Anaerobic Ammonium Oxidation in Waste Water -An Isotope Hydrological Perspective

Full text

5
Anaerobic Ammonium Oxidation
in Waste Water -
An Isotope Hydrological Perspective
Yangping Xing and Ian D. Clark
Department of Earth Science, University of Ottawa
Canada
1. Introduction
Excess nitrogen components must be removed from wastewater to protect the quality of the
water bodies that it will be eventually discharged to. A conventional wastewater treatment
system for nitrogen removal is often involved with two processes, nitrification and
denitrification. Nitrification is mostly achieved by complete oxidation of ammonium (NH4+)
to nitrite (NO2-) by the appropriate aerobic bacteria and then oxidation of the nitrite to
nitrate ion (NO3-) by another variety of aerobic bacteria. Subsequently, the formed nitrate
will be reduced to dinitrogen gas under anoxic conditions at the expense of organic carbon
and released into the atmosphere as a harmless product (van Dongen et al., 2001). The
introduction of oxygen into wastewater for nitrification requires a large amount of energy.
Furthermore, the carbon source is often limited in wastewater, so purchasing of carbon
source (typically methanol) is necessary too. A newly discovered anaerobic ammonium
oxidation (anammox) may circumvent the limitations and open up a new possibility for
nitrogen removal from wastewater. The alternative approach is a microbiological involved
activity which requires less energy and enables more efficiency on N removal.
2. The history and physiology of anammox
The discovery of anammox activity and anammox bacteria is quite recent. Even though
Richards (1965) has noticed NH4+ deficits in anoxic marine basins, and proposed that the
missing NH4+ was anaerobically oxidized to N2 by some unknown microbe using nitrate as
an oxidant, which was coined one of two “lithotrophs missing in nature” by Broda (1977).
Because there was no known biological pathway for this transformation, biological
anaerobic ammonium oxidation received littler further attention (Arrigo, 2005). It was not
until mid-1990s, work with bioreactors designed to remove NH4+ from wastewater provided
direct evidence for anaerobic ammonium oxidation, and the process was termed
“anammox” by Mulder and his colleagues ( 1995). A series of 15N-labellling experiment
were carried out to study the metabolic mechanism and intermediates of anammox reaction
(van de Graaf et al., 1995; 1997). It is a chemolithotrophic process in which 1 mol of NH4+ is
oxidized by 1 mol of NO2- to produce N2 gas in the absence of oxygen (Strous et al., 1999).
NH4+ + NO2- → N2 +2H2O (1)

Waste Water - Treatment and Reutilization
90
The pathway of N2 formation clearly distinguishes anammox from denitrification which
combines N from two NO3- molecules to form N2 and presents as an elegant shortcut in the
natural nitrogen cycles (Fig 1.) Physical purification of the anammox microbes from the
multispecies biofilms yielded a 99.6% pure culture that was capable of carrying out PCR
amplification of the DNA. The microbes responsible for anammox process were identified as
members of the bacterial order Planctomycetales (Strous et al., 1999). The first genome
sequence of a representative anammox bacterium was published in 2006 (Strous et al., 2006).
To date, five anammox genera have been described, Candidatus Brocadia, Candidatus
Kuenenia, Candidatus Scalindua, Candidatus Anammoxoglobus and Candidatus Jettenia.
A range of studies have been conducted for the detection of anammox bacteria and activities in
variable environments from natural to man-made ecosystems (Risgaard-Petersen et al., 2003;
Schmid et al., 2005). Anammox activity was found in marine environments, such as the Black
Sea, the coast of Namibia, Chile, Peru and some freshwater and estuarine systems like, Lake
Tanganyika and mangroves (Kuypers et al. 2003; 2005; Risgaard-Petersen et al., 2004; Meyer et
al., 2005; Thamdrup et al., 2006; Schubert et al., 2006; Hamersley et al., 2009).In addition to
widespread distribution, the activity of anammox bacteria in the environments also be
substantial. The maximum reported contribution of anammox is 67-79%, occurring in
sediments at a depth of 700m of the Norwegian Trench (Engström et al., 2005). Considerable
supporting evidences have confirmed that anammox has global importance (Kuene, 2008).
Owing to the availability of laboratory enrichment cultures, the physiology of anammox
bacteria has been relatively well characterized (Jetten et al, 2005). Anammox is characterized
by slow growth and its cell doubles only once per 11 days under optimum conditions and 2-
3 weeks on average (Strous et al., 2006). The low growth rate of anammox bacteria is not
caused by inefficient energy conservation but by a low substrate-conversion rate.
Furthermore, anammox bacteria are obligate anaerobes and their metabolism is reversibly
inhibited when oxygen concentration is above 2 µM and nitrite is higher than 10 mM (Strous
et al., 1997a). The temperature range suitable for anammox bacteria has been reported
between -2℃ (sea ice, Rysgaard & Glud, 2004) and 43℃ (Strous et al., 1999). A recent study
has observed anammox activity at temperature from 60℃ to 85℃ at hydrothermal vents
located along Mid-Atlantic Ridge (Byrne et al., 2008). At optimal condition, anammox
biomass could be enriched from activated sludge within hundred days. Enriched anammox
bacteria in active sludge or biofilm present as brownish or red granule (Fig 2.). Under the
microscope, the bacteria are observed as small coccoid cells with diameter of approximately
800 nm. They all possess one anammoxosome, a membrance bound compartment inside the
cytoplasm which is the locus of anammox catabolism. Further, the intracytoplasmic is
surrounded by unique lipids, called ladderanes (Sinninghe Damsté et al., 2004). Due to their
unique characteristics, ladderane lipids have also been used as a biomarker for the presence
of anammox bacteria (Kuypers et al., 2003). Besides, an interesting special feature is the
turnover of hydrazine (normally used as a high-energy rocket fuel and poisonous to most
living organisms) as an intermediate.
In addition, anammox bacteria have been found to be metabolically flexible, exhibiting
alternative metabolic pathways. For instance, anammox can subsequently reduce NO3- to
NO2- to NH4+, followed by the conversion of NH4+ and NO2- to N2 through anammox
pathway, allowing anammox bacteria to overcome NH4+ limitation. Anammox bacteria are
also a potential source of N2O production by nitric oxide detoxification (Kartal et al., 2007).
Apart from NO2- and NO3-, anammox bacteria also employ Fe3+, manganese oxides as
electron acceptors (Strous et al., 2006), which further expended the metabolic diversity of the
anammox bacteria.

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
91
Fig. 1. Anammox in the context of nitrogen
cycle (Modified from Kuyper, et al., 2003).
Fig. 2. Typical anammox granular sludge
(Photo modified from Van Loosdrecht, 2006).
3. The application of anammox in waste water
Since anammox was discovered in a denitrifying fluidized bed reactor for wastewater
treatment, it was realized that having a great potential for the removal of undesired NH4+ from
wastewater from the beginning. The introduction of anammox process to N-removal would
lead to a 90% reduction in operation costs because by using anammox process, nitrification
process normally employed in wastewater treatment can be stopped at the nitrite level which
can save aeration and carbon sources. For this reason, Mulder and colleagues patented the
process immediately, even without direct proof and understanding of its biological nature
(Mulder, 1992). In recent years, many research efforts dedicated to the application aspects of
anammox reaction. The feasibility of the anammox process for the removal of NH4+ from
sludge digester effluents was evaluated. Experiments with a laboratory-scale (2L) fluidized
bed reactor showed that the anammox process was capable to remove NH4+ and NO2-
(externally added) efficiently from the sludge digester effluent. And anammox biomass could
be enriched from activated sludge within 100 days (Strous et al., 1997 b; Jetten et al., 1997). The
possible reactors are sequencing batch reactors (SBR), moving bed reactor, blanket reactor or
gas-lift-loop reactor. In these studies, NO2- was supplied from a concentrated stock solution.
However, for application in real wastewater practice, a suitable system for biological NO2- has
to be developed. One such system is the combination of the anammox process and SHARON
(Sustainable high rate ammonium removal over nitrite) process. The principle of the combined
process is that the NH4+ in the sludge digester effluent is oxidized in the SHARON reactor to
NO2- for only 50% in the reaction I. The mixture of NO2- and NH4+ is ideally suited as influent
for the anammox process in reaction II. With this system sludge digester effluent can be
treated independently. In the study, the SHARON process was operated stably for more than 2
years. During the test period the overall NH4+ removal efficiency was 83% (Van Dongen et al.,
2001). In the earlier design, reactions I and II were carried out in consecutive reactors, but these
were later combined in a single oxygen-limited reactor where nitrite-producing bacteria and
anammox bacteria coexist. However, anammox bacteria grow slowly and because of the low
specific conversion rates of one reactor process, the bottleneck in this combination has been
insufficient biomass retention (Kartal et al., 2010). A granular-sludge reactor is developed to
achieve a high volumetric conversion rate due to a large surface area for mass transfer (Kartal

Waste Water - Treatment and Reutilization
92
et al., 2010). The selective production of granules has been successfully applied on
nitrifying/anammox sludge in a sludge blanket reactor, which substantially improved the
energy management of wastewater facilities. Granular-sludge system not only overcome the
limit of conversion rate, but also offers the possibility for application of anammox for
wastewater treatment at low temperature and concentrations. The upper limits of nitrogen
loading to anammox process were explored in gas lift reactors. The results showed that
anammox bacteria were able to remove 8.9 kg N m-3 reactor day-1 (Jetten et al., 2004). Due to
extensive explorations of anammox process and combinations with other processes in the
practices of application, there are numerous developed systems from SHARON-anammox,
OLAND (Oxygen-limited autotrophic nitrification-denitrification, Kuai & Verstraete, 1998) to
CANON (Completely autotrophic nitrogen removal over nitrite, Third et al., 2001) and
DEAMOX (Denitrifying ammonium oxidation, Kalyuzhnyi et al., 2006). Van der Star et al.,
(2007) have made an overview and suggested that a uniform naming of these process as
shown in table 1.
Process name proposed by
van der Star et al., (2007)
Source of
nitrite
Alternative process
name Reference
Two reactor
Nitritation-anammox
process
Nitritation of
NH4+
SHARONa,b-
anammox
Two stage OLAND
Van Dongen et al.,
2001
Wyffels et al., 2004
One- reactor
Nitritation-anammox
Nitritation of
NH4+
OLANDc
CANONd
Aerobic/anoxic
deammonification
SNAPe
DEMONf
DIBf,g
Kuai and
Verstraete, 1998
Third et al., 2001
Hippen et al., 2001
Lieu et al., 2005
Wett, 2006
Ladiges et al., 2006
One reactor denitrification-
anammox process
NO3- of
denitrification
Anammoxh
DEAMOXi
Mulder et al., 1995
Kalyuzhnyi et al.,
2006
a Sustainable high rate ammonium removal over nitrate; the name only refers to nitritation when nitrite
oxidation is avoided by choice of residence time and operation at elevated temperature.
b Sometimes the nitrification-denitrification over nitrite is addressed by this term.
c Oxygen-limited autotrophic nitrification denitrification.
d Completely autotrophic nitrogen removal over nitrite.
e Single-stage nitrogen removal using the Anammox and partial nitritation.
f Name refers to the deammonification process in an SBR under pH-control.
g Deammonification in Interval-aerated Biofilm systems.
h System where Anammox was found originally. The whole process was originally designated as
Anammox.
i Denitrifying ammonium oxidation: this name only refers to denitrification with sulphide as electron
donor.
Table 1. Process names for nitrogen removal systems involving the anammox process
(modified from van der Star et al., 2007).

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
93
To date, there are several full-scale installations of anammox applications in the wastewater
treatment plants. The first full scale reactor was built in Netherlands in 2002. The prototype
has been set up as part of a municipal wastewater treatment plant in Rotterdam and is
performing well. The internal circulation type reactor used in Rotterdam is especially suited
for use of granular sludge. As of 2006, three full scale processes intended for the application
of anammox have been built in Europe. In addition, anammox bacteria have been found that
can be enriched from various types of wastewater sludge, indicating that anammox bacteria
are indigenous in many treatment plants throughout the world (Op den Camp et al., 2006).
Therefore, the ubiquitous characteristic of anammox bacteria makes no real limit to its
application at normal wastewater treatment plants.
4. Tracing anammox in contaminated ground water- a case study
Groundwater contamination by NH4+ typically occurs because of surface activities such as
composting, landfilling (Erksine, 2000), disposal of animal wastes and animal carcasses
(Ritter & Chirnside, 1995; Umezawa et al., 2008), fertilizer storage (Barcelona& Naymik,
1984), and septic system effluent (Aravena & Robertson, 1998). NH4+ contaminated
groundwater is a likely site for anammox activity. NH4+ enters the groundwater system and
competes for exchange sites on soil particle surfaces; then nitrifying organisms in the oxic
zone oxidize NH4+ to NO2- and then to NO3-. Movement of the groundwater through the soil
matrix carries the products of partial nitrification (NH4+ and NO2-/NO3-) as the plume
spreads due to the effects of retardation by aquifer material (Erksine, 2000). It is expected
that contaminated groundwater environments will favor the anammox reaction when both
NO2- and NH4+ are present in areas of low oxygen. In landfills, NH4+ is rarely detected over
a few hundred meters away from the source, suggesting that attenuation of NH4+ is
occurring along the flowpath (Erksine, 2000), and this is likely to be the case regardless of
the source of NH4+. We think that groundwater provides anammox organisms with an ideal
environment for growth. Isotope evidence for anammox in groundwater has been shown by
Clark and colleagues (Clark et al., 2008), but the presence and activity of anammox
organisms has yet to be confirmed. In the case study, a series of geochemical, isotopic,
labelling experiments and microbiological techniques including FISH, PCR, are used to
assess whether anammox organisms are present and active in NH4+-contaminated
groundwater sites.
4.1 Isotopic evidence of anammox
Tracing the fate of NH4+ and NO3- in ground water is greatly aided by measurement of 15N
and 18O, which can be used to characterize sources of these compounds and the reaction
pathways they may have followed (Delwiche & Steyn, 1970; Hübner, 1986; Kendall, 1998).
The reactions of nitrogen species in the environment are associated with characteristic
fractionations that provide additional insights to subsurface processes and fate.
Transformation of NO3- to N2 by denitrifying bacteria is accompanied by a 15N fractionation
on the order of ε15NN2_NO3 = -15‰ to -20‰ (Wada et al., 1975; Böttcher et al. 1990). Böttcher
et al. (1990) also showed that 18O is also enriched in the residual NO3- product, with
ε18ON2_NO3 = -8‰. Accordingly, stable isotopes provide important constraints on plausible
reaction pathways for nitrogen species in the subsurface. Within the context of tracing
anammox in ground water through the use of stable isotopes, a detailed investigation was
undertaken at the site of a municipal water supply aquifer contaminated by the activities of

Waste Water - Treatment and Reutilization
94
a chemical plant and fertilizer blending operation (Fig 3.). Wastewater contribution comes
from the chemical company and fertilizer blending company with ammonium approaching
840 ppm N and nitrate up to 350 ppm N.
4.1.1 Field and analytical work
A program of field sampling and analytical work was carried out in 2003 and again in 2004,
involving sampling ground water from 62 piezometers and extraction wells both on two
companies sites. Total NH4+ concentrations were analyzed on unfiltered samples by
distillation and titration with sulphuric acid. NO3- and NO2- concentrations were measured
by liquid chromatography. The 2004 series of samples were analyzed for isotopes of NH4+
(15N) and NO3- (15N and 18O) in the G.G. Hatch Isotope Laboratory in University of Ottawa.
15N-NH4+ was measured by a diffusion procedure. The sample was first distilled at high pH
into a sulphuric acid solution, and concentrated to ammonium sulphate salt by evaporation.
The salt precipitation was analyzed as N2 gas by continuous flow isotope ratio mass
spectrometry using a Finigan MAT Delta Plus directly interfaced with a Carlo Erba
elemental analyzer (EA). Isotopes in NO3- were analyzed by quantitative conversion of NO3-
to N2O gas according to the bacterial denitrifier method (Sigman et al. 2001; Casciotti et al.
2002). The bacterial N2O was analyzed for both 15N and 18O by injection through a gas bench
interfaced with a Finnigan MAT Delta Advantage continuous flow mass spectrometer. The
Fig. 3. Air photo of study area showing the
direction ground water flow from the waste
water ponds from chemical company and
fertilizer company to the confined
municipal aquifer.
Fig. 4. δ15NNH4 vs. total NH4+ for waste of
water source area and treatment well ground
water. Conservative mixing envelope shown
with black line.

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
95
N2 gas concentrations were measured in ground water sampled in septum vials by purging
with Helium (He) and direct injection into a Finnigan MAT Delta Advantage mass
spectrometer.
4.1.2 Isotope results and discussion
The regional background geochemistry of the confined municipal upper aquifer was
measured in two wells. The concentrations of NH4+ and NO3- of background water is lower
than 1 ppm N, and the redox conditions are considered to be moderately reducing, with
dissolved oxygen less than 1 mg L-1 and Eh of 137mV. Ground water at fertilizer company
source area was dominated by NH4-NO3 with an average NH4+ 632 ppm N and 250ppm
NO3- ppm N. The NH4+ concentration at chemical company source area was lower with an
average value of 40.6ppm N and this water has no detectable NO3-. However within
municipal upper aquifer, concentration of NH4+ was highly diluted with a maximum of 7.3
ppm N in the water treatment wells. The comparison between the really measured NH4+
and the predicted concentration of NH4+ by a conservative mixing model indicated that a
significant loss of NH4+ in ground water aquifer. The missing NH4+ was calculated to be 30.7
and 21.2 ppm N in treatment well 1 and treatment well 2, respectively. In the same way,
NO3- was found a loss of 8.0 and 3.2 ppm N from the two wells. These values are minimum
estimate because NO3- is not retarded by sorption in the aquifer like NH4+ (For more
information, please refer to the publication (Clark et al., 2008).
The missing of NH4+ was believed as a reactive loss involved with anammox reaction which is
based on isotopic evolution of associated nitrogen species. In conservative mixing, δ15N will
reflect the concentration-weighted contribution of NH4+ from each primary source. In the
present case, if no reaction, δ15N of nitrogen species would be weighted results of wastewater
from chemical company, fertilizer company with dilution water from background water.
Fractionation of 15N during cation exchange is considered to be minor to negligible (Kendall
1998), and so retardation is not expected to affect the δ15N of NH4+ in the municipal aquifer. By
contrast, reactive loss of NH4+ by oxidation, whether through nitrification or anammox, will
impart a clear enrichment trend independent of any mixing relationships. A plot of δ15NNH4
against NH4+ concentration shows a strong contrast between the two waste water source areas
and background NH4+ in the municipal aquifer (Fig 4.). The values for δ15NNH4 for the high
NH4+ concentration sites near the former fertilizer company water storage pond average 5.8‰,
while those for the chemical company average -2.7‰, providing an 8.5‰ contrast between
the two. The conservative mixing envelope, calculated from binary mixing between each of the
three endmembers, is shown in figure 4. Nonconservative behaviour of samples from the
fertilizer company source area is observed by their trend toward δ15N-enriched values at lower
NH4+ concentrations. Similarly, samples from the chemical company plume show
nonconservative enrichment in 15N. This is consistent with the conservative mixing
calculations, showing reactive loss of NH4+ along the flowpath. Because cation exchange has
been shown to be essentially nonfractionating (Ceazan et al., 1989; Kendall, 1998; Buss et al.,
2004), this reactive loss must be through oxidation.
The usual pathway for NH4+ oxidation is nitrification by O2. This is an energetically
favourable reaction in oxic water. It follows a two-step reaction through nitrite by a mixture
of aerobic bacteria, including Nitrosomonas, Nitrobacter nitrosospira, and Nitrobacter
pseudomonas. However, according to our measurement, redox conditions are unfavourable
for aerobic bacteria, and so NH4+ loss by nitrification is unlikely in these ground water.
Further evidence against nitrification is found by the positive correlation between NO3- and

Waste Water - Treatment and Reutilization
96
NH4+ in this water. NH4+ loss by oxidation to NO3- would show an inverse correlation and
NO3- would remain the dominant species in the municipal aquifer. A third line of evidence
against NH4+ nitrification is found in the comparison of δ15N values in NH4+ coexisting with
NO3- in individual water samples. Essentially all samples, NO3- were enriched in 15N over
coexisting NH4+. These rules out NH4+ nitrification as a source for NO3-, which would
produce NO3- with lower δ15N than the NH4+ precursor (Kendall, 1998). Furthermore, the
δ15NNH4 enrichment trends with decreasing NH4+ concentration against the possibility of
NH4+ nitrification (Fig 5.). The positive correlation for NH4+ and NO3-, the enrichment in
15NNH4, and the greater enrichment for 15N in NO3- over NH4+ suggest that the loss of NH4+
is due to anaerobic oxidation by anammox bacteria. Two Rayleigh distillation trend lines
trace the enrichment in δ15NNH4 in the residual NH4+ from different initial concentrations.
The enrichment factor ε15NNH4_N2 = 4‰ used for these trend lines provided the best fit for the
range of source area data points and thus provides a first-order estimate of 15N fractionation
during anammox reaction. Additional evidence for anammox reaction in the NH4+-NO3-
ground water at fertilizer company source area is found in the overpressing of N2 gas in
these samples. Normalization of measured N2 concentrations to atmospherically derived
Argon gas (Fig 6.) showed that overpressing in N2 in excess of three times of atmospheric
saturation. The δ18O composition of NO3- further supported reactive loss of NO3-, where
enrichment of δ18O and δ15N was seen for most samples (Data not shown).
Fig 5. Evolution of δ15NNH4 during
anammox reaction for the high NH4+-NO3-
fertilizer company ground water. Trend
lines calculated from a Rayleigh distillation
with a reaction enrichment factor of 4‰.
Fig. 6. Excess N2in fertilizer compan
y
g
round
water from reactive loss of NO3- and/ or NH4-,
normalized to dissolved argon gas.
4.1.3 Summary
Anaerobic oxidation of the ammonium by anammox bacteria is concluded as the reason of
the strong attenuation of NH4+ and NO3- observed between the source areas and the
municipal ground water treatment wells. Several lines of evidence suggest the conclusion:
1. δ15N measurements of NH4+ show progressive enrichment with decreasing
concentration, demonstrating reactive loss by ammonium oxidation. Volatilization

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
97
along the flowpath is unlikely because it requires unsaturated conditions and because
of the neutral pH of the water (negligible un-ionized NH3).
2. NO3- concentrations decline along the flowpath and into the municipal aquifer. This
precludes nitrification for the observed loss of NH4+ for which an increase in NO3-
concentrations should be observed. The measured redox conditions are too low to
support aerobic nitrification of NH4+.
3. δ15NNO3 is consistently 5‰ to 10‰ enriched over that of δ15NNH4 for water carrying both
species, demonstrating that NH4+ loss is not by nitrification. Oxidation of NH4+ to NO3-
would produce NO3- with depleted δ15N values.
4. Strong correlations between δ15NNH4 and δ15NNO3 demonstrate reactive loss of both
species, consistent with anammox reaction. Enrichment of δ15NNO3 correlates with
enrichments in δ15NNH4, further supporting reactive loss of NO3- .
5. N2 overpressuring above atmospheric equilibrium is observed to increase with
increasing δ15NNH4 values along the flowpath from the FC source area. Increased N2 in
conjunction with enrichment in δ15NNH4 can occur only through anaerobic oxidation of
NH4+ to N2 by the anammox reaction.
4.2 Tracer experiments
Tracer experiments with 15N-labeled nitrogen species are commonly used for elucidating
nitrogen fate in both sediments and groundwater environments. Consumption of 15NH4+
and concomitant production of 15N-labeled N2 provided the first clear experimental
evidence for anammox activity in a fluidized bed reactor (van de Graaf et al., 1995). So far,
few labelling experiments have provided evidence of anammox in anoxic basin and in the
suboxic zone of sea and lakes (Dalsgaard et al., 2003; Kuypers et al., 2003; Schubert et al.,
2006; Hamersley et al., 2009), but there is no analogue application in groundwater systems
yet. 15N-labelling also provides a very sensitive technique for the determination of anammox
rates. And a simultaneous determination of anammox and denitrification, gives in sights to
the relative importance of the two N removal pathways (Thamdrup & Dalsgaard, 2002;
Risgaard- Peterson et al., 2003). In addition, potential isotopic fractionation associated with
anammox bacteria activity also indicates the presence of anammox reaction. From the
simultaneous attenuation of NH4+ and NO3-, and a progress enrichment of δ15N-NH4+ and
δ15N-NO3-, Clark et al., (2008) suggested that anammox may play a role in ground water. As
a follow-up study, a series of 15N labelling incubation experiments have been established to
investigate anammox activity and reaction rates at several ground water sites.
4.2.1 15N labelling experiments
For 15N-labelling experiments, the method was slightly modified from the previous
publication (Dalsgaard et al., 2003). Ground water or sediment and groundwater in an
industrial contaminated site Elmira and a turkey manure polluted site Zorra were collected
directly to 12-mL exetainers (Labco, UK). In terms of the mixture of sediment and ground
water incubation, around 4.5mL sediment and 7.5mL of groundwater were collected. In
order to minimize oxygenation, exetainer was submerged into a big container completely
filled with ground water and neither headspace nor bubbles in the vial. From each site,
triplicates were sampled for 15N labelling experiments. 15N labelling experiments were
conducted immediately after return to the laboratory (less than 2 hours). In brief, 3mL of
water was withdrawn by a syringe to make a headspace for helium (He) flushing. Each

Waste Water - Treatment and Reutilization
98
exetainer was flushed with He for at least 15min to remove background N2 and dissolved O2
and N2. 15N enriched compounds were added with syringe to a final concentration of
100µmol in 10ml of sample as 15NH4Cl and Na15NO3 (all >99% 15N, Sigma-Aldrich). Even
though the final concentration of enriched 15N was variable in previous studies, ranging
from 40 µmol to 10mmol L-1 (Dalsgaard et al, 2003; Thamdrup et al., 2006), the present
addition was in higher range because that the concentration of 14N species in study samples
were very high and sometime can reach to 20mmol L-1. An additional trial was carried out
without any tracer addition as control to confirm that the whole incubation system functions
well. 15N-labelling experiments were incubated in a dark incubation chamber at 15°C,
which is very close to the in situ temperature. 14N15N:14N14N and 15N15N: 14N14N were
determined by gas chromatography-isotope ratio mass spectrometry and expressed as
δ14N15N values ( 14 15
14 15 14 14
14 15 14 14
()sample
NN [ 1]1000
()standard
NN:NN
NN:NN
δ
=−×
; air was used as the standard)
(GG Hatch isotope laboratory, University of Ottawa). In terms of anammox contribution to
total N2 production, assuming that the 15NH4+ pool turns over at the same rate as the
ambient 14NH4+ pool, the total anammox N2 production can be calculated from the
production of 29N2 and the proportionate 15N labelling in the whole NH4+ pool (Thamdrup
& Dalsgaard, 2002; Thamdrup et al., 2006). The rates of anammox were extrapolated from
linear regression of 14N15N as a function of time in the incubation with 15NH4+ and the rates
of denitrification were determined from the slope of linear regression of 15N15N over time in
the incubation with 15NO3-.
4.2.2 Results and discussion
At both of sampling sites except a pristine background well (Pu86 having not been impacted
by NH4+ from the compost plume), the formation of 14N15N was observed in the incubation
trials with 15NH4+ (Fig 7 a and c). However, the formation of 14N15N was very slow, and the
concentration was lower than the detection limit after 72 hours incubation and the
enrichment signal δ15N/14N was only 22.1 ± 4.2‰. The incubation experiments were
extended to 3 months. The highest δ15N/14N increased to 14,278.03‰ at the end of
incubation. At Elmira site, 14N15N accumulated linearly and stably with time without a lag
phase, which indicates that anammox was the active process and no intermediates were
involved in the reaction (Galán et al., 2009). Furthermore, the production of only 14N15N
rather 15N15N was a clear evidence for the stoichiometry of N2 production through
anammox (van de Graaf et al., 1995; Jetten et al., 2001). At Zorra site, the formation of 14N15N
reached the maximum at 1500hours incubation and started to decline. This is maybe due to
the lack of another N donor NO3- which concentration was low at Zorra site. In control
incubations without added tracer there was no production of 15N-enriched N2, indicating the
eligibility of the incubation system. At Elmira sites, the average 14N15N formation rate was
0.014±0.003µmol L-1 h-1, and the rate at Zorra site was 0.02±0.0021 µmol L-1 h-1. The rate of
14N15N production essentially corresponded to the anammox rate (van de Graaf et al., 1995;
Thamdrup & Dalsgaard 2002; Dalsgaard et al., 2003). So, according to the equation from
Thamdrup & Dalsgaard (2002), the calculated anammox reaction was 0.04±0.008 µmol L-1 h-1
at Elmira and 0.021±0.0022 µmol L-1 h
-1 at Zorra. Compared to Dalsgaard et al., (2003)
reported reaction rates 42 to 61mmol N m-2 d-1 in anoxic water column of Golfo Dulce, the
reaction rate in ground water was much lower. However, many lower rates have been
found in the oxygen-deficient water such as in eastern South Pacific (≤0.7nmol L-1 h-1;

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
99
Thamdrup et al., 2006) and in the Black Sea (~0.007µmol d-1; Kuypers et al., 2003). Our
results were very close the reported reaction rates in freshwater lakes, ranging from 6 to 504
nmol N2 L-1 d-1 (Hamersley et al., 2009).
The pronounced accumulation of 15N15N in the incubation of 15NO3- indicated that active
and strong denitrification process (Fig 7b and d). The production of 15N15N was the major
product at Zorra sites with an order magnitude higher than the mass of 14N15N. In the
incubation of 15NH4+, using the calculated anammox produced N2 as a numerator and the
total produced N2 (14N14N+14N15N+ insignificant 15N15N) as a denominator, at Elmira sites
32.7% of N2 gas was attributed to anammox; 21.4% for Zorra sites. 15NO3- tracer labelling
experiment showed that anammox accounted for 44.79% of N2 production at Elmira sites
and 29.03% at Zorra sites. The two techniques demonstrated a fair agreement at both of
study sites. To date, the reported relative contribution of anammox to N2 production was
variable with a wild range from below detection to 67% (Thamdrup & Dalsgaard 2002;
Dalsgaard et al., 2005). The contribution of anammox activity to N cycle was fairly
corresponding to the percentage of anammox bacteria biomass (bacteria biomass data will
be shown following). In conclusion, 15N labelling experiments directly and clearly proved
that the presence and activity of anammox in ground water.
Fig. 7. Formation of 14N15N (open square) and 15N15N (solid square) in 3mL of headspace of
incubation vials with samples from Elmira site(a and b) and Zorra site(c and d) after
addition of 15NH4+ and 15NO3-.

Waste Water - Treatment and Reutilization
100
4.3 Microbiological analyses
Molecular methods have been extensively utilized to identify the presence of anammox
bacteria in environmental and wastewater samples. Fluorescence in situ hybridization
(FISH) targeting the 16S rRNA gene has been used extensively, and described in detail by
Schmid et al. (2005). Anammox bacteria have also been identified using PCR, using a variety
of primers, often based on FISH probes, targeting the group as a whole or specific members
(Schmid et al. 2005; Penton & Tiedje, 2006). Quantitative PCR (q-PCR) has been used for
direct quantification of all known anammox-like bacteria in water columns (Hamersley et al.
2009), in wastewater enrichment cultures (Tsushima et al., 2007) and in terrestrial
ecosystems (Humbert et al., 2010).
4.3.1 Microbiological methods
For the present study, between 240 mL and 1 L of groundwater was collected and filtered via
piezometer for DNA extraction; filtrate was collected on a 0.22μm filter surface (Millipore).
Filters were stored at –70oC until DNA extraction. Nucleic acids were extracted from the filter
surface using a phenol chloroform extraction technique, described previously by Neufeld et
al., (2007). General bacterial 16S rRNA gene primers for denaturing gradient gel
electrophoresis (DGGE; GC-341f and 518r; Muyzer et al., 1993) and anammox-specific 16S
rRNA gene primers (An7f and An1388r; Penton et al., 2006) were used for PCR along with a
series of reaction conditions (Moore et al, submitted). PCR products were cloned using a
TOPO-TA cloning kit (Invitrogen) according to the manufacturer’s instructions. DNA
sequencing was performed at the Biochemistry DNA sequencing facility at the University of
Washington (ABI 3700 sequencer), at The Center for Applied Genomics in Toronto (ABI
3730XL sequencer), and at the sequencing facility at the University of Waterloo (Applied
Biosystems 3130xl Genetic Analyzer). DNA chromatograms were manually edited for base
mis-calls and were visually inspected and trimmed to ensure only quality reads were
included. Redundant sequences were removed using Jalview. Alignment and building
phylogenetic trees were done with MEGA4.0 (Tamura et al., 2007). Sequences were aligned
with known anammox reference sequences obtained from Genbank (DQ459989, AM285341,
AF375994, DQ317601, DQ301513, AF375995, AF254882, AY257181, and AY254883) and a
Planctomycete outgroup (EU703486). Phylogenetic trees were built using the neighbor joining
method and the maximum composite likelihood model. Total bacterial community pie charts
were constructed using phylum assignments provided by the Ribosomal Database Project and
NCBI Blast. Anammox specific qPCR used An7f and An1388r (Penton et al., 2006) and general
bacterial qPCR used 341f and 518r (Muyzer et al., 1993).
Fluorescently labelled oligonucleotide probes: EUB 338 (specific for all bacteria cells),
Amx368 (specific for all anammox species) and Kst- 0157-a-A-18 (specific for an anammox
species “Kuenenia Stuttgartiensis”) all labelled with different fluorescent color were used to
ground water and sediment samples in order to determine the abundance of the specific
anammox bacteria cells in samples. Several protocols have been used and a suitable protocol
for this type of environmental samples was modified. In order to give a quantitative point
view of total cell versus anammox, cell counting was established. Total cell counting was
carried by DAPI (4',6-diamidino-2-phenylindole) staining, which is a special fluorescent
stain that binds strongly to the DNA’s of only all bacterial cells (Tekin, in preparation).
4.3.2 Results and discussion
Planctomycete abundance in the total bacterial community increased with depth at Zorra
according to clone library data, and planctomycetes reached 5.2 and 20.8% of the total

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
101
bacterial community at depths greater than 5 m below ground surface. Large Illumina
libraries (~100 000) sequences indicated that anammox organisms made up ~10% of the
bacterial community at Zorra. Quantitative PCR using anammox specific primers (An7f
An1388r; Penton et al. 2006) confirmed that the abundance of anammox organisms increased
with the observed increase in planctomycete abundance at Zorra site. The number of
anammox 16S rRNA gene copies at Elmira was lower on average than that of Zorra. A
pristine background well (having not been impacted by NH4+ from the compost plume)
showed two orders of magnitude fewer anammox gene copies per nanogram of genomic
DNA than at impacted area. Clone libraries targeting the 16S rRNA genes of anammox
bacteria were used to examine the communities of anammox performing organisms at field
sites. All Anammox organisms were present at the two contaminated groundwater sites
however the community compositions differ (Fig 8). At Zorra site, Can. Brocadia dominated
anammox community, where the vast majority of anammox sequence also grouped with
known Can. Brocadia reference sequence, and a few clones grouped with known Can.
Scalinadua. FISH images also showed the presence of anammox bacteria in both of two
ground water sites (Data not shown).
Fig. 8. (a) Phylogenetic tree of environmental anammox sequences aligned with known
anammox reference sequences. Numbers in brackets represent the number of clones
identifying with each cluster. (b) Distribution of anammox related 16S rRNA gene
sequences found at each field site, by genus. (Modified from Moore et al., in preparation).
Anammox organisms are very hard to culture due to extremely slow growth rates, so there
is a high reliance on molecular techniques for finding and identifying these organisms in
mixed communities. PCR of environmental DNA extracts with general bacterial primers to
generate clone libraries has been shown to underestimate the proportion of anammox
organisms in the environment due to mismatches with “universal” primers (Jetten et al.,
2009; Penton et al., 2006; Schmid et al., 2007). Anammox organism abundance may be
greater than estimated by molecular methods due to known mismatches of anammox
organisms with several “anammox,” “planctomycete” or “universal bacterial” primer sets.
Anammox organisms have at least 10 mismatches with 27f and 2 mismatches with 1492r,

Waste Water - Treatment and Reutilization
102
primers used to create general bacterial 16S rRNA gene libraries for Zorra where the
abundance of planctomycetes was estimated to be between 5.2 and 20.8% of the total
bacterial population at 7.5 m. In summary, the results of microbiological investigation
provided further evidence for anammox presence in ground water and additional insight of
anammox bacteria community in ground water environments.
5. Anammox and denitrification in waste water
From a geochemical perspective, anammox and denitrification have the same implication,
i.e., they both lead to a loss of fixed nitrogen, albeit with a somewhat different
stoichiometry. The biogeochemical relationship between anammox bacteria and denitrifies
appears quite complex. They always coexist in the same environment where they can be
competitor to each other and also can play as a booster too.
In some environments with low NH4+, anammox depends on ammonification, which may
connect with denitrifies’ function on N-containing organics. In addition, the electron
acceptor of anammox NO2- also highly relies on the production of denitrification. Therefore,
the combination of anammox and denitrification is introduced in most of application in
waste water treatment as above stated. Under the assumption that NO2- consumption by
anammox can be described by Michaelis-Menten kinetics (Dalsgaard et al., 2003), the
apparent half-saturation concentrations, Km for NO2- during anammox in natural
environments has been constrained to <3 µM (Trimmer et al., 2003). Since maximum NO2-
concentrations in natural environments are only few µmol per liter, tighter competition for
NO2- may affect the balance between anammox and denitrification (Kuyper et al., 2006). The
competition ability relies on the availability of organic matter and the physiology of
bacteria. Anammox bacteria is regarded as autotrophic, so the activity of anammox bacteria
may not be directly associated with organic matter. In contrast, organic matter provides
both of energy and substrates to denitrification which sometime limits denitrification
activity, especially in waste water treatment (Ruscalleda et al., 2008), but denitrifies grow
faster than anammox bacteria which make the organisms easily outgrown in the
competition. Similarly, NH4+ sometime derives from ammonification as mentioned above
which more complicate the relationship of the two processes.
With more studies, more and more scientists argue that it is possible that anammox account
for a substantial 30-50% of N2 production in the ocean or oxygen minimum zone.
Theoretically, 29% of N2 production during the complete mineralization of Redfieldian
organic matter through denitrification and anammox, is produced through anammox
(Dalsgaard et al., 2003; Devol, 2003). Kuyper et al., (2006) supposed the number can exceed
48%. However, Gruber (2008) think this conclusion can not be easily extrapolated, since the
dependence of anammox on denitrification, but he also pointed out that there is ample room
for surprises since how little we know about the process and the associated organisms.
6. Conclusions and outlook
Over 40 years have passed since the anaerobic oxidation of ammonium with nitrite
reduction was first proposed. However, our understanding of anammox is till far from
complete. Anammox research is still in a very early state. All over the world, research
groups are working on diverse aspects of the molecular biology, biochemistry,
ultrastructure, physiology and metabolism and ecology of anammox process. As well as

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
103
assessing the impact of the activity on the environment and their application in waste water
treatment. A lot of interesting facts have been revealed and certainly more will come in
future. Identifying the genomes of anammox bacteria will help to cultivate these bacteria in
pure cultures what wasn’t achieved until now. Pure cultures could optimize the application
of anammox in wastewater treatment plants and facilitate the research on the anammox
bacteria. Several important questions remain to be answered are: how important the
anammox process is in freshwater ecosystems, especially contaminated aquifer? How do
anammox organisms interact with other nitrogen involved bacteria? From an isotope
hydrological perspective, the relevant fractionation factors have yet to be established. Also,
the limited applications on waste water treatment indicate that a further understanding of
anammox is needed.
7. Acknowledgements
We are grateful for the significant contributions from J. Neufeld, T. Moore, E, Tekin, D.
Fortin and to G.G Hatch isotope laboratory and geochemistry laboratories at University of
Ottawa and University of Waterloo. This work was supported by NSERC awarded to Dr. I.
Clark.
8. References
Abma, W.R.; Driessen, W. Haarhuis, R. & van Loosdrecht, M.C.M. (2010). Upgrading of
swage treatment plant by sustainable and cost-effective separate treatment of
industrial wastewater. Water Sci. Technol. 61., 1715-1722.
Aravena, R. & Robertson, W.D. (1998). Use of multiple isotope tracers to evaluate
denitrification in ground water: Study of nitrate from a large-flux septic system
plume. Ground Water. 36., 975-982.
Arrigo, K.R. (2005). Marine microorganisms and global nutrient cycles. Nature, 437., 15., 349-
355.
Barcelona, M.J. & Naymik, T.G. (1984). Dynamics of a fertilizer contaminant plume in
groundwater. Environ. Sci. Technol. 18., 4.,257-261.
Böttcher, J.; Strebel, O. Voerkelius, S. & Schmidt, H.-L. (1990). Using isotope fractionation of
nitrate nitrogen and nitrate oxygen for evaluation of denitrification in a sandy
aquifer. Journal of Hydrology. 114, 413–424.
Broda, E. (1977). Two kinds of lithotrophs missing in nature. Z. Allg. Mikrobiogie. 17., 491-
493.
Buss, S.R.; Herbert A.W., Morgan, P. Thornton, S.F. & Smith, J.W.N. (2004). A review of
ammonium attenuation in soil and groundwater. Quarterly Journal of Engineering
Geology and Hydrogeology 37, 347–359.
Byrne, N.; Strous, M. Crépeau, V. Kartal, B. Birrien, J.L. Schmid, M. Lesongeur, F. Schouten,
S. Jaeschke, A. Jetten, M.S.M. Prieur, D. & Godfroy, A. (2008). Presence and activity
of anaerobic ammonium- oxidizing bacteria at deep-sea hydrothermal vents. ISME
Journal. 3., 117-123.
Casciotti, K.L.; Sigman, D.M. Galanter Hastings, M. Böhlke, J.K. & Hilkert, A. (2002).
Measurement of the oxygen isotopic composition of nitrate in marine and fresh
waters using the denitrifier method. Analytical Chemistry. 74, 4905–4912.

Waste Water - Treatment and Reutilization
104
Ceazan, M. L.; Thurman, E.M. & Smith, R. L. (1989). Retardation of ammonium and
potassium transport through a contaminated sand and gravel aquifer. The role of
cation exchange. Environmental Science & Technology. 23., 1402-1408.
Clark, I. ; Timlin, R. Bourbonnais, A. Jones, K. Lafleur, D. & Wickens, K. (2008). Origin and
fate of industrial ammonium in anoxic ground water 15N evidence for anaerobic
oxidation (anammox). Ground Water Monit. Remediat. 28., 3., 73-82.
Dalsgaard, T.; Canfield D, E. Petersen, J. Thamdrup, B. & Acuña-González, J. (2003).
Anammox is a significant pathway of N2 production by the anammox reaction in
the anoxic water column of Golfo Dulce, Costa Rica. Nature. 422., 606-08.
Dalsgaard, T.; Thamdrup, B. & Canfield, D.E. (2005) Anaerobic ammonium oxidation
(anammox) in the marine environment. Res Microbiol. 156: 457–464.
Delwiche, C.C. & Steyn, P.L. (1970). Nitrogen isotope fractionation in soils and microbial
reactions. Environmental Science & Technology. 4., 45–67.
Devol, A. H. (2003). Solution to a marine mystery. Nature, 422., 575-576.
Engström, P.; Dalsgaard, T. Hulth, S. & Aller, R.C. (2005). Anaerobic ammonium oxidation
by nitrite (anammox): Implications for N2 production in coastal marine sediments.
Geochim Cosmochim Acta. 69., 2057–2065.
Erksine, A.D. (2000). Transport of ammonium in aquifers: retardation and degradation.
Quart. J.Engin. Geol.Hydrogeol. 33., 161-170.
Galán, A.; Molina, V. Thamdrup, B. Woebken, D. Lavik, G. Kuypers, M.M.M. & Ulloa, O.
(2009). Anammox bacteria and the anaerobic oxidation of ammonium in the oxygen
minimum zone off northern Chile. Deep-Sea Research (II). 56., 1021-1031.
Gruber, N. (2008). The marine nitrogen cycle: overview and challenges. In: Nitrogen in the
marine environment, 2nd edition. Capone, D.G. (Ed.). Elsevier Publisher, London, UK.
Hamersley, M. R.; Moebken, D. Boeherer, B. Schultze, M. Lavik, G. & Kuypers, M. M.M.
(2009). Water column anammox and denitrification in a temperate permanently
stratified lake (Lake Rassnitzer, Germany). Systematic and Applied Microbiology. 32.,
571-582.
Hippen, A.; Rosenwinkel, K.-H. Baumgarten, G. & Seyfried, C.F. (1997). Aerobic de-
ammonification: a new experience in the treatment of wastewaters. Water Sci.
Technol. 35., 111–120.
Humbert, S.; Tarnawski, S. Fromin, N. Mallet, M-P. Aragno, M. & Zopfi, J. (2010). Molecular
detection of anammox bacteria in terrestrial ecosystems: distribution and diversity.
The ISME Journal 4, 450–454.
Hübner, H. (1986). Isotope effects of nitrogen in the soil and biosphere. In: Handbook of
Environmental Isotope Geochemistry, Vol. 2, The Terrestrial Environment. B, Fritz, P. &
Ch- Fontes, J.( Ed.), 361–425. Elsevier, Amsterdam, Netherlands.
Jetten, M.S.M.; Cirpus, I. Kartal, B. van Niftrik, L. van de Pas-Schoonen, K.T. Sliekers, O.
Haaijer, S. van der Star, W. Schmid, M. van de Vossenberg, J. Schmidt, I. Harhangi,
H. van Loosdrecht, M. Kuenen, J.G. Op den Camp, H.& Strous, M. (2005). 1994-
2004: 10 years of research on anaerobic oxidation of ammonium. Biochemical Society
Transaction. 33., 1., 119-123.
Jetten, M.S.M.; Horn, S.J. & van Loosdrecht, M.C.M. (1997) Towards a more sustainable
municipal wastewater treatment system. Water Sci. Technol. 35., 171-180.

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
105
Jetten, M.; Wagner, M. Fuerst, J. van Loosdrecht, M. Kuenen, J.G. & Strous, M. (2001).
Microbiology and application of the anaerobic ammonium oxidation (anammox)
process. Current Opinion in Biotechnology. 12., 283–288.
Jetten, M.S.M.; van Niftrik, L. Strous, M. Kartal, B. Keltjens, J.T. & Op den Camp, H.J.M.
(2009). Biochemistry and molecular biology of anammox bacteria. Crit. Rev.
Biochem. Mol. Biol. 44(2-3)., 65-84.
Kartal, B.; Kuenen, J.G. & van Loosdrecht, M.C.M. (2010). Sewage treatment with anammox.
Science. 328., 702-703.
Kartal, B.; Kuypers, M.M.M. Lavik, G. Schalk, J. Op den Camp, H.J.M. Jetten, M.S.M. &
Strous, M. (2007). Anammox bacteria disguised as denitrifiers: nitrate reduction to
dinitrogen gas via nitrite and ammonium. Environ Microbiol. 9., 635–642.
Kayuzhnyi, S.; Gladchenko, M. Mulder, A. & Versprille, B. (2006). DEAMOX-New biological
nitrogen removal process based on anaerobic ammonia oxidation coupled to
sulphide-driven conversion of nitrate into nitrite. Wat. Res. 40., 3637-3645.
Kendall, C. (1998). Tracing nitrogen sources and cycling in catchments, In: Isotope Tracers in
Catchment Hydrology, Kendall C. & McDonnell, J.J. (Ed.), 526–531. Elsevier,
Amsterdam, Netherlands.
Kuai, L.P. & Verstraete, W. (1998). Ammonium removal by the oxygen-limited antotrophic
nitrification-denitrification system. Appl. Environ. Microbiol. 64., 4500-4506.
Kuenen, J. G. (2008). Anammox bacteria: from discovery to application. Nature. 6., 320-326.
Kuypers, M.M.M.; Sliekers, A.O. Lavik, G. Schmid, M. Jørgensen, B.B. Kuenen, J.G.
Sinninghe Damsté, J.S. Strous, M. & Jetten, M.S.M. (2003). Anaerobic ammonium
oxidation by anammox bacteria in the Black Sea, Nature. 422., 608-11.
Kuypers, M.M.M.; Lavik, G. Woebken, D. Schmid, M. Fuchs, B.M. Amann, R. Barker
Jøregensen, B. & Jetten, M.S.M. (2005). Massive nitrogen loss from the Benguela
upwelling system through anaerobic ammonium oxidation. Proc. Natl Acad. Sci.
USA.102., 6478–6483.
Kuyper, M.M.M.; Lavik, G. & Thamdrup, B. (2006). Anaerobic ammonium oxidation in
marine environment, In: Past and present water column anoxia. Neretin, L.N. (Ed.),
NATO Science series. Springer. Dordrecth, The Netherlands.
Ladiges, G.; Thierbach, Beier R.D., & Focken, M. (2006). Versuche zur zweistufigen
Deammonifikation im Hamburger Kla¨rwerksverbund. [Attempts to two-stage
deammonification in the wastewatertreatment union of Hamburg]. 6. Aachener
Tagung mit Informationsforum: Stickstoffru¨ ckbelastung -Stand der Technik 2006-,
Aachen (Ger), ATEMIS GmbH.p. Fachbeitrag 13 (13p).
Lieu, P.K.; Hatozaki, R. Homan, H. & Furukawa, K. (2005). Singlestage nitrogen removal
using Anammox and partial nitritation (SNAP) for treatment of synthetic landfill
leachate. Jpn. J. Water Treat. Biol. 41 (2)., 103.
Meyer, R.L.; Risgaard-Petersen, N. & Allen, D.E. (2005). Correlation between anammox
activity and microscale distribution of nitrite in a subtropical mangrove sediment.
Appl. Environ. Microbiol. 71., 10., 6142-6149.
Moore, T.; Xing, Y.P. Tekin, E. Lazenby, B. Schiff, S. Robertson, W. Timlin, R. Lanza, S. M.
Ryan, C. Aravena, R. Fortin, D. Clark, I. & Neufeld, J.D. Characterization of
groundwater-associated communities of anaerobic ammonium-oxidizing bacteria.
(Submittedd to Applied and Environmental Microbiology).
Mulder, A. (1992). Anoxic ammonia oxidation. Patent number: 5078884. USA.

Waste Water - Treatment and Reutilization
106
Mulder, A.; van de Graff, A. A. Robertson, L.A. & Kuenen, J. G. (1995). Anaerobic
ammonium oxidation discovered in a denitrifying fluidized bed reactor. FEMS
Microbiol. Ecol. 16., 177-184.
Muyzer, G.; Dewaal, E.C. & Uitterlinden, A.G. (1993). Profiling of complex microbial-
populations by denaturing gradient gel-electrophoresis analysis of polymerase
chain reaction-amplified genes-coding for 16s ribosomal-RNA. Appl. Environ.
Microbiol. 59(3): 695-700.
Neufeld, J.D.; Schafer, H. Cox, M.J. Boden, R. McDonald, I.R. & Murrell, J.C. (2007). Stable-
isotope probing implicates Methylophaga spp and novel Gammaproteobacteria in
marine methanol and methylamine metabolism. ISME Journal. 1., 480-491.
Op den Camp, H.J.M.; Kartal, B. Guvent, D. van Niftrik, L.A.M.P. Haaijer, S.C.M. van der
Star, W.R.L. van de Pas-Schoonen, K.T. Cabezas, A. Ying, Z. Schmid, M.C. Kuypers,
M.M.M. van de Vossenberg, J. Harhangi, H.R. Picioreanu, C. van Loosdrecht,
M.C.M. Kuenen, J.G. Strous, M. & Jetten, M.S.M. (2006). Global impact and
application of the anaerobic ammonium-oxidation (anammox) bacteria. Biochemical
Society Transactions. 34., 174-178.
Penton, C.R.; Devol, A.H. & Tiedje, J.M. (2006). Molecular evidence for the broad
distribution of anaerobic ammonium-oxidizing bacteria in freshwater and marine
sediments. Appl. Environ. Microbiol. 72., 6829-6832.
Richard, F.A. (1965). Anoxic basins and fjords. In: Chemical oceanography, vol 1. Riley, J.P.&
Skirrow, G. (Ed.), Academic Press, London, 611-645.
Risgaard-Petersen, N.; Meyer, R.L. Schmid, M. Jetten, M.S.M. Enrich-Prast, A. Rysgaard, S.
& Revsbech, N.P. (2004). Anaerobic ammonium oxidation in an estuarine sediment.
Aquat Microb Ecol. 36., 293–304.
Risgaard-Peterson, N.; Nielsen, P. L. Rysgaard, S. Dalsgaard, T. & Meyer, R. L. (2003).
Application of the isotopic paring technique in sediments where anammox and
denitrification coexist. Limnology and Oceanography: method. 1., 63-73.
Ritter, W.F.& Chirnside, A.E.M. (1995). Impact of dead bird disposal pits on groundwater
quality on the delmarva peninsula. Bioresour. Technol. 53., 105-111.
Ruscalleda, M.; López, H. Ganiqué, R. Puig, S. Balaguer, M.D. & Colprim, J. (2008).
Heterotrophic denitrification on granular anammox SBR treating urban landfill
leachate. Water Sci. technol. 58., 1749-1755.
Rysgaard, S. & Glud, R.N. (2004). Anaerobic N2 production in Arctic sea ice. Limnol.
Oceanogr. 49., 1., 86-94.
Schmid, M.C.; Maas, B. Dapena, A. van de Pas-Schoonen, K. van de Vossenberg, J. Kartal, B.
van Niftrik, L. Schmidt, I. Cirpus, I. Kuenen, J.G. Wagner, M. Sinninghe Damsté, J.
S. Kuypers, M.M.M. Revsbech, N.P. Mendez, R. Jetten, M.S.M. & Strous, M.
(2005). Biomarkers for the in situ detection of anaerobic ammonium oxidizing
(anammox) bacteria. Appl. Environ. Microbiol. 71., 1677–1684.
Schmid, M.C.; Risgaard-Petersen, N. van de Vossenberg, J. Kuypers, M.M.M. Lavik, G.
Petersen, J. Hulth, S. Thamdrup, B. Canfield, D. Dalsgaard, T. Rysgaard, S. Sejr,
M.K. Strous, M. den Camp, H.J.M.O. & Jetten, M.S.M.( 2007). Anaerobic
ammonium-oxidizing bacteria in marine environments: widespread occurrence but
low diversity. Environ. Microbiol. 9., 1476-1484.

Anaerobic Ammonium Oxidation in Waste Water - An Isotope Hydrological Perspective
107
Schubert, C. J.; Durish-Kaiser, E. Wehrli, B. Thamdrup, B. Lam, P.& Kuypers, M.M.M.
(2006). Anaerobic ammonium oxidation in a tropical freshwater system (Lake
Tanganyika). Environmental Microbiology. 8., 1857-1863.
Sigman, D.M.; Casciotti, K.L. Andreani, M. Barford, C. Galanter, M. & Böhlke, J.K. (2001). A
bacterial method for the nitrogen isotopic analysis of nitrate in marine and fresh
waters. Analytical Chemistry. 73., 4145–4153.
Sinninghe Damsté, J.S.; Rijpstra, W.I.C. Schouten, S. Fuerst, J.A. Jetten, M.S.M. & Strous, M.
(2004). The occurrence of hopanoids in planctomycetes: Implications for the
sedimentary biomarker record. Org Geochem. 35., 561–566.
Strous, M.; Fuerst, J. A. Kramer, E.H.M. Logemann, S. Muyzer, 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., 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. Appl. Microbiol. Biotechnol. 50., 589-596.
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. Médigue, C. Collingro, A. Snel, B. Dutilh, B.E. et al., (2006).
Deciphering the evolution and metabolism of an anammox bacterium from a
community genome. Nature. 440., 790–794.
Strous, M.; van Gerven, E. Kuenen, J.G. & Jetten, M.S.M. (1997a). Effects of aerobic and
microaerobic conditions on anaerobic ammonium-oxidizing (anammox) sludge.
Appl. Environ. Microbiol. 63., 2446-2448.
Strous, M.; van Gerven, E. Ping, Z. Kuenen, J.G. & Jetten, M.S.M. (1997b). Ammonium
removal from concentrated waste streams with the Anaerobic Ammonium
Oxidation (Anammox) process in different reactor configurations. Water Res. 31.,
1955-1962.
Tekin, E. (2010) Anammox in contaminated ground water. Thesis (in preparation).
University of Ottawa.
Tamura, K., Dudley, J., Nei, M., and Kumar, S. (2007). MEGA4: Molecular evolutionary
genetics analysis (MEGA) software version 4.0. Mol. Biol. Evol. 24., 1596-1599.
Thamdrup, B. & Dalsgaard, T. (2002). Production of N2 through anaerobic ammonium
oxidation coupled to nitrate reduction in marine sediments. Applied and
Environmental microbiology. 68., 1312-1318.
Thamdrup, B.; Dalsgaard, T. Jensen, M.M. Ulloa, O. Farías, L. & Escribano, R. (2006).
Anaerobic ammonium oxidation in the oxygen-deficient water off northern Chile.
Limnol. Oceanogr. 51., 2145-2156.
Third, K. A.; Slickers, A.O. Kuenen, J.G. & Jetten, M.S.M. (2001). The CANON system
(completely autotrophic nitrogen-removal over nitrite) under ammonium
limitation interaction and competition between three groups of bacteria. System.
Appl. Microbe. 24., 588-596.
Trimmer, M.; Nicholls, J.C. & Deflandre, B. (2003). Anaerobic ammonium oxidation
measured in sediments along the Thames estuary, United Kingdom. Appl. Environ.
Microb.69., 6447-6454.
Tsushima, I,; Kindaichi, T. & Okabe, S. (2007). Quantification of anaerobic ammonium-
oxidizing bacteria in enrichment cultures by real-time PCR. Water Res. 41.,785–794

Waste Water - Treatment and Reutilization
108
Umezawa, Y.; Hosono, T.; Onodera, S. Siringan, F. Buapeng, S. Delinom, R. Yoshimizu, C.
Tayasu, I. Nagata, T. & Taniguchi, M. (2008). Sources of nitrate and ammonium
contamination in groundwater under developing Asian megacities. Sci. Total
Environ. 404., 361-376.
van de Graaf, A. A.; de Bruijn, P. Robertson, L.A. & Kuenen, J.G. (1997). Metabolic pathway
of anaerobic ammonium oxidation on the basis of N-15 studies in a fluidized bed
reactor. Microbology-UK. 143., 2415-2421.
van de Graaf, A. A.; Mulder, A. de Bruijn, P. Jetten, M.S.M. Robertson, L.A. & Kuenen, J.G.
(1995). Anaerobic oxidation of ammonium is a biologically mediated process. Appl.
Environ. Microbial. 61., 1246-1451.
Van der Star, W.R.I.; Abma, W.R. Blommers, D., Mulder, J.W. Tokutomi, T. Strous, M.
Picioreanu, C. & van Loosdrecht, M.C.M. (2007). Startup of reaction for anoxic
ammonium oxidation: Experiences from the first full-scale Anammox reactor in
Rotterdam. Wat. Res. 41., 4149-4163.
Van Loosdrecht, M.C.M. (2008). Innovative nitrogen removal. In: Biological Wastewater
Treatment Principles, Modelling and Design. Henze, M. et al., (Eds.). IWA Publishing,
London, UK.
Van Dongen, U.; Jetten, M.S.M. & van Loosdrecht, M.C.M. (2001). The SHARON-Anammox
process for treatment of ammonium rich wastewater. Water Sci.Technol. 44., 1.,153-
160.
Wada, E.; Kadonaga, T. & Matsuo, S. (1975). 15N abundance in nitrogen of naturally
occurring substances and global assessment of denitrification from isotopic view
point. Geochemical Journal. 9., 139–148.
Wett, B. (2006). Solved upscaling problems for implementing deammonification of rejection
water. Water Sci. Technol. 53., 12., 121–128.
Wyffels, S.; Boeckx, P. Pynaert, K. Zhang, D. van Cleemput, O. Chen, G. & Verstraete, W.
(2004). Nitrogen removal from sludge reject water by a two-stage oxygen-limited
autotrophic nitrification denitrification process. Water Sci. Technol. 49., (5–6)., 57–64.