate composition in a
granule (rp=0.75mm, SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
In practice, a small variation of bulk oxygen concentration is quite possible especially for a
big reactor. Therefore careful regulation of oxygen concentration level at the peak point is
crucial to achieve maximum process performance.
At oxygen levels lower than 0.30 gO2.m-3, unconverted ammonium is accumulated due to low
ammonium oxidizing bacteria. For oxygen levels higher than 0.30 gO2.m-3 nitrite oxidizing
bacteria (XNOB) starts to grow, resulting in nitrate production and reduce conversion to
nitrogen gas. Figure 12b reveals that the highest anammox bacteria are at 0.40 gO2.m-3
oxygen level and above this anammox bacteria decrease with oxygen concentration.
Ammonium oxidizing bacteria and the nitrite oxidizing bacteria are increasing with bulk
37
Chapter IV: Results and discussion
oxygen concentration. Therefore only nitrate accumulation is at higher oxygen concentration
(higher than 2.00 gO2.m-3). A higher bulk oxygen concentration also gives advantages to
heterotrophs to grow. Around 2% heterotrophic bacteria are present in a granule at oxygen
levels higher than 1.5 gO2.m-3. They utilize the dissolved organic substrate that is produced
from the decay of active biomass.
3.1.3. Sensitivity analysis for the density of heterotrophs
It is generally accepted that the density of heterotrophs is lower than that one of autotrophs.
In this study the density of heterotrophs (ρH) is 20000 gVSS.m-3 that corresponds to 26666
gCOD.m-3 and density of autotrophs (ρA) 60000 g VSS.m-3 correspond to 80000 g COD.m-3
(van Benthum et al., 1995) have been assumed. Henze et al. (2000) assumed that
heterotrophic biomass density was same as autotrophic biomass density. Therefore attention
should be focused on assessment of the sensitivity of heterotrophic biomass density on
reactor performance and to analyze it a series of simulation have also been performed
assuming the heterotrophic density (ρH) 26666 gCOD.m-3 and 80000 g COD.m-3. The
simulation results of this performance are shown in Figure 13. From this sensitivity analysis
of heterotrophic biomass density it is found that there is almost no effect on the steady state
process performance especially within the desired bulk oxygen concentration range for
nitrogen removal.
300
[gN.m-3]
250
rho_H=80000, NH4+
rho_H=80000, NO2rho_H=80000, NO3rho_H=80000, N2
rho_H=26666, NH4+
rho_H=26666, NO2rho_H=26666, NO3rho_H=26666, N2
200
150
100
50
0
0
0.5
1
1.5
2
O2 [g.m-3]
Figure 13. Influence of heterotrophic biomass density on reactor performance (rp=0.75mm,
SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
38
Chapter IV: Results and discussion
3.2.
Influence of the granule size
The granule size is a key factor affecting the nitrogen removal in a granular sludge reactor
(Volcke et al., 2010). The granule size may fluctuate even at the steady state of reactor
operation and may change the aerobic to anaerobic volume ratio in granules at the fixed bulk
oxygen concentration level (de Kreuk et al., 2007). Therefore, only depending on single
granules and their aerobic and anoxic zone inside the granule may not be reliable for nitrogen
removal and could result in unstable nitrogen removal efficiency (Chen et al., 2011). To
examine the reactor performance, more specifically the microbial community structure in the
granules and nitrogen components in bulk as well as the nitrogen removal efficiency with
granule size, simulations for different granule size were carried out. As the total volume of
the granules is fixed at 100 m3, the number of granules and surface area decrease with
increasing granule radius.
3.2.1. Dynamics of nitrogen compounds and steady state biomass profile
Figure 14a describes the dynamic of the nitrogen components and Figure 14b the steady state
profile of biomass and particulate inerts in the granules for different granule size with a fixed
bulk oxygen level (1.0 gO2.m-3). From the Figure 14b, it is found that at the steady state
profile of biomass and particulate inerts are present within 0.30 mm thickness from the outer
surface of the granules and anammox is present in inner part of this thickness. At very
smaller granule size (radius < 0.25 mm), there is no anammox due to the oxygen penetration.
In this granular size, it is hardly to form anaerobic layer due to easy oxygen diffusion through
the granules. Most of the active biomasses in these granules are ammonium oxidizing
bacteria (XAOB), nitrite oxidizing bacteria (XNOB) and heterotrophs (XH). With increasing
granule size, an anaerobic layer is started to form that increase anammox bacteria.
From Figure 14a, the nitrite accumulation is for an intermediate period of time for 1.00 mm
granular size. But the amount and period of nitrite accumulation is varied with granular size.
Nitrite accumulation is observed at granule size 0.50 mm to 1.75 mm. Above this granule size
there is no nitrite accumulation. For 0.50 mm to 1.00 mm granule accumulation is for
intermediate period and for 1.00 mm to 1.75 mm it is upto steady state.
39
Chapter IV: Results and discussion
(a)
(b)
rp= 0.25 mm
rp = 0.25 mm
100
250
200
SNO2
SNO3
150
SN2
100
SS
SNtot
50
0
0
2000
4000
6000
time [day]
rp = 1.00 mm
8000
10000
XI
0.05
0.1
0.15
z [mm]
0.2
0.25
rp =1.00 mm; time =10000
80
SNH
SNO2
SNO3
150
SN2
100
XAOB
[kg COD.m-3]
[g N.m-3 or g COD.m-3]
XH
rp= 1.00 mm
SS
60
2000
4000
6000
time [day]
XNOB
XAN
40
SNtot
50
XI
XH
20
8000
0
0
10000
rp= 2.00 mm
rp = 2.00 mm
Xtot
0.2
0.4
0.6
z [mm]
0.8
1
rp =2.00 mm; time =10000
100
250
80
SNH
200
SNO2
SNO3
150
SN2
100
XAOB
[kg COD.m-3]
[g N.m-3 or g COD.m-3]
20
Xtot
300
SS
60
2000
4000
6000
time [day]
XNOB
XAN
40
XI
XH
20
SNtot
50
Xtot
8000
rp = 2.50 mm
0
0
10000
rp= 2.50 mm
300
0.5
1
z [mm]
1.5
2
rp =2.50 mm; time =10000
100
250
80
SNH
200
[kg COD.m-3]
[g N.m-3 or g COD.m-3]
XAN
100
200
SNO2
SNO3
150
SN2
100
SS
2000
4000
6000
time [day]
XAOB
60
XNOB
XAN
40
XI
XH
20
SNtot
50
0
0
XNOB
40
250
0
0
XAOB
60
0
0
300
0
0
rp =0.25 mm; time =10000
80
SNH
[kg COD.m-3]
[g N.m-3 or g COD.m-3]
300
Xtot
8000
10000
0
0
0.5
1
1.5
z [mm]
2
2.5
Figure 14. Kinetics of bulk concentration of nitrogen components (a) and biomass
community in a granule (b) for different granule size at oxygen level 1.0 g.m-3, initial
ammonium concentration 300 g.m-3 and 30°C temperature.
40
Chapter IV: Results and discussion
3.2.2. Steady state reactor performance and biomass composition
The overall steady state reactor performance in terms of nitrogen removal and amount of
biomass in a granule are shown in Figure 15. Due to the diffusion of oxygen to a very small
size granule, anammox bacteria are not present. For presence of ammonium oxidizing
bacteria (XAOB) and nitrite oxidizing bacteria (XNOB) and absence of anammox, all the
ammonium is firstly converted to nitrite and followed to nitrate. Therefore the process with
very small size granules is not able to produce nitrogen gas.
According to the figure 15, the anammox bacteria increase with increasing granule size and
therefore nitrogen removal is increase. There is a clear peak point of 2.00 mm granular size
where maximum nitrogen removal is found. At this point maximum nitrogen removal is 261
gN.m-3 that belongs to 87% nitrogen removal for bulk oxygen concentration level 1.00
gO2.m-3. For granules larger than 2.00 mm radius, steady state nitrogen removal is decreased
even though the amount of anammox and XAOB in a granule is higher. This is because of the
lower number of granules in the system at higher granule size whereas total volume of
biomass is fixed (100 m3). That means, lower amount of total anammox and XAOB in the
reactor. Lower number of granules as well as lower XAOB become the limiting factor for the
conversion of ammonium to nitrite. Therefore ammonium accumulation increases and
nitrogen removal decrease with granule size higher than 2.00 mm. Liu et al. (2005) and Chiu
et al. (2007) reported that large-size aerobic granules were not favorable for biological
removal of nitrogen due to potential mass transfer limitation.
(a) Steady state N, SO2 = 1.0 g.m-3; SNHin=300 gN.m-3
(b)
300
-3
250
200
150
100
5
XAOB
SNO2
4
XNOB
SNO3
SN2
SNtot
SS
XAN
3
XI
2
XH
1
50
0
0
x 10
SNH
[gCOD.m-3]
[g N.m-3 or g COD.m -3]
6
0
1
2
Granule radius [mm]
3
-1
0
0.5
1
1.5
2
Granule radius [mm]
2.5
3
Figure 15. (a) Steady state bulk concentration and (b) amount of biomass with granular size
(SO2 = 1.00gO2.m-3, SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
41
Chapter IV: Results and discussion
For intermediate granule size (1.00 – 2.00 mm) and oxygen level 1.0 g.m-3, the nitrite that is
produced by ammonium oxidizing bacteria (XAOB) is not fully to convert either nitrate or
nitrogen gas. There is steady state nitrite accumulation in this intermediate granular size.
Vlaeminck et al., (2010) also found nitrite accumulation by experimental studies and
mentioned that the reduction of XNOB activities with increasing granule size was the main
reason behind this type of nitrite accumulation.
3.2.3. Interaction between granule size and oxygen concentration
In previous subsection, ammonium accumulation is found at higher granule size due to lower
amount of total ammonium oxidizing bacteria (XAOB). Providing more oxygen gives
advantages to XAOB to grow. Therefore increasing bulk oxygen concentration converts more
ammonium to nitrite followed by anammox bacteria to nitrogen gas and increase the nitrogen
removal performance.
100
90
1.5
80
1
70
Oxygen level
N removal
NH4 removal
0.5
0
0
1
2
Maximum
removal (%)
Bulk oxygen (gO2.m-3)
2
60
50
3
Granule radius (mm)
Figure 16. Influence of granule radius on nitrogen removal and corresponding bulk oxygen
concentration (SS= 0gCOD.m-3, NH4(in) =300 gN.m-3, T=30°C).
For 0.75 mm granule the maximum nitrogen removal is observed at 0.30 gO2.m-3 bulk
oxygen concentration level (Figure 12a) and for 2.00 mm size it is 1.00 gO2.m-3 (Figure 15a).
To determine the process optimum granule size with bulk oxygen concentration level for the
maximum nitrogen removal, simulation is demonstrated for 0.75 mm to 2.50 mm granule
radius at series of bulk oxygen concentration levels (0.30 – 2.00 gO2.m-3). The results are
shown in Figure 16. It is observed that for maximum nitrogen removal, higher bulk oxygen is
42
Chapter IV: Results and discussion
required for higher granule size. There is very little difference (93% to 89%) among
maximum nitrogen removal at granule size increase from 0.75 to 2.50 mm. This reveals that
increasing granule size has very low influence on maximum performance of granule sludge
reactor and the stress for changing granular size mostly be minimized by changing oxygen
concentration level.
According to the figure 16, the maximum ammonium (NH4+) removal is varied from 99.5%
to 97.5% with radius chance from 0.75 to 2.50 mm. The difference between the maximum
nitrogen removal and maximum ammonium removal is the accumulation of nitrite and
nitrate. The nitrate (NO3-) accumulation is varied from 14.5 to 16 gN.m-3 and nitrite (NO2-) is
0.60 to 3.00 gN.m-3 with increasing granule size 0.75 mm to 2.50mm.
3.3.
Role of temperature
Temperature is a key parameter in nitrogen removal process because it increases the
microbial efficiency and plays an important role on process performance. But increasing
biological efficiencies are up to a certain temperature level, above which efficiency decreases
and the microorganisms die. In this model the decreasing efficiency and threshold for
microorganism die off due to high temperature is not considered. Therefore, in this study
temperature range is maintained between 10 - 40°C to overcome any error results related to
decreasing microbial efficiency and death at high temperature.
3.3.1. Effect of temperature at fixed oxygen level
The steady state nitrogen removal performances as well as nitrogen compounds in bulk and
composition of microbial community with temperature for the dissolved oxygen
concentration level 0.50 gO2.m-3 are shown in the Figures 17. From the figure 17b, the
anammox bacteria increase up to 20°C and then decrease whereas the ammonium oxidizing
bacteria (XAOB) decrease with temperature. The nitrite oxidizing bacteria (XNOB) increases but
after 35°C it decreases.
The process performance cannot be fully described by the change of microbial community
structure with temperature. Because the process response is not only depends on the amount
of bacteria but also their efficiency that largely change with temperature. Due to lower
microbial efficiency at lower temperature, unconverted ammonium and nitrite are found.
43
Chapter IV: Results and discussion
Increasing temperature increases microbial efficiencies and decreases the ammonium and
nitrite accumulation. From the Figure 17a, the highest performance, in terms of nitrogen
removal is observed for temperature range 15 to 30°C at bulk oxygen concentration 0.50
gO2.m-3. Yamamoto et al. (2006) also mentioned that in partial nitritation process,
successfully started up and maintained with higher performance for nitrogen removal was
between 15 to 30°C temperature and the performance was sharply decreased below 15°C.
Above 30°C temperature the nitrogen removal is decrease due to very low amount of
anammox bacteria. On the other hand nitrate accumulation is increasing with temperature.
The lower decreasing rate of nitrite oxidizing bacteria (XNOB) compared to anammox bacteria
is the main reason behind increasing nitrate accumulation at higher temperature.
(b)
(a)
200
150
100
1
SNH
SNO2
fraction
[g N.m-3 or g COD.m -3]
250
SNO3
SN2
SS
50
0
10
0.8
XAOB
0.6
XNOB
XAN
0.4
0.2
20
30
Temperature [oC]
40
0
10
XI
XH
15
20
25
30
Temperature [oC]
35
40
Figure 17. Response of nitrogen removal process with temperature, (a) steady state nitrogen
compounds and (b) steady state microbial community (rp=0.75mm, SO2= 0.50gO2.m-3, SS=
0gCOD.m-3, NH4(in) =300 gN.m-3).
3.3.2. Interaction of bulk oxygen with temperature
With changing temperature, the bulk oxygen concentration level also plays a very important
role in the reactor performance. Figure 18 shows the relation between temperature and reactor
performance with bulk oxygen concentration. At lower temperature the possibility of nitrite
(NO2-) accumulation is high. The nitrite (NO2-) accumulation decreases and nitrate (NO3-)
accumulation increases with temperature.
Relatively higher oxygen concentration is required to reach maximum nitrogen removal at
lower temperature. The maximum nitrogen removal is observed within bulk oxygen
44
Chapter IV: Results and discussion
concentration range 0.25 to 0.35 gO2.m-3 and the desired temperature range 10 to 40°C. For
the bulk oxygen concentration below 0.35 gO2.m-3, better nitrogen removal is for temperature
40°C compare to 10 and 20°C but above this oxygen level the process with temperature 20°C
shows better performance and for above 1.60 gO2.m-3 oxygen, 10°C shows better in terms of
nitrogen removal. Therefore the nitrogen removal at an optimum temperature is also a
function of bulk oxygen concentration.
300
250
[gN.m-3]
200
150
100
50
0
0
0.5
1
1.5
Bulk oxygen concentration [gO2.m-3]
2
N2 at T=10
NO2 at T=10
NO3 at T=10
N2 at T=20
NO2 at T=20
NO3 at T=20
N2 at T=40
NO2 at T=40
NO3 at T=40
Figure 18: The combined effect of temperature and bulk oxygen concentration on nitrogen
removal (rp=0.75mm, SS= 0gCOD.m-3, NH4(in) =300 gN.m-3).
3.4. Effect of ammonium surface load
The simulation is performed with changing influent ammonium concentration with a series of
bulk oxygen concentration level (0.10 to 1.00 gO2.m-3) to find out the optimum bulk oxygen
level for maximum nitrogen removal. The corresponding results are shown in figure 19. In
this figure, the ammonium concentrations express as ammonium surface load. The influent
ammonium concentrations change from 200 to 900 gN.m-3 at fixed biomass volume (100 m3),
constant granular size (0.75 mm) and constant flow rate (2500 m3.d-1) correspond to 1.25 to
5.625 gN.m-2.d-1 ammonium surface loads.
45
Chapter IV: Results and discussion
From Figure 19, it is clear that for maximum nitrogen removal, required bulk oxygen
concentration increases with ammonium surface load. Total amount of maximum nitrogen
removal increase from 188 gN.m-3 to 654 gN.m-3 for increasing ammonium surface loads
from 1.25 to 5.625 gN.m-2.d-1 but corresponding removal efficiency (N removal) decrease
from 94% to 72%. Hao et al. (2001) did a simulation to test the CANON process in a biofilm
model and found decreasing nitrogen removal efficiency from 90% to 40% for increasing
ammonium surface load from 0.62 to 4.94 gN.m-2.d-1.
100
80
0.6
60
40
0.3
Oxygen level
N removal
NH4 removal
20
0
Maximum
removal (%)
Bulk oxygen (gO2.m-3)
0.9
0
0
2
4
Ammonium surface load
6
(gN.m-2.d-1)
Figure 19. Relationship among ammonium surface loads with corresponding dissolved
oxygen concentration level for maximum nitrogen removal (rp=0.75mm, SS= 0gCOD.m-3,
T=30°C).
In partial nitritation anammox process, oxygen is required to partially convert the ammonium
to nitrite by ammonium oxidizing bacteria (XAOB). Therefore for high ammonium surface
load, high bulk oxygen concentration is desired to get maximum performance in terms of
nitrogen removal. The removal efficiency is highest at lower dissolved oxygen level and
lower ammonium surface load. When the granular surface and size are relatively fixed, the
maximum capacity of nitrogen removal is thus fixed. For the limiting factor of the surface
area with fixed capacity, a higher ammonium surface load is corresponding to lower nitrogen
removal efficiency.
According to Figure 19, ammonium (NH4+) removal efficiency is almost same (99%) for all
ammonium surface loading conditions. The difference between the nitrogen and ammonium
46
Chapter IV: Results and discussion
removal efficiency belongs to nitrite (NO2-) and nitrate (NO3-) accumulation. There is higher
nitrate (NO3-) accumulation compare to nitrite (NO2-) at low ammonium surface load (for
1.875 gN.m-2.d-1 surface load; 1.00 gN.m-3 NO2- and 28.50 gN.m-3 NO3-) but at higher
surface load, nitrite (NO2-) accumulation is much higher (for 5.625 gN.m-2.d-1 ammonium
surface load; 185 gN.m-3 NO2- and 46 gN.m-3 NO3-).
4. Influence of influent organic substrate on reactor performance
In previous simulation no organic substrate included in influent, heterotrophic growth was
only on decay products. In this section influent organic substrate concentrations (0 – 100
gCOD.m-3) is included in the model to analyze the influence of it on heterotrophic growth
and reactor performance in terms of nitrogen removal. Assume that all influent organic
substrate is soluble (SS) and readily biodegradable.
4.1. Effect of organic substrate at fixed oxygen level
An investigation to find out the effect of substrate organic materials on nitrogen removal has
performed and the outcome is exposed in Figure 20. From Figure 12a, it was found that in the
case without any influent organic substrate maximum nitrogen removal was achieved at bulk
oxygen concentration 0.30 gO2.m-3 for an influent containing 300 gN.m-3 ammonium (NH4-)
and with 0.75 mm granule. At these conditions, nitrogen removal is initially increased and
then decreased with increasing influent organic substrate concentration (Figure 20a). For
increasing influent organic substrate concentration level from 0 to 40 gCOD.m-3 the nitrogen
removal increases from 280 gN.m-3 to 294 gN.m-3 that corresponds to 93.4% to 98% removal.
During this process nitrate (NO3-) accumulation decreases. The organic substances give
advantages the heterotrophic bacteria to grow and take part in denitrification. Heterotrophic
bacteria denitrify nitrate to nitrite, which can be further denitrified by heterotrophs or by
anammox. In both cases nitrogen gas (N2) is formed. Further increasing the influent organic
substrate concentration, decreases the nitrogen removal performance and increases
ammonium (NH4-) accumulation.
Figure 20b shows that anammox (XAN) and ammonium oxidizing bacteria (XAOB) are
decreasing whereas heterotrophic bacteria (XH) are increasing with influent organic substrate.
47
Chapter IV: Results and discussion
Increasing heterotrophic growth compete with anammox for ammonium and with ammonium
oxidizing bacteria (XAOB) for oxygen. For influent organic substrate concentrations higher
than 40 gCOD.m-3, bulk oxygen concentration becomes a limiting factor for XAOB, due to
high oxygen consumption by heterotrophic bacteria. It causes ammonium accumulation and it
increases with substrate organic materials.
(a)
(b)
1
250
200
SNO2
150
SNO3
XNOB
SN2
0.6
XAN
0.4
XI
XH
SNtot
100
0.2
SS
50
0
0
XAOB
0.8
SNH
fraction
[g N.m-3 or g COD.m -3]
300
20
40
60
SSin [gCOD.m-3]
80
100
0
0
20
40
60
SSin [gCOD.m-3]
80
100
Figure 20. Influence of input substrate organic materials on steady state reactor performance.
(a) bulk concentration of nitrogen components, and (b) biomass and particulate fraction in a
granule (rp=0.75mm, SO2= 0.30 gO2.m-3, NH4(in) =300 gN.m-3, T=30°C).
4.2. Effect of oxygen concentration at fixed influent organic substrate
The investigation is also performed in case of already existing higher (100 gCOD.m-3)
influent organic substrate. The simulation results are shown in Figure 21. The highest
anammox (XAN) bacteria are found at the bulk oxygen level 0.5 gO2.m-3 (Figure 21b) but the
highest nitrogen removal is observed at a range of bulk oxygen concentration and that is 0.40
to 0.50 gO2.m-3 (Figure 21a).
Below 0.40 gO2.m-3 oxygen level, there is not sufficient oxygen to grow enough ammonium
oxidizing bacteria (XAOB) to convert all the ammonium to nitrite resulting lower anammox
bacteria in a granule. Therefore, when decreasing oxygen concentration from 0.40 gO2.m-3
the nitrogen removal decreases and ammonium accumulation increases. Above this oxygen
level the significant amount of nitrite oxidizers (XNOB) started to grow and reduces the
conversion to nitrogen gas and increase the nitrate accumulation.
48
Chapter IV: Results and discussion
Comparing the nitrogen removal without any influent organic substrate (Figure 12a) at same
level of bulk oxygen concentration, the nitrogen removal efficiency is higher in presence of
influent organic substrate and heterotrophic bacteria plays the vital role for this better
performance.
(a)
(b)
1
250
SNH
200
SNO2
150
100
50
0
0
0.8
XAOB
SNO3
fraction
[g N.m-3 or g COD.m -3]
300
SN2
SNtot
SS
0.5
1
O2 [gO2.m-3]
1.5
0.6
XNOB
0.4
XAN
0.2
XH
0
0
XI
0.5
1
O2 [gO2.m-3]
1.5
Figure 21. Influence of oxygen concentration for a fixed input substrate concentration on
steady state reactor performance. (a) bulk concentration of nitrogen components, and (b)
biomass and particulate fraction in a granule (rp=0.75mm, Ss= 100 gCOD.m-3, NH4(in) =300
gN.m-3, T=30°C).
4.3. Interaction between organic substrate and oxygen concentration
To find out the optimum bulk oxygen level for maximum nitrogen removal with influent
organic substrate, the simulations are performed with changing influent organic substrate
from 40 to 1000 gCOD.m-3 and a series of bulk oxygen concentration level (0.20 to 2.30
gO2.m-3). All corresponding results are shown in Figure 22. According to Figure 22 higher
bulk oxygen concentration is needed for maximum nitrogen removal at higher organic
substrate concentration. The maximum nitrogen removal decrease from 294 gN.m-3 (98%) to
276 gN.m-3 (92%) for increasing organic substrate concentration from 40 to 1000 gCOD.m-3.
High organic substrates produce higher heterotrophic bacteria (XH). Due to the competition
between heterotrophic bacteria (XH) and ammonium oxidizing bacteria (XAOB) high bulk
oxygen concentration is required to get optimum amount of X AOB at high organic substrate.
Therefore to achieve the maximum nitrogen removal efficiency, the bulk oxygen
concentration increases with increasing influent organic substrate.
49
Chapter IV: Results and discussion
Figure 22 also shows that maximum ammonium (NH4+) removal efficiency is almost the
same (99%) for all organic substrate concentration. The difference between the nitrogen and
ammonium removal efficiency is for nitrite (NO2-) and nitrate (NO3-) accumulation. At high
organic substrate concentration, nitrite (NO2-) accumulation is higher than nitrate (NO3-)
accumulation.
100
2
90
1.5
80
1
Oxygen level
N removal
NH4 removal
0.5
0
70
Maximum
removal (%)
Bulk oxygen (gO2.m-3)
2.5
60
50
0
200
400
600
800
1000
Influent organic substrate (gCOD.m-3)
Figure 22. Relation among the maximum nitrogen removal (N removal) at given influent
organic substrate, corresponding ammonium removal (NH4 removal) and optimum oxygen
level (rp=0.75mm, NH4(in) =300 gN.m-3, T=30°C).
4.4. Effect of organic substrate on dynamics of nitrogen removal
Additional organic substrate gives advantages to heterotrophic bacteria to grow that increase
steady state nitrogen removal. From the dynamics of heterotrophic growth (Figure 5) it is
found that there are different types of heterotrophic growth and their fraction is changing over
time. The different types of heterotrophic growth over influent organic substrate effect on
dynamics of nitrogen removal. To analyze this dynamics behavior of nitrogen removal,
simulation is run for three influent organic substrate concentration (0, 50 and 100 gCOD.m-3)
with 0.75 mm granule size, 0.50 gO2.m-3 bulk oxygen concentration and 300 gN.m-3 influent
ammonium concentration. The results are summarized in Figure 23.
From Figure 23 it is found that organic substrate reduces the nitrogen removal at initial
period but higher nitrogen removal at steady state. An important point in this analysis is that
increasing organic substrate reduces the time to reach steady state. Increasing organic
50
Chapter IV: Results and discussion
substrate concentration from 0 to 100 gCOD.m-3 the steady state nitrogen removal increase
from 212 gN.m-3 to 292 gN.m-3 and time required to reach the steady state is reduced from
1140 days to 420 days. Adding too high organic materials again make an adverse effect on
removal efficiency (already explained in Figure 20). To get the maximum performance in
terms of nitrogen removal, there is an optimum organic substrate concentration based on bulk
oxygen level.
300
250
gN.m-3
200
150
SS=00
SS=50
SS=100
100
50
0
0
400
800
1200
1600
2000
Time (day)
Figure 23. Evaluation of the effect of bulk substrate organic materials on nitrogen removal
performance (rp=0.75 mm, O2=0.50gO2.m-3, NH4(in)=300g.m-3, T=30°C).
51
Chapter V: Conclusions and perspectives
In this work the effect of heterotrophic growth on biomass decay product and autotrophic
nitrogen removal in a granular sludge reactor is investigated. Besides this, the influence of
organic substrate present in the influent is studied.
A mathematical model describing the partial nitritation-anammox process was developed and
implemented in the AQUASIM simulation program. A simulation study has been conducted
to analyze the process performance with changing operational parameters, indicating the
crucial factors affecting the nitrogen removal process and indicating optimum process
parameters to achieve optimum performance in terms of maximum nitrogen removal.
1. Steady state and dynamic model behaviour
The steady state nitrogen removal performance is higher if heterotrophic growth on decay
products is taken up in the autotrophic nitrogen removal model. Heterotrophic bacteria take
part in the nitrogen removal process either by producing nitrogen gas from nitrite or by
producing nitrite from nitrate. Within artificial distinct three types of heterotrophic bacteria
according to consuming substrate, the nitrate consuming heterotrophic bacteria (XH,NO3) is
most desirable as it reduces nitrate to nitrite which can further be converted through
anammox bacteria, resulting in higher nitrogen removal. The conversion of nitrite to nitrogen
gas by heterotrophic bacteria (XH,NO2) reduces the autotrophic as well as total nitrogen
removal. At steady state, nitrite consuming heterotrophs are not present in the reactor and the
overall nitrogen removal increases with the presence of nitrate consuming heterotrophs.
The model performance is not sensitive to heterotrophic biomass density. Therefore the
changing of heterotrophic biomass density does not effect on nitrogen removal performance.
Due to growth of granules, growth of active biomass and changing the positions of active
biomass in the granules, a very long period of time is required to reach the steady state in
granular sludge reactor. Moreover the growth rate of anammox bacteria is very slow. The
time required for reaching steady state for soluble compounds is lower compared to time
required for biomass steady state. The steady state is reached earlier for high initial granule
52
Chapter V: Conclusions and perspectives
size with a high fraction of anammox bacteria. In reality, full performance of the process is
required within few days. For getting full efficiency within few days in a real treatment plant,
it is better to use cultured granules with a high fraction of anammox or granules with a high
anammox from existing treatment plants.
2. Influence of operational parameters and influent organic substrate
The bulk oxygen concentration is the main control variable for the partial nitritationanammox process. There is always an optimum bulk oxygen concentration with a peak point
of maximum nitrogen removal. For any deviation from this point, nitrogen removal
decreases. With decreasing oxygen concentration, ammonium accumulation increases but
with increasing oxygen level results in nitrate accumulation.
The performance of a granular sludge reactor is significantly affected through the granule
size. There is also a sharp peak point of nitrogen removal at an optimal granule size. Nitrogen
removal decreases when the granule size deviates from this point. But this optimum granule
size depends on bulk oxygen concentration. There is an optimum combination of granular
size with bulk oxygen concentration that leads to the maximum nitrogen removal. But the
removal efficiency is higher for the optimum combination of bulk oxygen concentration with
smaller granule size.
There is a range of temperatures with similar maximum nitrogen removal efficiencies in
partial nitritation-anammox process. A temperature below this range results in decreasing
nitrogen removal and increasing ammonium and nitrite concentrations, whereas a
temperature above this range results in reduced nitrogen removal and increased nitrate
accumulation.
Influents with high ammonium load needs high bulk oxygen concentration for maximum
nitrogen removal. But removal efficiency is higher at low ammonium surface load condition.
The nitrogen removal firstly increases and then decreases with influent organic substrate
concentration in partial nitritation-anammox process. The influent organic substrate gives an
advantage to heterotrophic bacteria and up to a certain level it increases the steady state
nitrogen removal. Presence of organic substrate reduces the time needed to reach steady state.
Without any influent organic substrate the maximum nitrogen removal is at a clear sharp
53
Chapter V: Conclusions and perspectives
point with an oxygen concentration but when the influent contains organic substrates, the
maximum removal is in a range of oxygen concentrations. Therefore influent organic carbon
makes the process easy to control at optimum level. The influent organic substrate that is
required for maximum nitrogen removal is related to bulk oxygen concentration. High
organic substrates need a high oxygen level for maximum performance, but higher organic
substrate concentration corresponds to lower nitrogen removal efficiency.
3. Future works
Overall, the results obtained in this work imply that there is an optimum bulk oxygen
concentration level for maximum nitrogen removal at a certain granule size, ammonium
surface load and influent organic substrate concentration. So it is important to maintain this
optimal bulk oxygen concentration level with combination of other process parameters for
partial nitritation-anammox process in a granular sludge reactor to get the best performance
with respect to nitrogen removal.
Since simulation studies have been completed, now it is time for the model calibration and
validation. A measurement campaign at a full scale granular sludge reactor for partial
nitritation-anammox is currently carried out. The obtained experimental results will be used
for calibration and validation of the model studied in this work.
Emission of nitrous oxide from wastewater treatment plants is attracting a lot of research
interest, since it is a potent greenhouse gas (Kampschreur et al., 2009). It is necessary to
include greenhouse gas emission in the model and optimize the process parameters towards
minimum greenhouse gas emission.
Several process parameters with heterotrophic growth in an autotrophic nitrogen removal
were investigated, but intermediate products (such as nitric oxide, nitrous oxide) and their
inhibition were not considered. Therefore further analysis of the process should be pursued in
order to improve the process performance in terms of nitrogen removal in a granular sludge
reactor by including intermediate products and their inhibition characteristics.
54
References
Abma W.R., Driessen W., Haarhuis R., van Loosdrecht, M.C.M. (2010). Upgrading of
sewage treatment plant by sustainable and cost-effective separate treatment of industrial
wastewater. Water Science and Technology, 61(7), 1715 - 1722.
Ahn Y.H. (2006). Sustainable nitrogen elimination biotechnologies: a review. Process
Biochemistry, 41(8), 1709 - 1721.
Alpkvist E., Picioreanu C., van Loosdrecht M.C.M., Heyden A. (2006). Three-dimensional
biofilm model with individual cells and continuum EPS matrix. Biotechnology and
Bioengineering, 94(5), 962 - 979.
Anthonisen A.C., Loehr R.C., Prakasam T.B.S., Srinath E.G. (1976). Inhibition of
nitrification by ammonia and nitrous acid. Journal of Water Pollution Control Federation,
48(5), 835 - 852.
Barnes D., Bliss P.J. (1983). Biological Control of Nitrogen in Wastewater Treatment, E. &
F.N. Spon, London, UK.
Baumann B., Snozzi M., van der Meer J.R., Zehnder A.J.B. (1996). Development of stable
denitrifying cultures during repeated aerobic-anaerobic transient periods. Water
Research, 31(8), 1947 - 1954
Beccari M., Dipinto A.C., Ramadori R., Tomei M.C. (1992). Effects of dissolved oxygen and
diffusion resistances on nitrification kinetics. Water Research, 26(8), 1099 - 1104.
Carrera J., Vicent T., Lafuente J. (2004). Effect of influent COD/N ratio on biological
nitrogen removal (BNR) from high-strength ammonium industrial wastewater. Process
Biochemistry, 39 (12), 2035 - 2041.
Cema G., Wiszniowski J., Zabczynski S., Zablocka-Godlewska E., Raszka A., SurmaczGórska
J.
(2007).
Biological
nitrogen
removal
from
landfill
leachate
by
deammonification assisted by heterotrophic denitrification in a rotating biological
contactor (RBC). Water Science and Technology, 55(8-9), 35 - 41.
Chamchoi N., Nitisoravut S., Schmidt J.E. (2008). Inactivation of ANAMMOX communities
under concurrent operation of anaerobic ammonium oxidation (ANAMMOX) and
denitrification. Bioresource Technology, 99(9), 3331 - 3336.
Chen F.Y., Liu Y.Q., Tay J.H., Ning P. (2011). Operational strategies for nitrogen removal in
granular sequencing batch reactor. Journal of Hazardous Materials, 189(1-2), 342 - 348.
55
References
Chiu Z.C., Chen M.Y., Lee D.J., Wang C.H., Lai J.Y. (2007). Oxygen diffusion in active
layer of aerobic granule with step change in surrounding oxygen levels. Water Research,
41(4), 884 - 892.
Dapena-Mora A., Fernandez I., Campos J.L., Mosquera-Corral A., Mendez R., Jetten M.S.M.
(2007). Evaluation of activity and inhibition effects on Anammox process by batch tests
based on the nitrogen gas production. Enzyme and Microbiol Technology, 40(4), 859 865.
de Kreuk M.K., Picioreanu C., Hosseini M., Xavier J.B., van Loosdrecht M.C.M. (2007).
Kinetic model of a granular sludge SBR: influences on nutrient removal. Biotechnology
and Bioengineering, 97(4), 801- 815.
de Kreuk M.K., Pronk M., Van Loosdrecht M.C.M. (2005). Formation of Aerobic Granules
and Conversion Processes in an Aerobic Granular Sludge Reactor at Moderate
Temperature. Water Research, 39(18), 4476 - 4484.
Dosta J., Fernández I., Vázquez-Padín J.R., Mosquera-Corral A., Campos J.L., Mata-Álvarez
J., Méndez R. (2008). Short- and long-term effects of temperature on the Anammox
process. Journal of Hazardous Materials, 154(1-3), 688 - 693.
Effler S.W., Brooks C.M., Auer M.T., Doerr S.M. (1990). Free ammonia and toxicity criteria
in a polluted urban lake. Research Journal of The water Pollution Control Federation,
62(6), 771 - 779.
Egli K., Fanger U., Alvarez P.J.J., Siegrist H., Van Der Meer J.R., Ehnder A.J.B. (2001).
Enrichment and characterization of an anammox bacterium from a rotating biological
contactor treating ammonium-rich leachate. Archive of Microbiology, 175(3), 198 - 207.
EPA. (1975). Process Design Manual for Nitrogen Control. US EPA Office Tech. Transfer,
Washington, DC.
Fux C., Boehler M., Huber P., Brunner I., Siegrist H. (2002). Biological treatment of
ammonium-rich wastewater by partial nitritation and subsequent anaerobic ammonium
oxidation (anammox) in a pilot plant. Journal of Biotechnology, 99(3), 295 - 306.
Grunditz C., Dalhammar G. (2001). Development of nitrification inhibition assays using pure
cultures of Nitrosomonas and Nitrobacter. Water Research, 35(2), 433 - 440.
Guven D., Dapena A., Kartal B., Schmid M.C., Maas B., van de Pas- Schoonen K. , Sözen S.,
Mendez R., Op den Camp H.J.M., Jetten M.S.M., Strous M., Schmidt I. (2005).
Propionate oxidation by and methanol inhibition of anaerobic ammonium-oxidizing
bacteria. Applied and Environmental Microbiology, 71(2), 1066 - 1071.
56
References
Hao X.D., Heijnen J.J., van Loosdrecht M.C.M. (2002). Sensitivity analysis of a biofilm
model describing a one-stage completely autotrophic nitrogen removal (CANON)
process. Biotechnology Bioengineering, 77(3), 266 - 277.
Hao X.D., van Loosdrecht, M.C.M. (2004). Model-based evaluation of COD influence on a
partial nitrification-Anammox biofilm (CANON) process. Water Science and
Technology, 49(11-12), 83 - 90.
Harremoës P., Sinkjaer O. (1995). Kinetic interpretation of nitrogen removal in pilot scale
experiments. Water Research, 29(3), 899 - 905.
Hawkins S., Robinson K., Layton A., Sayler G. (2010). Limited impact of free ammonia on
Nitrobacter spp. Inhibition assessed by chemical and molecular techniques. Bioresource
Technology, 101(12), 4513 - 4519.
Hellinga C., Schellen A., Mulder J.W., van Loosdrecht M.C.M., Heijnen J.J. (1998). The
SHARON process: an innovative method for nitrogen removal from ammonium-rich
wastewater. Water Science and Technology, 37(9), 135 - 142.
Hellinga C., van Loosdrecht M.C.M., Heijnen J.J. (1999). Model based design of a novel
process for nitrogen removal from concentrated flows. Mathematical and Computer
Modelling of Dynamic System, 5(4), 351 - 371.
Henze M., Gujer W., Mino T., van Loosdrecht M. (2000). Activated Sludge Models ASM1,
ASM2, ASM2d and ASM3. IWA scientific and technical report no 9. London: IWA
Publishing, 130.
Hidaka T., Yamada H., Kawamura M., Tsuno H. (2002). Effect of dissolved oxygen
conditions on nitrogen removal in continuously fed intermittent-aeration process with
two tanks. Water Science and Technology, 45(12), 181 - 188.
Hippen A., Rosenwinkel K.H., Baumgarten G., Seyfried C.F. (1997). Aerobic
deammonification: A new experience in the treatment of wastewaters. Water Science and
Technology, 35(10), 111 - 120.
Hsieh Y.L., Tseng S.K., Chang Y.J. (2003). Nitrogen removal from wastewater using a
double-biofilm reactor with a continuous-flow method. Bioresource Technology, 88(2),
107 - 113.
Hunik J.H., Tramper J., Wijffels R.H. (1994). A strategy to scale-up nitrification processes
with immobilized cells of Nitrosomonas europaea and Nitrobacter agilis. Bioprocess
Engineering, 11(2), 73 - 82.
57
References
Isaka K., Sumino T., Tsuneda S. (2007). High nitrogen removal performance at moderately
low temperature utilizing anaerobic ammonium oxidation reactions. Journal of
Bioscience and Bioengineering, 103(5), 486 - 490.
Jubany I., Lafuente J., Baeza J.A., Carrer J. (2009). Total and stable washout of nitrite
oxidizing bacteria from a nitrifying continuous activated sludge system using automatic
control based on Oxygen Uptake Rate measurements. Water Research, 43(11), 2761 2772.
Jung Y., Kang S.H., Chung Y.C., Ahn D.H. (2007). Factors affecting the activity of
anammox bacteria during start up in the continuous culture reactor. Water Science and
Technology, 55(1–2), 459 - 468.
Kampschreur M.J., Temmink H., Kleerebezem R., Jetten M.S.M., van Loosdrecht M.C.M.
(2009). Nitrous oxide emission during wastewater treatment. Water Research, 43(17),
4093 - 4103.
Kartal B., Kuenen J.G., van Loosdrecht M.C.M. (2010). Sewage Treatment with Anammox.
Science, 328(5979), 702 - 703.
Kindaichi T., Ito T., Okabe S. (2004). Ecophysiological interaction between nitrifying
bacteria and heterotrophic bacteria in autotrophic nitrifying biofilms as determined by
microautoradiography- fluorescence in situ hybridization. Applied and Environmental
Microbiology, 70(3), 1641 - 1650.
Koch G., Egli K., van der Meer J.R., Siegrist H. (2000). Mathematical modelling of
autotrophic denitrification in a nitrifying biofilm of a rotating biological contactor. Water
Science and Technology, 41(4-5), 191 - 198.
Komorowska-Kaufman M., Majcherek H., Klaczyn´ski E. (2006). Factors affecting the
biological nitrogen removal from wastewater. Process Biochemistry, 41(5), 1015 - 1021.
Kuai L., Verstraete W. (1998). Ammonium removal by the oxygen-limited autotrophic
nitrification-denitrification system. Applied and Environmental Microbiology, 64(11),
4500 - 4506.
Kuenen J.G. (2008). Anammox bacteria: from discovery to application. Nature Reviews
Microbiology, 6(4), 320 - 326.
Kumar M., Lin J.G. (2010). Co-existence of anammox and denitrification for simultaneous
nitrogen and carbon removal - strategies and issues. Journal of Hazardous Materials,
178(1-3), 1 - 9.
58
References
Lackner S., Terada A., Smets B.F. (2008). Heterotrophic activity compromises autotrophic
nitrogen removal in membrane aerated biofilms: Results of a modeling study. Water
Research, 42(4-5), 1102 - 1112.
Liu H., Yang F., Shi S., Liu X. (2010). Effect of substrate COD/N ratio on performance and
microbial community structure of a membrane aerated biofilm reactor. Journal of
Environmental Sciences, 22(4), 540 - 546.
Liu Y.Q., Liu Y., Tay J.H. (2005). Relationship between size and mass transfer resistance in
aerobic granules. Letters in Applied Microbiology, 40(5), 312 - 315.
Love N.G., Smith R.J., Gilmore K.R., Randall C.W. (1999). Oxime inhibition of nitrification
during treatment of an ammonia-containing industrial wastewater. Water Environment
Research, 71(4), 418 - 426.
Luo A., Zhu J., Ndegwa P.M. (2002). Removal of carbon, nitrogen, and phosphorus in pig
manure by continuous and intermittent aeration at low redox potentials. Biosystems
Engineering, 82(2), 209 - 215.
Manser R., Gujer W., Siegrist H. (2005). Consequences of mass transfer effects on the
kinetics of nitrifiers. Water Research, 39(19), 4633 - 4642.
Matsumoto S., Katoku M., Saeki G., Terada A., Aoi Y., Tsuneda S., Picioreanu C., van
Loosdrecht M. C. M. (2010). Microbial community structure in autotrophic nitrifying
granules characterized by experimental and simulation analyses. Environmental
Microbiology, 12(1), 192 - 206.
Metcalf and Eddy. (2003). Inc., Revised by Tchobananoglous G., Burton F.L. Wastewater
engineering: treatment, disposal and reuse. McGraw-Hill, McGraw-Hill series in water
resources and environmental engineering, New York, USA.
Moussa M.S., Hooijmans C.M., Lubberding H.J., Gijzen H.J., van Loosdrecht M.C.M.
(2005). Modelling nitrification, heterotrophic growth and predation in activated sludge.
Water Research, 39(20), 5080 - 5098.
Münch E.V., Lant P., Keller J. (1996). Simultaneous nitrification and denitrification in
bench-scale sequencing batch reactors. Water Research, 30(2), 277 - 284.
Mulder A., van de Graaf A.A., Robertson L.A., Kuenen J.G. (1995). Anaerobic ammonium
oxidation discovered in a denitrifying fluidized bed reactor. FEMS Microbiology
Ecology, 16(3), 177 - 184.
Muller A., Wentzel M.C., Loewenthal R.E., Ekama G.A. (2003). Heterotroph anoxic yield in
anoxic aerobic activated sludge systems treating municipal wastewater. Water Research,
37(10), 2435 - 2441.
59
References
Okabe S., Kindaichi T., Ito T. (2005). Fate of 14C-labeled microbial products derived from
nitrifying bacteria in autotrophic nitrifying biofilms. Applied and Environmental
Microbiology, 71(7), 3987 - 3994.
Peng Y., ChenY., Wang S., Peng C., Liu M., Song X. Cui Y. (2004). Nitrite accumulation by
aeration controlled in sequencing batch reactors treating domestic wastewater. Water
Science and Technology, 50(10), 35 - 43.
Philips S., Laanbroek H.J., Verstraete W. (2002). Origin, causes and effects of increased
nitrite concentrations in aquatic environments. Reviews in Environmental Science and
Biotechnology, 1(2), 115 - 141.
Picioreanu C., van Loosdrecht M.C.M., Heijnen J.J. (1997). Modelling the effect of oxygen
concentration on nitrite accumulation in a biofilm airlift suspension reactor. Water
Science and Technology, 36(1), 147 - 156.
Qiao S., Yamamoto T., Misaka M., Isaka K., Sumino T., Bhatti Z., Furukawa K. (2010).
High-rate nitrogen removal from livestock manure digester liquor by combined partial
nitritation-anammox process. Biodegradation, 21(1), 11 - 20.
Reichert P. (1994). Aquasim – A tool simulation and data-analysis of aquatic system. Water
Science and Technology, 30(2), 21 - 30.
Rittmann B.E., Laspidou C.S., Flax J., Stahl D.A., Urbain V., Harduin H., van der Waarde
J.J., Geurkkink B., Henssen M.J.C., Brouwer H., Klapwijk A., Wetterauw M. (1999).
Molecular and modeling analysis of the structure and function of nitrifying activated
sludge. Water Science and Technology, 39(1), 51 - 59.
Rittmann B.E., Stilwell D., Ohashi A. (2002). The transient-state, multiple-species biofilm
model for biofiltration processes. Water Research, 36(9), 2342 - 2356.
Sabumon P.C. (2007). Anaerobic ammonia removal in presence of organic matter: a novel
route. Journal of Hazardous Materials, 149(1), 49 - 59.
Satoh H., Ono H., Rulin B., Kamo J., Okabe S., Fukushi K. (2004). Macroscale and
microscale analyses of nitrification and denitrification in biofilms attached on membrane
aerated biofilm reactors. Water Research, 38(6), 1633 - 1641.
Semmens M.J., Dahm K., Shanahan J., Christianson A., (2003). COD and nitrogen removal
by biofilms growing on gas permeable membranes. Water Research, 37(18), 4343–4350.
Shammas N.K. (1986). Interactions of temperature, pH, and biomass on the nitrification
process. Journal Water Pollution Control federation, 58(1), 52 - 59.
Siegrist H. (1996). Nitrogen removal from digester supernatant—Comparison of chemical
and biological methods. Water Science and Technology, 34(1-2), 399 - 406.
60
References
Sinninghe-Damste J.S., Strous M., Rijpstra W.I., Hopmans E.C., Geenevasen J.A.J., Van
Duin A.C.T., Van Niftrik L.A., Jetten M.S.M. (2002). Linearly concatenated cyclobutane
lipids form a dense bacterial membrane. Nature, 419(6908), 708 - 712.
Strous M., Fuerst J.A., Kramer E.H.M., Logemann S., Muyze G., Van De Pas- Schoonen
K.T., Webb R., Kuenen J.G., Jetten M.S.M. (1999). Missing lithotroph identified as new
planctomycete. Nature, 400(6743), 446 - 449.
Strous M., Heijnen J.J., Kuenen J.G., Jetten M.S.M. (1998). The sequencing batch reactor as
a powerful tool for the study of slowly growing anaerobic ammonium-oxidizing
microorganisms. Applied Microbiology and Biotechnology, 50(5), 589 - 596.
Strous M., Kuenen J.G., Jetten M. (1999). Key physiology of anaerobic ammonium
oxidation. Applied and Environmental Microbiology, 65(7), 3248 - 3250.
Strous M., Pelletier E., Mangenot S., Rattei T., Lehner A., Taylor M.W., Horn M., Daims H.,
Bartol-Mavel D., Wincker P., Barbe V., Fonknechten N., Vallenet D., Segurens B.,
Schenowitz-Truong C., Medigue C., Collingro A., Snel B., Dutilh B.E., Op den Camp
H.J., van der Drift C., Cirpus I., van de Pas-Schoonen K.T., Harhangi H.R., van Niftrik
L., Schmid M., Keltjens J., van de Vossenberg J., Kartal B., Meier H., Frishman D.,
Huynen M.A., Mewes H.W., Weissenbach J., Jetten M.S.,Wagner M., Le Paslier D.
(2006). Deciphering the evolution and metabolism of an anammox bacterium from a
community genome. Nature, 440(7085), 790 - 794.
Suzuki I., Dular U., Kwok S.C. (1974). Ammonia or ammonium ion as substrate for
oxidation by Nitrosomonas europaea cells and extracts. Journal of Bacteriology, 120(1),
556 - 558.
Tang C., Zheng P., Wang C., Mahmood Q. (2010). Suppresion of anaerobic ammonium
oxidizers under high organic content in high-rate Anammox UASB reactor. Bioresource
Technology, 101(6), 1762 - 1768.
Tarre S., Beliavski M., Denekamp N., Gieseke A., de Beer D., Green M. (2004). High
nitrification rate at low pH in a fluidized bed reactor with chalk as the biofilm carrier.
Water Science and Technology, 49(11-12), 99 - 105.
Terada A., Hibiya K., Nagai J., Tsuneda S., Hirata A. (2003). Nitrogen removal
characteristics and biofilm analysis of a membrane-aerated biofilm reactor applicable to
high-strength
nitrogenous
wastewater
treatment.
Journal
of
Bioscience
and
Bioengineering, 95(2), 170 - 178.
61
References
Third K.A., Paxman J., Schmid M., Strous M., Jetten M.S.M., Cord-Ruwisch R. (2005).
Enrichment of Anammox from activated sludge and its application in the CANON
process. Microbial Ecology, 49(2), 236 - 244.
Udert K.M., Kind E., Teunissen M., Jenni S., Larsen T.A. (2008). Effect of heterotrophic
growth on nitritation/anammox in a single sequencing batch reactor. Water Science &
Technology, 58(2), 277 - 284.
van Benthum W.A.J., van Loosdrecht M.C.M., Tijhuis L., Heijnen J.J. (1995). Solids
retention time in heterotrophic and nitrifying biofilms in a biofilm airlift suspension
reactor. Water Science and Technology, 32(8), 53 - 60.
van de Graaf A.A., de Bruijn P., Robertson L.A., Jetten M.S.M., Kuenen J.G. (1996).
Autotrophic growth of anaerobic ammonium-oxidizing micro-organisms in a fluidized
bed reactor. Microbiology-UK, 142(8), 2187 - 2196.
van de Graaf A.A., de Bruijn P., Robertson L.A., Jetten M.S.M., Kuenen J.G. (1997).
Metabolic pathway of anaerobic ammonium oxidation on the basis of N-15 studies in a
fluidized bed reactor. Microbiology-UK, 143(7), 2415 - 2421.
van Hulle S.W.H., Vandeweyer H.J.P., Meesschaert B.D., Vanrolleghem P.A., Dejans P.,
Dumoulin A. (2010). Engineering aspects and practical application of autotrophic
nitrogen removal from nitrogen rich streams. Chemical Engineering Journal, 162(1), 1 20.
van Hulle S.W.H., Volcke E.I.P., López Teruel J., Donckels B., van Loosdrecht M.C.M.,
Vanrolleghem P.A. (2007). Influence of temperature and pH on the kinetics of the
SHARON nitritation process. Journal of Chemical Technology and Biotechnology,
82(5), 471 - 480.
van Loosdrecht M.C.M., Henze M. (1999). Maintenance, endogenous respiration, lysis,
decay and predation. Water Science and Technology, 39(1), 107 - 117.
van Loosdrecht M.C.M., Salem S. (2006). Biological treatment of sludge digester liquids.
Water Science and Technology, 53(12), 11 - 20.
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 biomass activity for one stage partial nitritation and anammox.
Applied and Environmental Microbiology, 76(3), 900 - 909.
Volcke E.I.P., Picioreanu C., De Baets B., van Loosdrecht M.C.M. (2010). Effect of granule
size on autotrophic nitrogen removal in a granular sludge reactor. Environmental
Technology, 31(11), 1271 - 1280.
62
References
Vrtovšek J., Roš M. (2006). Denitrification of Groundwater in the Biofilm Reactor with a
Specific Biomass Support Material. Acta Chimica Slovenica, 53(3), 396 - 400.
Wang J.L., Jing K. (2005). The characteristics of anaerobic ammonium oxidation
(ANAMMOX) by granular sludge from an EGSB reactor. Process Biochemistry, 40(5),
1973 - 1978.
Wett B. (2006). Solved upscaling problems for implementing deammonification of rejection
water. Water Science and Technology, 53(12), 121 - 128.
Wiesmann U. (1994). Biological Nitrogen Removal from Wastewater. Advances in
Biochemical Engineering Biotechnology, 51, 113 - 154.
Williamson K., McCarty P.L. (1976). Verification Studies of the Biofilm Model for Bacterial
Substrate Utilization. Journal of Water Pollution Control Federation, 48(2), 281 - 296.
Yamamoto T., Takaki K., Koyama T., Furukawa K. (2006). Novel partial nitritation
treatment for anaerobic digestion liquor of swine wastewater using swim-bed
technology. Journal of Bioscience and Bioengineering, 102(6), 497 - 503.
Yang S.F., Tay J.H., Liu Y. (2004). Respirometric activities of heterotrophic and nitrifying
populations in aerobic granules developed at different substrate N/COD ratios. Current
Microbiology, 49(1), 42 - 46.
Yoo H., Ahn K.H., Lee K.H., Kwak Y.U., Song K.G. (1999). Nitrogen removal from
synthetic wastewater by simultaneous nitrification and denitrification (SND) via nitrite in
an intermittently-aerated reactor. Water Research, 33(1), 145 - 154.
63