REVIEW

Presence and detection of anaerobic ammonium-oxidizing(anammox) bacteria and appraisal of anammox processfor high-strength nitrogenous wastewater treatment: a review

Akihiko Terada • Sheng Zhou • Masaaki Hosomi

Received: 30 September 2010 / Accepted: 15 February 2011

� Springer-Verlag 2011

Abstract Until now, anaerobic ammonium oxidation

(anammox) has been widely applied as an alternative

method to the conventional nitrification–denitrification

pathway for biological nitrogen removal from wastewater.

Since their discovery in a denitrifying fluidized bed reactor

in the Netherlands in the early 1990s, anammox bacteria

have also been detected in natural environments. Anam-

mox is one of the newly found drivers known to contribute

to the biogeochemical nitrogen cycle. In the marine envi-

ronment, more than 50% of nitrogenous compounds are

reportedly converted into nitrogen gas via the anammox

pathway. These observations were made using state-of-the-

art techniques for detecting anammox bacteria based on

their lipids, small-subunit ribosomal RNA genes, func-

tional genes, and unique reaction pathways. The research

objectives for anammox bacteria are quite diverse, ranging

from the application of anammox processes to various

wastewater types, to anammox biochemistry and phylog-

eny, to elucidating how anammox bacteria have evolved.

Since the genome of the anammox bacterium Kuenenia

stuttgaritiensis was deciphered, anammox bacteria have

proved to be quite versatile. The next challenge is to enrich

knowledge of anammox bacterial physiology and phylog-

eny to improve their use in engineered and natural envi-

ronmental systems and minimize nitrogen loads to

downstream water bodies. Furthermore, rapid startup of the

anammox process for engineered systems is required to

broadly harness the benefits of anammox bacteria. This

review article summarizes the physiology and phylogeny

of anammox bacteria, detection methods of anammox

bacteria and reactions, the behavior of anammox bacteria

in natural environments, and recent developments in their

use for engineered systems.

Keywords Anaerobic ammonium oxidation � Anaerobic

ammonium-oxidizing bacteria � Biofilm � Microbial

ecology � Planctomycetes

Introduction

Nitrogenous compounds, discharged in municipal and

industrial wastewaters, are drivers of eutrophication, which

enhances excessive growth of green algae and cyanobac-

teria, in many developing and developed countries. The

excessive growth of the algae and cyanobacteria reduces

the oxygen concentration in the receiving waters and,

additionally, some cyanobacteria excrete toxins (Sangolkar

et al. 2006). This leads to decreases in the dissolved oxygen

concentration, water stagnation and the eventual death of

fish and shellfish in the water body. Nitrite and nitrate

contamination of water and groundwater is an emerging

problem worldwide, and originates from anthropogenic

nitrogen pollution.

To reduce the nitrogen loading to water bodies, bio-

logical nitrogen removal which comprises nitrification and

denitrification has been an essential biochemical step.

Nitrification is a key process for nitrogen removal in most

municipal and industrial wastewater treatment plants

(WWTPs) and constructed wetlands as well as natural

environments. The term ‘‘nitrification’’ refers to the aero-

bic, sequential oxidation of ammonium to nitrite, and

nitrite to nitrate. These two steps are catalyzed by spe-

cialized chemolithoautotrophic prokaryotes known as

A. Terada (&) � S. Zhou � M. Hosomi

Department of Chemical Engineering,

Tokyo University of Agriculture and Technology,

2-24-16 Naka-cho, Koganei, Tokyo 184-8588, Japan

e-mail: akte@cc.tuat.ac.jp

123

Clean Techn Environ Policy

DOI 10.1007/s10098-011-0355-3

ammonia-oxidizing bacteria (AOB) (Wiesmann 1994),

ammonia-oxidizing archaea (AOA) (Konneke et al. 2005;

Francis et al. 2007), and nitrite-oxidizing bacteria (NOB)

(Wiesmann 1994). Currently, no microorganisms have been

reported to be capable of mediating both autotrophic nitrifi-

cation steps. Denitrification is defined as the anoxic process in

which nitrite and nitrate are reduced to gaseous nitrogen

oxides (nitric oxide (NO), nitrous oxide (N2O), and free

nitrogen (N2)) by facultative anaerobic heterotrophic deni-

trifying bacteria. These conventional nitrification–denitrifi-

cation pathways have reportedly been central to biological

nitrogen removal in engineered systems and natural envi-

ronments (Ahn and Choi 2006; Van Hulle et al. 2010) and are

considered to be of major importance in the global nitrogen

cycle (Jetten 2008; Jetten et al. 2009; Zhu et al. 2010).

In 1995, anaerobic ammonium oxidation (anammox)

was accidently discovered in a fluidized bed denitrifying

reactor where influent ammonium was suddenly lost con-

comitant with nitrate reduction following long-term oper-

ation of the reactor (Mulder et al. 1995). The discovery of

the anammox reaction was corroborated from the thermo-

dynamic calculations of Broda (1977), which revealed that

ammonia oxidation in the absence of oxygen but in the

presence of nitrite/nitrate is thermodynamically feasible. It

has been 15 years since two papers on the discovery of the

anammox reaction were published in 1995 (Mulder et al.

1995; van de Graaf et al. 1995).

Initially, the electron acceptor for the anammox reaction

was considered to be nitrate; but this was corrected to

nitrite and the following anammox stoichiometry was

proposed in 1998 (Strous et al. 1998):

1NHþ4 þ 1:32NO�2 þ 0:066HCO�3 þ 0:13Hþ ! 1:02N2

þ 0:26NO�3 þ 0:066CH2O0:5N0:15 þ 2:03H2O

ð1Þ

A few papers have also reported nitrogen loss in the

absence of organic carbon to support the anammox reaction

(Kuai and Verstraete 1998; Siegrist et al. 1998); however, the

conclusions as to whether anammox is a biologically

mediated reaction or not has been disputed for some time.

This is because the anammox bacteria have an extremely

slow growth rate, making it difficult to confirm the anammox

reactions reported by other researchers. However, anammox

bacteria were successfully enriched to 70% in a sequencing

batch reactor (SBR) (Strous et al. 1998) and concentrated up

to 99.6% by density-gradient centrifugation (Strous et al.

1999a), leading to the conclusion that anammox is a

biological reaction. Strous et al. (1999a) successfully

enriched a new autotrophic member of the order

Planctomycetales (in the phylum Planctomycetes), and

identified it as anammox bacteria. This paper revolutionized

research on better uses of anammox bacteria for nitrogen

removal in engineered systems. Since then, many engineered

systems for biological nitrogen removal have been developed

and have been reviewed elsewhere (Kuenen and Jetten 2001;

Ahn 2006). Currently, anammox-related technology has the

potential to achieve an energy-neutral or even energy-positive

process (Kartal et al. 2010).

The number of scientific papers related to anammox,

published in journals registered in the ISI Web of Science

(Thomson Reuters Inc.) has been increasing year by year as

shown in Fig. 1, broadening the research field from engi-

neered systems to the exploration of anammox to natural

environments. Studies on natural environments have required

us to reconsider the global nitrogen cycle, as anammox is one

of the major biogeochemical processes in marine and estua-

rine environments (Francis et al. 2007). Many surprising and

exciting discoveries regarding anammox ubiquity in aquatic

environments have been reported (Jetten et al. 2003), mainly

due to rapidly developing molecular techniques which mainly

focus on small-subunit ribosomal RNA (16S rRNA) and

functional genes in a culture-independent manner (Wagner

et al. 2006; Amann and Fuchs 2008). This review article

summarizes the physiology and phylogeny of anammox

bacteria, detection methods for anammox bacteria and reac-

tions, the presence and behavior of anammox bacteria in

natural systems, and recent developments on the use of these

bacteria for wastewater treatment.

Involvement of anaerobic ammonium oxidation

in the biological nitrogen cycle

Microbiologically, nitrogen compounds are involved in

five biological processes: nitrification which involves two

0

20

40

60

80

100

120

140

160

1994 1996 1998 2000 2002 2004 2006 2008 2010

The

num

ber

of I

SI p

ublic

atio

n

Year

Discovery of anammox in marine sediment (Thamdrup and Dalsgaard, 2002)

Completion of the genome sequence of K. stuttgartiensis(Strous et al., 2006)

Discovery of anammox in the ocean (Dalsgaard et al., 2003; Kuypers et al., 2003)

Discovery of anammox (Mulder et al., 1995)

Proof of anammox as biologically mediated process (Strous et al., 1999)

The number of papers on anammox research

The number of papers on anammox research for bioreactors

Fig. 1 Increase in the number of publications related to anaerobic

ammonium oxidation since 1995. Comments represent ground-

breaking papers on anammox research

A. Terada et al.

123

metabolic processes, i.e., ammonium and nitrite oxidation,

denitrification, dissimilatory nitrate reduction to ammo-

nium (DNRA), and anammox. In engineered systems, the

nitrogenous constituents, mainly organic nitrogen and

ammonium, are converted to nitrogen gas through con-

ventional nitrification–denitrification and partial nitrifica-

tion to nitrite (nitritation) via anammox pathways (Ahn

2006). Typical nitrogen removal processes are summarized

in Fig. 2. The nitrification–denitrification pathway has been

commonly utilized in engineered systems such as con-

structed wetlands (Knight et al. 1999; Chavan et al. 2008;

Lee et al. 2009a) and WWTPs (van Loosdrecht 2008).

Recently, intensive efforts have focused on nitrogen

removal via nitrite rather than nitrate due to the lower

energy input requirement. Nitrogen removal via nitrite, a

short-cut nitrogen removal process, allows for a lower

oxygen requirement with a stoichiometric value of 3.43

g-O2/g-N (via nitrate: 4.57 g-O2/g-N); a lower organic

carbon requirement as a denitrification electron donor with

a stoichiometric value of 1.71 g-COD/g-N (via nitrate:

2.86 g-COD/g-N); and less sludge production with 0.73

g-volatile suspended solid (VSS)/g-N (via nitrate: 1.19

g-VSS/g-N) compared with nitrogen removal via nitrate

(Rittmann and McCarty 2001). Applications using the

anammox process are more beneficial than nitrogen

removal via nitrite. The partial ammonium oxidation to

nitrite, the stoichiometric value suitable for the subsequent

anammox reaction of an ammonium/nitrite molar ratio of

1/1.32 (Strous et al. 1998), requires much less oxygen at

1.75 g-O2/g-N (Terada et al. 2007), no organic carbon as

an electron donor, and produces less sludge at 0.14 g-VSS/

g-N (Strous et al. 1998). Therefore, this approach reduces

power requirements by 64% (due to reduced aeration),

COD demands by 100% and sludge production by 80–90%

compared with nitrogen removal via nitrate, creating sig-

nificant benefits for the treatment of high-strength nitrog-

enous wastewaters (van Loosdrecht 2008). Thus, nitritation

coupled to anammox is an innovative process for biological

nitrogen removal from wastewater (Sliekers et al. 2002;

Pynaert et al. 2003; van der Star et al. 2007; Kuenen 2008).

One engineering challenge for nitrogen removal via a

nitritation–anammox pathway is to stabilize nitritation

performance. Suppression of nitrite oxidation is a key for

success of the nitritation–anammox pathway. Indeed,

nitrification system perturbation/malfunction, e.g., exces-

sive oxygen supply, potentially leads to conversion of

nitrite into nitrate which cannot be used for anammox

bacteria. This eventually impairs the anammox activity,

resulting in system failure of a nitritation–anammox pro-

cess (Strous et al. 1999b; Third et al. 2005; Terada et al.

2007). Given that the metabolic workload during nitrifi-

cation is divided between two physiologically and phylo-

genetically distinct groups of microorganisms, AOB

(or AOA) and NOB (Francis et al. 2007), AOB need to

over-grow NOB so that nitrite can successfully accumulate

in the bioreactor. Several technical approaches have been

used to preferentially enrich AOB over NOB by employing

their different eco-physiological properties, e.g., tempera-

ture (Balmelle et al. 1992; Jetten et al. 1997), solid reten-

tion time (SRT) (Hellinga et al. 1998), free ammonia (FA)

(Kim et al. 2006; Ganigue et al. 2007), and dissolved

oxygen (Bernet et al. 2005; Ahn et al. 2008; Terada et al.

2010). Not only suspended cultures, but also biofilms can

take advantage of these strategies for successful nitritation.

Especially, biofilm systems operating at high FA concen-

trations and/or limited oxygen loading (Pynaert et al. 2004)

plus the inoculum composition (Terada et al. 2010) can

allow for nitritation, but the challenge of fine adjustment

and control of the oxygen loading remains. In addition, the

long-term nitritation stability is still questionable, which

warrants future challenge (Fux et al. 2004). The details of

various criteria on successful nitritation are summarized in

several review articles (Schmidt et al. 2003; Khin and

Annachhatre 2004; Ahn 2006; Sun et al. 2010; Van Hulle

et al. 2010).

Phylogeny of anammox bacteria

The 16S rRNA sequence of an anammox microorganism,

Candidatus Brocadia anammoxidans, was identified for the

first time by Strous et al. (1999a), showing that Ca. Bro-

cadia anammoxidans is affiliated to the phylum Plancto-

mycetes. Numerous research has been carried out since

then. As shown in Fig. 3, which shows the 16S rRNA gene-

based phylogenetic tree as reconstructed by ARB software

NH4+

NO2- NO3

-

NON2O

NitrificationAerobic

DenitrificationAnoxic

NOB

Denitrifying bacteria

Nitritation-anammox

N2

AnammoxAnammox bacteria

AOB and AOA

DNRAOrg-N

NH2OH

PONDON

Nit

roge

n fi

xati

on

Fig. 2 Microbial nitrogen transformation in natural and engineered

systems. Methane-dependent denitrification is excluded here. AOAammonia-oxidizing archaea, AOB ammonia-oxidizing bacteria,

DNRA dissimilatory nitrate reduction to ammonium, DON dissolved

organic nitrogen, PON particulate organic nitrogen, and NOB nitrite-

oxidizing bacteria

Review of anammox detection, presence, and application

123

(Ludwig et al. 2004), anammox bacteria have been divided

into five genera, Ca. Brodacia, Ca. Kuenenia, Ca. Scalin-

dua, Ca. Anammoxoglobus, and Ca. Jettenia (Jetten et al.

2005; Kartal et al. 2007b; Quan et al. 2008). Generally, Ca.

Scalindua have been detected as the predominant anammox

bacteria in marine environments such as sediment and the

oxygen minimum zone (OMZ) (Kuypers et al. 2003; Sch-

mid et al. 2003; Dong and Reddy 2010; Li et al. 2010b),

and high-temperature oil resorviors (Li et al. 2010a) while

the other anammox genera have generally been found in

bioreactor systems for wastewater treatment (Egli et al.

2001; Pynaert et al. 2003; Kindaichi et al. 2007; van der

Star et al. 2008), hot spring (Jaeschke et al. 2009a), and

terrestrial environments (Amano et al. 2007; Dale et al.

2009; Humbert et al. 2010), exhibiting a ubiquitous pres-

ence (Kuenen 2008; Schmid et al. 2007). These anammox

bacteria are deeply branched in the Planctomycetes phy-

lum, with low 16S rRNA similarities to other Planctomy-

cetales genera, e.g., Gemmata, Isosphaera, Planctomyces,

and Pirellula (below 80%) as shown in Fig. 3 (Schmid

et al. 2005). Even in the branch of anammox bacteria, each

distinct genus has as low as a 90% similarity in 16S rRNA

sequences. The detailed reason for this lack of similarity

remains unresolved, but the Planctomycetes phylum seems

to have evolved very slowly and may be less sensitive to

evolutional changes. This has the potential to permit the

16S rRNA gene of each individual species and genera to

become separated. It is also possible that Planctomycetes

could be the first bacterial group to emerge (Brochier and

Philippe 2002). After the deciphering the genome of Ca.

Kuenenia stuttgartiensis (Strous et al. 2006), the detailed

anammox reaction mechanisms and the metabolic versa-

tility of anammox bacteria have been investigated in depth.

The detailed results can be found in a review article (Jetten

et al. 2009).

Anammox versatility has been recently reported, with

Ca. Brocadia fulgida and Ca. Anammoxoglobus propio-

nicus capable of taking up acetate (Kartal et al. 2008) and

propionate (Kartal et al. 2007b) along with ammonium,

respectively. These anammox bacteria are likely to pre-

vail over other anammox bacteria when such volatile fatty

acids are present with ammonium. Furthermore, Ca.

Kuenenia is able to mediate DNRA and, potentially, so

are the other anammox bacterial genera (Kartal et al.

2007a). These different physiological characteristics

could be a key for determining the predominant anammox

bacteria in both natural environments and engineered

systems.

Detection of anammox bacteria and reactions

The discovery of anammox reactions in natural environ-

ments and advances in anammox-related technologies are

attributed to the development of diversified methods

focusing on anammox bacterial cells/genome and anam-

mox reactions. The overview of methods which have been

used for detecting anammox reactions and bacteria are

shown in Fig. 4. In this chapter, each method used for

anammox detection is introduced.

Planctomyces

Candidatus Scalindua wagneri, AY254882Candidatus Scalindua brodae, AY257181

Candidatus Kuenenia stuttgartiensis, AF375995

Candidatus Anammoxoglobus propionicus, DQ317601

Candidatus Jettenia asiatica, DQ301513

Candidatus Brocadia fulgida, DQ459989

Candidatus Brocadia anammoxidans, AF375994

Isosphaera

Gemmata

Pirellula

0.10

To outgroups

Fig. 3 Phylogenetic tree based

on 16S rRNA gene for the

phylum Planctomycetesconsisting of five genera of

previously reported anaerobic

ammonium-oxidizing bacteria

and the genera Planctomyces,

Pirellula, Gemmata, and

Isophaera. The tree

reconstruction was carried out

by neighbor-joining analysis

with a 50% conservation filter

for bacteria. The bar represents

10% sequence divergence. This

tree was reconstructed using the

ARB software package (Ludwig

et al. 2004)

A. Terada et al.

123

Polymerase chain reaction for 16S rRNA and functional

genes

As of January 2011, the isolation of an anammox bacterium

has yet to be successful. Therefore, culture-independent

methods, targeting either 16S rRNA or functional genes,

are the most widely used approaches. To detect anammox

bacteria in environmental samples, PCR amplification with

general 16S rRNA gene-targeted primers and subsequent

quantitative or phylogenetic analyses are a central method,

in common use. Bacterial diversity studies have also been

used, and these are summarized elsewhere (Muyzer 1999;

Fromin et al. 2002; Schutte et al. 2008).

The 16S rRNA or functional gene-based approach

without cultivation is a very powerful technique when

anammox bacteria have low activity or samples contain a

lot of inert particulates, both of which hamper the appli-

cation of the fluorescence in situ hybridization (FISH)

method which will be explained in ‘‘Fluorescence in situ

hybridization’’ section (Schmid et al. 2005). Generally,

there are two drawbacks to the PCR method: low DNA

extraction yield and the production of artifacts after PCR.

The low DNA recovery after extraction is generally

dependent on the method and type of samples. Different

protocols for DNA extraction are reported to provide dif-

ferent degrees of DNA yield, potentially underestimating

the microbial community in respect of AOA (Urakawa

et al. 2010). There are currently no reports on the rela-

tionship between the degree of DNA yield and microbial

community diversity and density for anammox bacteria.

However, the precaution of selection of an appropriate

DNA extraction method should be undertaken. Numerous

primer sets are available for targeted anammox bacteria in

different environmental systems (Schmid et al. 2000, 2003,

2005; Penton et al. 2006).

The use of 16S rRNA as a biomarker of anammox

bacteria lacks a direct relationship to their physiologies,

which might be a disadvantage when investigating the

microbial community responsible for anammox. The

information obtained from 16S rRNA only identifies the

existence of DNA for targeted bacteria but not their

activities. The use of the 16S–23S rRNA intergenic spacer

region (ISR) could be a good candidate to identify anam-

mox activity. Anammox bacteria, particularly Ca. ‘Broca-

dia anammoxidans’ and Ca. ‘Kuenenia stuttgartiensis’, are

reported to have an ISR of approximately 450 bp, which

divides the 16S and 23S rRNA genes (Schmid et al. 2001).

In contrast, most other members of the Planctomycetales,

i.e., the genera Pirellula, Planctomyces, Isophaera, and

Gemmata do not have that ISR but possess unlinked 16S

and 23S rRNA genes. The difference can be detected by

molecular based quantification methods such as FISH

(Schmid et al. 2001) or quantitative PCR (Park et al. 2010).

The benefit of using the ISR region to detect anammox

Environmental Samples

Target:anammox reaction

In situ activity in a sediment or biofilm

Microsensor technique

Elucidation of anammox reaction

pathway

Tracer technique by stable isotope

Target:anammox bacteria

DNA/RNA in a bacterial cell

16S rRNA gene

Intergenic spacer region gene

Identification (cloning, finger printing),

quantification (quantitative PCR)

Functional gene

Identification (cloning, finger printing),

quantification (quantitative PCR)

Bacterial cell

16S rRNA gene

Intergeneicspacer region gene

FISH, CARD-FISH etc

Protein (cytochrome c)

Confocal Raman microscopy

Membrane lipid

GC/MS or NMR

Advantage

in situ measurement of chemical species without sample destruction

Able to distinguish anammox reaction from denitrification by 15N labeled NH4

+ or NO2

-

Able to identify and quantify anammox bacteria based on bacterial phylogeny without cultivation

Able to identify and quantify anammox bacteria based on bacterial function without cultivation

Able to detect activity based on transcriptome analysis (mRNA)

in situ detection of anammox bacteria without cultivation and sample destruction

Able to quantify anammox bacteria with spatial information kept

in situ detection of anammox bacteria based on cytochrome c

No need for pretreatment like FISH

Able to explicitly distinguish anammox bacteria (‘ladderane lipid’ is only a biomaker for anammox bacteria)

Diisadvantage

Difficult to specify a single reaction (necessary to link microsensor technique with a molecular method)

Incapable of maintaining spatial information (if a sample is an aggregate, sediment or biofilm)

Need to consider biases due to PCR amplification

Need to consider bias due to PCR amplification

Difficult to detect bacteria with lowactivity (FISH)

Potential artifacts due to amplification of negative signal (CARD-FISH)

Extremelyexpensive to procure a confocal Raman microscopy

Extremelyexpensive to procure an apparatus

Complicated to set up the apparatus

Fig. 4 Overview of methodologies for anammox reactions and bacterial detection

Review of anammox detection, presence, and application

123

bacteria occurs because the ISR allows monitoring of the

anammox bacteria activity in their natural and engineered

environments. For instance, Park et al. (2010) reported that

the ISR tracks and precedes the autotrophic nitrogen

removal rate, suggesting that ISR expression is a molecular

biomarker for predicting anammox reactor performances.

Functional gene markers, genes which encode key

enzymes involved in a specific metabolic pathway, are

another alternative to reflect bacterial physiology (Junier

et al. 2010). For anammox bacteria, the hydrazine oxido-

reductase (hzo) gene, which encodes for hydrazine oxido-

reductase (HZO), is a representative functional gene

biomarker. HZO, a member of the octaheme cytochrome

c hydroxylamine oxidoreductase protein family (Klotz

et al. 2008), is reported to be a key player in the anammox

biochemical process (Klotz and Stein 2008). The role of

HZO is to dehydrogenate hydrazine, an intermediate in the

anammox reaction, and to convert it into N2. HZO has been

purified from an anammox-enriched bioreactor and its role

has been investigated (Shimamura et al. 2007). The iden-

tification of the hzo gene has been used as a biomarker for

anammox bacteria (Quan et al. 2008; Schmid et al. 2008;

Park et al. 2010). A new primer set for the hzo gene has

been investigated with several environmental samples, and

is of promise for identifying the gene in ubiquitous natural

environments (Li et al. 2010b). These studies do not target

genuine anammox bacterial activity. Hence, transcriptome

analysis is essential to identify the gene activity by moni-

toring messenger RNA (mRNA), which may illuminate

anammox activity more accurately. It has been reported

that hzo mRNA, in addition to ISR, is a potential molecular

biomarker for predicting anammox reactor performances

(Park et al. 2010). Additionally, the cytochrome cd1-con-

taining nitrite reductase gene (nirS), which is unique to Ca.

Scalindua but distinct from the heterotrophic denitrifier,

nirS, has also been used with transcriptome analysis

(mRNA monitoring) to quantify the activity of Ca. Scal-

indua (Lam et al. 2009).

Application of real time quantitative PCR

Although conventional PCR amplification cannot quantify

the amount of a target gene, qPCR, and competitive PCR

(cPCR) allow targeted bacterial density to be quantified

based on 16S rRNA, ISR, or functional genes. The prin-

ciples behind these techniques are described elsewhere

(Sharkey et al. 2004; Smith and Osborn 2009). Competitive

PCR is less likely to be applied due to its lower accuracy

and the availability of cheaper qPCR apparatus. Several

researchers have applied the qPCR technique to quantify

anammox 16S rRNA (Tsushima et al. 2007a, b; Quan et al.

2008; Li et al. 2009) and functional gene (Lam et al. 2009;

Park et al. 2010) copy numbers. Quantitative PCR provides

higher throughput, more reliability and more sensitive

quantification compared with quantitative FISH. Therefore,

this technique has been used to quantify the maximum

growth rate of anammox bacteria by using a growth curve

for the bacteria (Tsushima et al. 2007a; van der Star et al.

2007). The doubling time of anammox bacteria was first

reported to be 11 days (Strous et al. 1998) but estimations

by qPCR have indicated a broader range for the maximum

growth rate of anammox bacteria, ranging from 3.6 to

17 days (Tsushima et al. 2007a; van der Star et al. 2007).

However, the disadvantage of qPCR is the fact that the

copy number of a targeted gene is quantified but not the

actual bacterial cell number. Bacteria have different copy

numbers of 16S rRNA genes in each bacterial cell

depending on the bacterial species (Klappenbach et al.

2001; Lee et al. 2009b). The 16S rRNA gene copies cannot

be directly converted into cell numbers, indicating that the

copy numbers acquired by qPCR do not necessarily indi-

cate the density of the bacteria in the microbial community.

This also occurs for anammox bacteria: they may have

different 16S rRNA or hzo operon copy numbers depend-

ing on species, which warrants future study. Therefore, the

results obtained by qPCR should be double-checked using

known amounts of the clone-derived positive control to

confirm the putative effect of competition/inhibition of a

positive control with an actual sample (Tsushima et al.

2007a).

FISH

FISH relies on DNA/RNA hybridizations occurring within

whole microbial cells in situ. In situ hybridizations with

DNA oligonucleotides designed for the detection of spe-

cific bacteria are performed with fluorescent-labeled com-

pounds. Currently, FISH is well known as a powerful

diagnostic tool with widespread environmental and medi-

cal applications, and has been reviewed by numerous

researchers (Wagner et al. 2006; Amann and Fuchs 2008).

Numerous FISH probes have been described in literature

for the detection of phylogenetically different bacteria

based on the 16S rRNA gene in mixed bacterial commu-

nities, most of which are listed in probeBase (Loy et al.

2003) (http://www.microbial-ecology.de/probebase/). With

FISH, abundant rRNA molecules in bacterial ribosomes are

probed with fluorescently labeled DNA oligonucleotides

which target specific sites on the rRNA of targeted Bac-

teria, Archaea, and Eukarya. The DNA probes hybridize

with their target sites by fluorescence after either excess or

non-hybridized probes are removed. The target fluorescent

bacterial cells can be detected by epi-fluorescent micros-

copy, confocal laser scanning microscopy or flow cytom-

etry. There are fluorochrome variations with different

excitation/emission wavelengths. Multiple staining of

A. Terada et al.

123

different probes with different fluorochromes permits

simultaneous detection of different bacteria. Precaution is

required to check if the signal from one probe interferes the

signal from another, but the general protocol has been

validated (Amann 1995). Anammox bacterial probes are

available from phylum-specific (e.g., Neef et al. 1998) to

member-specific (e.g. Schmid et al. 2003) and genus-spe-

cific probes (e.g., Schmid et al. 2000, 2001; Kuypers et al.

2003; Kartal et al. 2007a, b), which are chosen depending

on the research objective. Probes that are widely used for

the detection of anammox bacteria are reviewed elsewhere

(Schmid et al. 2005; Daims et al. 2009).

The advantages of the FISH technique are to allow a

reliable result to be obtained; a reduced likelihood of

misinterpret artifacts than PCR-related techniques because

FISH does not amplify bacterial cells; and to provide the

spatial location of bacteria and their density in a limited

region such as biofilm or bacterial aggregate without the

destruction of the sample. A requirement for the success of

FISH is that a sample does not contain a large amount of

inert material, which can interfere with a positive signal

due to their auto-fluorescent nature. The basic morphology

of anammox bacteria are coccoid and the bacteria tend to

form a cluster-like structure (Egli et al. 2001) as also seen

with nitrifying bacteria (Schramm et al. 1996; Okabe et al.

1999).

When the activity in a targeted bacterial cell is low (i.e.,

low numbers of rRNA), the FISH technique may not pro-

vide bacterial cell detection using microscopy. In this case,

fluorescent signal amplification shows promise for detect-

ing bacterial cells. Catalyzed reporter deposition (CARD)-

FISH is one technique for dramatically increasing the

sensitivity of the FISH signal. This technique was first

introduced to microbial ecology in 2002 (Pernthaler et al.

2002). With CARD-FISH, the probes are not directly

labeled with fluorochromes but are covalently bound to

horseradish peroxidase (HRP). In an amplification step, this

enzyme catalyzes the formation of fluorochrome-labeled

tyramide radicals. These highly reactive molecules bind to

tyrosine rich regions of the ribosome and other proteins in

the vicinity. With this additional catalytic step, a cell

containing only a single ribosome can theoretically be

detected since the fluorescence signal is amplified many

times by the massive deposition of labeled tyramide. The

sensitivity can reportedly be amplified by a factor of 26–41

(Hoshino et al. 2008).

To summarize, CARD-FISH is ideal when conventional

FISH techniques have a low signal and/or high background

from hybridized cells. In particular, CARD-FISH has been

used to detect anammox bacteria in oligotrophic environ-

ments such as lakes (Hamersley et al. 2009) and marine

environments (Woebken et al. 2007; Galan et al. 2009).

The CARD-FISH methods have been optimized for the in

situ detection of anammox bacteria in biological waste-

water treatment (Pavlekovic et al. 2009).

Other modified FISH techniques have also been devel-

oped to minimize probe penetration problems and increase

hybridization efficiency. For example, peptide nucleic acid

FISH (PNA FISH) (Perry-O’Keefe et al. 2001) and locked

nucleic acid FISH (LNA FISH) (Kubota et al. 2006) could

be used for the detection of anammox bacteria under oli-

gotrophic conditions. The detailed principles underlying

these methods can be found in a review article (Amann and

Fuchs 2008).

Confocal Raman microscopy

In microbiology, it is feasible to differentiate bacteria down

to the strain level by Raman microscopy (Kirschner et al.

2001). The use of confocal Raman microscopy allows for

identification of the three-dimensional distribution of

anammox bacteria in aggregates, e.g., biofilms. Anammox

bacteria (enriched Ca. Brocadia anammoxidans) have been

identified using their Raman vibrational signature in a non-

invasive manner (Patzold et al. 2006). The modified tech-

nique, i.e., confocal resonance Raman microscopy

(CRRM), permits the use of the Raman resonance effects

on cytochrome c, which is abundant in an anammox cell

comprising more than 10% of the protein content of

anammox bacteria (Cirpus et al. 2005), so that anammox

bacteria can be non-invasively identified (Patzold et al.

2008). Unlike FISH, CRRM does not require pretreatment,

i.e., bacterial cell fixation by paraformaldehyde. Further-

more, it allows for the acquisition of reliable species-level

identification, which is an alternative for detecting anam-

mox bacteria in biofilms or granules if economically

feasible.

Lipid analysis as a biomarker for anammox bacteria

The lipid membrane composition of anammox bacteria has

been investigated by gas chromatography mass spectrom-

etry (GC–MS) and high field nuclear magnetic resonance

(NMR) spectrometry (Damste et al. 2002, 2004). Anam-

mox bacteria contain a membrane-bound intracytoplasmic

compartment, i.e., anammoxosome, where ammonium and

nitrite are converted to N2 gas via hydrazine, a toxic and

energy-rich intermediate (Damste et al. 2002). Unprece-

dented lipids, known as ladderane lipids, were discovered

in both Ca. Brocadia anammoxidans and Ca. Kuenenia

stuttgartiensis cells (Damste et al. 2002). Details can be

found in the review article (Jetten et al. 2009). The

ladderane lipids in anammox bacteria can be used as a

biomarker to show their presence. They have been detected

in natural environments, such as oceanic OMZs (Kuypers

et al. 2003, 2005) and marine sediments (Jaeschke et al.

Review of anammox detection, presence, and application

123

2009b, 2010). Recently, it has been revealed that the

composition of ladderane lipids in anammox bacteria

depends on applied temperatures, which may indicate the

possibility of determining the origin of ladderane lipids in

sediments (Rattray et al. 2010).

Microsensors

Microsensors, in the form of microelectrodes and mic-

rooptodes (fiberoptic microsensors), are very powerful

tools to determine a hot spot where anammox reactions

take place in sediments or biofilms at the micrometer level.

The tip diameter for microsensors can be fabricated down

to a few micrometers. Because the minimum resolution of

microsensors is twice as large as the tip diameter, the

interval of microsensor measurement, the minimum sample

depth is at the micrometer level. The microsensor tech-

nique allows the concentrations of various dissolved

compounds to be measured without sample destruction.

To measure the depth profiles of nitrogen species, i.e.,

ammonium, nitrite, and nitrate ions, ion-selective micro-

electrodes based on Ag/AgCl have been used with engi-

neered biofilms (Schramm et al. 1996; Okabe et al. 1999).

To measure oxygen, a Clark-type dissolved oxygen

microelectrode is commercially available (http://www.

unisense.com/). Decreases in the ammonium and nitrite

concentrations concomitant with nitrate production at the

same location indicate evidence of anammox reactions in a

biofilm (Kindaichi et al. 2007; Tsushima et al. 2007b; Cho

et al. 2010). The combination of nitrite (or nitrate) and

dissolved oxygen microelectrodes has been used to identify

hot spots of anammox reactions in a biofilm during com-

plete autotrophic nitrogen removal (coupling of partial

nitritation and anammox) (Nielsen et al. 2005), in man-

groves (Meyer et al. 2005) and in estuary sediments

(Risgaard-Petersen et al. 2004). The combination of FISH

with microsensor techniques provides information on the

identity and activity of anammox bacteria (Nielsen et al.

2005; Kindaichi et al. 2007; Cho et al. 2010).

15N isotope labeling technique

The use of the stable isotope 15N is a method for quanti-

fying anammox activity in an environmental sample. The

initial observation of nitrogen loss in a bioreactor system

has been confirmed by tracer experiments by applying 15N-

labeled ammonium (Mulder et al. 1995; van de Graaf et al.

1995). Since then, isotope experiments with 15N-labeled

ammonium (or 15N-labeled nitrate) and non-labeled nitrite

(14N) have been performed to detect anammox activity

(Thamdrup and Dalsgaard 2002; Hamersley et al. 2007;

Schmid et al. 2007; Dale et al. 2009). Given the anammox

reaction pathway, combining 15N-labeled ammonium with

14N-labeled nitrite or nitrate results in the expected mixed

labeled 14?15N2, which is evidence of the anammox reac-

tion and can be differentiated from the denitrification

pathway, which produces 14?14N2. In addition to the

detection of anammox activity, the use of 15N-labeled

nitrate has revealed anammox versatility participating in

DNRA (Kartal et al. 2007a). Furthermore, an 15N tracer

study has been used to test anammox inhibition in marine

sediments (Jensen et al. 2007).

Anaerobic ammonium oxidation in natural

environments

Although anammox was speculated as a possible pathway

for producing N2 in the ocean several decades ago, the first

observation of anammox bacteria was confirmed in a

wastewater treatment reactor (Mulder et al. 1995). Since

then, a number of studies have reported the presence of

anammox in natural environments such as marine sedi-

ments (Thamdrup and Dalsgaard 2002; Dalsgaard et al.

2003; Kuypers et al. 2003), marine sponges (Mohamed

et al. 2010), estuarine/tidal river sediments (Trimmer et al.

2003; Rysgaard et al. 2004; Meyer et al. 2005; Dale et al.

2009), deep-sea hydrothermal vents (Byrne et al. 2009), hot

spring (Jaeschke et al. 2009a), and some freshwater eco-

systems (Schubert et al. 2006; Erler et al. 2008; Hamersley

et al. 2009). The discovery and potential contribution of

anammox bacteria to N2 production is important because it

may necessitate a re-evaluation of nitrogen transformation

processes in the global nitrogen budget. For example, some

evidence from marine ecosystems suggests that anammox

may in fact be the dominant nitrogen loss process in anoxic

marine water columns (Dalsgaard et al. 2003; Kuypers

et al. 2003).

Inland freshwater ecosystems such as lakes, rivers, and

wetlands have an important role in the ecological systems

on earth and may provide suitable habitats for diverse

nitrogen cycle bacteria including anammox bacteria. In

particular, ecosystems with low dissolved oxygen concen-

trations in surface water and the availability of both

ammonium and nitrite/nitrate are considered as prime

environments for the anammox process. The presence

of the anammox process has been confirmed in lakes

(Schubert et al. 2006) and anammox bacteria have also been

detected in inland rivers (Zhang et al. 2007). Furthermore,

nitrogen transformations in soil planted with forage rice

plants (Zhou et al. 2009) and bulrushes (Dong and Reddy

2010) have also indicated the occurrence of the anammox

process. As anthropogenic eutrophication increases nutrient

loading to freshwater bodies, there is an increasing need to

understand the anammox process in freshwater ecosystems

in more detail to prevent eutrophication.

A. Terada et al.

123

Diversity of anammox bacteria in various natural

ecosystems

The first marine anammox bacteria (Ca. Scalindua sor-

okinii) were reported in the Black Sea (Kuypers et al.

2003). Since then, Ca. Scalindua-like bacteria have been

detected in various marine ecosystems (Risgaard-Petersen

et al. 2004; Amano et al. 2007; Galan et al. 2009). A new

cluster, Ca. Scalindua arabiaca, was also observed in the

Arabian Sea (Woebken et al. 2008). Compared with the Ca.

Brocadia and Ca. Kuenenia genera that usually dominate in

wastewater treatment systems, Ca. Scalindua are mainly

present in natural environments, particularly in anoxic

marine environments. Studies on 11 anoxic marine sedi-

ments and water column samples from different seas found

that all detected anammox species were affiliated with the

genus Ca. Scalindua, and were grouped into Ca. Scalindua

brodae and Ca. Scalindua sorokinii (Schmid et al. 2007).

Similarly, most of the anammox bacteria identified by

cloning the 16S rRNA gene from the Seto Inland Sea are

also proximate to these two genera (Amano et al. 2007).

Tidal river estuaries and freshwater ecosystems appear

to harbor a greater diversity of anammox bacteria com-

pared with anoxic marine environments (Dale et al. 2009;

Humbert et al. 2010). These environments have various

zones where different nitrogen loading, oxygen concen-

tration, organic content, and salinity are present. In the

Cape Fear River estuary in the United States, an increase in

anammox bacterial diversity was observed along a salinity

gradient. Ca. Scalindua are the most halotolerant to high

salinity of the anammox genera, while Ca. Brocadia and

Ca. Kuenenia are less tolerant to salinity (Dale et al. 2009).

A similarly high anammox bacterial diversity was found in

the Yodo River estuary, where sequences affiliated with

Ca. Scalindua, Ca. Brocadia, and Ca. Kuenenia have been

recovered (Amano et al. 2007), and Ca. Scalindua wagneri

was predominant in the samples recovered from the same

river (Amano et al. 2007). In addition to a Ca. Scalindua-

affiliated clone identified by the 16S rRNA gene, Ca.

Brocadia-affiliated counterparts have also been recovered

from sediments of the inland Xinyi River in China,

showing a high anammox diversity in the samples (Zhang

et al. 2007). The anammox 16S rRNA gene sequences

recovered from Lake Rassnitzer in Germany are proximal

to Ca. Scalindua sorokinii/brodae, Ca. Scalindua wagneri,

and Ca. Brocadia fulgida (Hamersley et al. 2009), while

Ca. Scalindua brodae was identified in Lake Tanganyika in

Africa (Schubert et al. 2006). These results suggest a high

anammox bacterial diversity in freshwater ecosystems,

which is comparable to estuarine sediments.

Despite the detection of anammox bacteria at psychro-

philic and mesophilic temperatures (4–37�C) in natural

environments, anammox bacteria are recently reported to

be able to thrive at even higher temperature than 43�C

(Byrne et al. 2009; Jaeschke et al. 2009a; Li et al. 2010a),

which is reported to be the highest temperature to allow

enriched anammox culture Ca. Brocadia anammoxidans in

a laboratory to maintain the activity (Strous et al. 1999b).

These reports have revealed that anammox bacteria were

present and active at 60–85�C in deep-hydrothermal vents

(Byrne et al. 2009), at 52�C in California and Nevada hot

springs (Jaeschke et al. 2009a) and at 55–75�C in petro-

leum reservoirs (Li et al. 2010a). These anammox bacteria

are affiliated to not only the known anammox bacterial

genera, i.e., Ca. Brocadia, Ca. Kuenenia, and Ca. Scalin-

dua but also, intriguingly, new deep-branching phylotypes

(Byrne et al. 2009; Jaeschke et al. 2009a; Li et al. 2010a).

The phylogeny and physiology of the new anammox clades

should be elucidated in a stably controlled environment,

leading to better understanding of the nitrogen cycle in

natural environments and, in future, use of these bacteria

for engineered systems.

Contribution of anammox to total N2 flux in various

natural ecosystems

Nitrogen production via the anammox reaction has been

detected using 15N isotope labeling techniques in various

natural environments, which are typically performed with

homogenized anoxic sediment or anoxic water. The con-

tribution of anammox to total N2 flux ranges from a small

percentage to approximately 70% depending on different

environmental conditions (Table 1). At the sites in the

Celtic Sea (water depth: 500–2000 m), the highest contri-

bution of anammox to total N2 production was 65%, while

the mean anammox contribution to total N2 production at

an Irish Sea shelf (water depth: 50–100 m) was 33% (Ja-

eschke et al. 2009b). In Skagerrak, at two typical conti-

nental shelf sites (water depths: 380 and 695 m) in the

Danish belt seaway, N2 production via anammox accoun-

ted for 24 and 67% of the total N2 production, indicating

that anammox is occasionally more dominant than deni-

trification for the removal of fixed nitrogen from deep

sediments (Thamdrup and Dalsgaard 2002). Similarly, the

anammox contribution to N2 production accounted for

19–35% of the total N2 production in the water column at

Golfo Dulce (water depth: 200 m), Costa Rica (Dalsgaard

et al. 2003), and accounted for 1–35% in coastal sediments

in East and West Greenland (water depth: 36–100 m)

(Rysgaard et al. 2004).

Sediments in estuaries also allow anammox bacteria to

grow and contribute to nitrogen loss; however, the degree

of contribution of anammox reaction to nitrogen loss is

quite diverse. The contribution of anammox to total N2

production ranged from 5 to 24% in the sediment of

Review of anammox detection, presence, and application

123

Randers Fjord (water depth: 1 m) (Risgaard-Petersen et al.

2004). On the other hand, marginal anammox activities

were observed relative to denitrification in an eutrophic

coastal bay in Aarhus (water depth: 16 m) (Thamdrup and

Dalsgaard 2002) and a shallow-water estuary in Norsminde

Fjord (water depth: 0.5 m) (Risgaard-Petersen et al. 2004).

Estuaries also provide spatial gradients of nitrogen com-

pounds and dissolved oxygen, creating different niches for

anammox bacteria. In the Thames River estuary, anammox

was responsible for 8% of the total N2 production at the

inland side but this decreased along the estuarine gradient

to \1% at the mouth of the estuary, congruent with the

decrease in anammox activity (Trimmer et al. 2003). A

similar trend was discovered in a subtropical tidal river

system (Logan/Albert River), where the anammox rate

increased with increasing distance from the mouth of the

river (Meyer et al. 2005). In the Cape Fear River estuary,

the anammox activity was somewhat lower than in the

Thames River, but the contribution of anammox to the total

N2 production was higher (3.8–16.5%) (Dale et al. 2009).

These findings indicate that environmental conditions such

as salinity variation probably influence the diversity, dis-

tribution, and eventually, activity of anammox bacteria in

estuaries. Different anammox bacterial communities,

comprised of Ca. Brocadia, Ca. Kuenenia, Ca. Jettenia or

Ca. Scalindua, were detected at sites along an estuarine

gradient with the highest anammox rate found at the site

where Ca. Scalindua dominated with the highest anammox

bacterial abundance (Dale et al. 2009).

Recently, a few reports have been published relating to

anammox in inland freshwater ecosystems. In Lake

Tanganyika, the rate of N2 production through anammox

in an anoxic water column at both 100 and 110 m depths

was up to 0.01 nmol N ml-1 h-1 (Schubert et al. 2006),

similar to those measured in Golfo Dulce (Dalsgaard et al.

2003). The contribution of anammox to total N2 produc-

tion at the different depths was 9 and 13%, respectively.

It appears that anammox activity depends on the season.

In another anoxic water column at Lake Rassnitzer, the

anammox activity in January and October (max.

0.0225 nmol N ml-1 h-1) was significantly lower than in

May (Hamersley et al. 2009). The anammox activity

varied from 0.13 to 3.91 nmol ml-1 h-1 in a constructed

wetland and the presence of oxygen determined the

anammox contribution to N2 production from 0.1 to 30%,

indicating the significance of oxygen for anammox

activity (Erler et al. 2008). However, information on the

contribution of anammox activity to total N2 production

in inland ecosystems has yet to be well understood and

compared with marine and estuarine environments.

Table 1 Anammox activity in various natural ecosystems

Location Anammox rate

(nmol N ml-1 h-1)

A/T (%) NH4?

(lM)

NOx-

(lM)

Salinity

(psu)

Org. C

(%)

Reference

Marine ecosystems

Celtic Sea 1.3–2.8a 65 \130 \140 ND 0.04–0.49 Jaeschke et al. (2009a, b)

Irish Sea 2.1–25.7a 33 \130 \100 ND 0.82–1.36 Jaeschke et al. (2009a, b)

Arctic Ocean 0.17–15 1–35 ND 0.3–15.3 31.7–33.2 0.3–3.2 Rysgaard et al. (2004)

Golfo Dulce 0.042–0.061b 19–35 0–0.3 0.5–15 35 ND Dalsgaard et al. (2003)

Skagerrak 1.25–4.1 24–67 1.4–3.6 35 34.7 ND Thamdrup and Dalsgaard (2002)

Estuaries and tidal rivers

Aarhus Bay 3.46 2 ND \19 ND ND Thamdrup and Dalsgaard (2002)

Randers Fjord 3.8–5.9 5–20 ND 15–300 3–15 4.2 Risgaard-Petersen et al. (2004)

Norsminde Fjord 0 0 ND \430 3–23.5 4.4 Risgaard-Petersen et al. (2004)

Cape Fear 0.033–0.33c 3.8–16.5 4–5.2 5.5–6.9 0.4–9.9 ND Dale et al. (2009)

Thames River 0.4–9.9 \1–8 ND ND 2–30% 0.26–3.6 Trimmer et al. (2003)

Logan/Albert River 0–8 0–9 0.3–53 0.8–93 ND ND Meyer et al. (2005)

Yodo River 0.14–0.33 1–2 ND ND ND ND Amano et al. (2007)

Inland water bodies

Lake Tanganyika 0–0.01 9–13 0–27 0–10 ND ND Schubert et al. (2006)

Lake Rassnitzer 0–0.0225 20 or 50 0–300 0–200 3–23 ND Hamersley et al. (2009)

Constructed wetland 0.04–3.91 0.1–30 ND 10 ND ND Erler et al. (2008)

A/T contribution of anammox to total N2 production, ND no dataa nmol 29N2 ml-1 (overnight)b lmol N m-2 day-1

c nmol N g-1 h-1

A. Terada et al.

123

Variations in the contribution of anammox to N2 pro-

duction relative to denitrification is attributed to changes in

sediment reactivity, i.e., the availability of organic matter

in sediment for denitrifying bacteria. Following investiga-

tions in Aarhus Bay and Norsminde Fjord estuary, anam-

mox activity seems to become insignificant when sediment

reactivity increases in enriched estuarine sediments.

Although anammox rates in Randers Fjord were compa-

rable with those measured previously in the Skagerrak

sediment (Thamdrup and Dalsgaard 2002), the contribution

of anammox to total N2 production in Randers Fjord was

significantly lower than in Skagerrak sediment. This indi-

cates that the denitrification rates in Randers Fjord are

10–15 times higher than in the Skagerrak sediment, prob-

ably due to the greater availability of organic matter in the

Randers Fjord sediment.

The supply of nitrite is another key driver for both

anammox and denitrification. In natural environments,

nitrite is provided by ammonium oxidation by AOB or

AOA, or by nitrate reduction by heterotrophic bacteria

(Meyer et al. 2005). When nitrite is available, the anam-

mox bacteria consume ammonium from sediments

(Thamdrup and Dalsgaard 2002). The presence and

absence of anammox bacteria in Randers Fjord and

Norsminde Fjord, respectively, stems from differences in

the availability of nitrite in the suboxic zone of the sedi-

ment (Risgaard-Petersen et al. 2004). In OMZs, the nitro-

gen cycle including anammox has been substantially

revised. Lam et al. (2009) have identified that anammox

bacteria obtained[67% of nitrite from nitrate reduction in

the Eastern Tropical South Pacific OMZ, which is much

larger contribution than nitrite from ammonium oxidation.

Furthermore, DNRA is important in OMZs to provide

anammox bacteria with ammonium, which concludes the

complicated picture of nitrogen cycling in the Eastern

Tropical South Pacific OMZ (Lam et al. 2009).

Anaerobic ammonium oxidation in engineered systems

The paper by Strous et al. (1999a), which concluded that

the anammox reaction is biologically mediated by anam-

mox bacteria, paved the way for the application of anam-

mox-related technologies and studies on anammox

biochemistry and phylogeny. As shown in Fig. 1, the

number of publications especially related to the application

of anammox-related bioreactors for wastewater treatment

has increased explosively since 2000. Comparisons of

bioreactor system configurations, objectives and perfor-

mance are tabulated in Table 2. Numerous biochemical

processes have been developed and various wastewater

types have been used to test the feasibility of anammox

processes.

Bioreactor systems involving enrichment of anammox

bacteria

Currently, there are five known genera of anammox bac-

teria in the Planctomycetes phylum (Kuenen 2008; Jetten

et al. 2009). Although an anammox bacterium has not been

successfully isolated, the physiological properties of these

anammox bacteria have been revealed. The enrichment of

Ca. Kuenenia stuttgartiensis was successful up to 73%,

allowing the genome of the uncultured bacterium Kunuenia

stuttgartiensis to be assembled (Strous et al. 2006). Deci-

phering an environmental genome illustrates the evolu-

tional history of the bacterium and allows speculation into

the genetic blueprint of the bacterium properties. Taking

advantage of the special properties of anammox bacteria

allows the targeted anammox genus to be enriched pref-

erentially in a bioreactor system. For example, the adap-

tation of anammox biomass to high salinity conditions

allows Ca. Scalindua to be enriched in an SBR (Kartal

et al. 2006). This observation has been followed by reports

where Ca. Scalindua has been preferentially grown by

inoculating marine sediments (van de Vossenberg et al.

2008; Kawagoshi et al. 2009). The addition of acetate to

ammonium, nitrite, and nitrate provides an environment in

a bioreactor where Ca. Brocadia fulgida can outcompete

other bacteria (Kartal et al. 2008), while the presence of

both propionate and ammonium enriches Ca. Anam-

moxoglobus propionicus (Kartal et al. 2007b). Two species

affiliated with Ca. Scalindua and enriched in SBRs at 15

and 20�C, are also able to consume acetate, formate, and

propionate along with nitrate (van de Vossenberg et al.

2008).

Optimum temperature at which anammox bacteria show

the highest anammox activity depends on enriched species.

Ca. Brocadia-enriched biomass exhibited the highest

anammox activity at 40�C (Strous et al. 1999b) whereas the

highest activity was observed at 37�C in Ca. Kuenenia-

enriched biomass in a rotating biological contactor (RBC)

(Egli et al. 2001) and in multi-groups of anammox bacteria

entrapped in a polyethylene glycol gel carrier (Isaka et al.

2008). The optimum temperature range disagrees with the

counterpart at which the highest anammox activity is

attained in natural environments. Anammox bacteria dis-

covered in marine sediments exhibited the highest activity

at 25�C and lost the activity at 37�C (Dalsgaard and

Thamdrup 2002), indicating presence of anammox bacteria

with physiologically and phylogenetically different prop-

erties. Although no anammox activity was observed above

45�C in engineered systems, anammox bacteria are repor-

ted to be present even above 50�C in deep-hydrothermal

vents (Byrne et al. 2009), hot springs (Jaeschke et al.

2009a), and petroleum reservoirs (Li et al. 2010a). The

enrichment of these anammox bacteria and elucidation of

Review of anammox detection, presence, and application

123

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eret

al.

1995

)

70–840/7

0–/8

40

1.5

0.7

–1.5

0.1

584

(NH

4?

)/99

(NO

2-

)

San

dpar

ticl

esw

ere

use

das

a

carr

ier.

pH

8,

Tem

p.

36�C

Str

ous

etal

.(1

997)

Fix

edbed

2.5

910

-3

Synth

etic

Den

itri

fyin

g(f

rom

Muld

eret

al.

1995

)

70–840/7

0–/8

40

1.0

0.7

0.1

582

(NH

4?

)/99

(NO

2-

)

Sin

tere

dgla

ssw

asm

ounte

d.

pH

8,

Tem

p.

36�C

Str

ous

etal

.(1

997)

SB

R15

910

-3

Synth

etic

Den

itri

fyin

g(f

rom

Muld

eret

al.

1995

)

63–420/6

3–420

3.2

3n.d

.0.0

54

n.d

.T

he

anam

mox

stoic

hio

met

ry

was

eval

uat

ed.

pH

7–8,

Tem

p.

32–33

�C

Str

ous

etal

.(1

998)

SB

R1

910

-3

Synth

etic

Munic

ipal

WW

TP

28–400/1

5–520

1.4

8–

0.1

882

pH

7.8

–8,

Tem

p.

35�C

Dap

ena-

Mora

etal

.

(2004a)

SB

R1

910

-3

Synth

etic

Nit

rify

ing

gra

nule

s50–375/5

0–375

0.7

5–

0.4

478

pH

7.8

–8,

Tem

p.

35�C

Dap

ena-

Mora

etal

.

(2004b

)

SB

R2

910

-3

Synth

etic

Anam

mox

(80%

)257/2

03

0.4

60.3

15

0.2

5–

pH

7.8

,T

emp.

30�C

Sli

eker

set

al.

(2002)

SB

R20

910

-3

Synth

etic

Mix

edbio

mas

s

from

leac

hat

ean

d

urb

antr

eatm

ent

pla

nts

250–1268/

263–1661

1.6

0–

–79.9

–99.9

Anam

mox

abundan

ce85.8

%

atth

een

d.

pH

7.2

–8.7

,

Tem

p.

36.0

±0.3

�C

Lopez

etal

.(2

008

)

Gas

-lif

t7

910

-3

Synth

etic

Nit

rify

ing

gra

nule

s28–900/1

4–1100

2.0

–0.9

988

pH

7.8

–8,

Tem

p.

30�C

Dap

ena-

Mora

etal

.

(2004b

)

Gas

-lif

t1.8

910

-3

Synth

etic

Anam

mox

(80%

)?

nit

rifi

ers

1360/1

370

–8.9

–93

pH

7.5

Sli

eker

set

al.

(2003)

UA

SB

6.4

910

-3

Synth

etic

Met

han

ogen

ic

gra

nule

s

65–510/6

8–700

–2.9

––

pH

7.5

,T

emp.

30�C

Imaj

oet

al.

(2004)

UA

SB

0.2

Synth

etic

Met

han

ogen

ic

gra

nule

s

65–510/6

8–700

–6.4

0.1

2–0.1

386.5

–92.3

pH

7.5

,T

emp.

30�C

Imaj

oet

al.

(2004)

Upfl

ow

colu

mn

bed

14

910

-3

Synth

etic

Den

itri

fyin

g

bio

mas

s

250/2

50

1.6

1.0

3–

69

±8

Use

of

non-w

oven

shee

t.pH

7.8

9±

0.5

6,

Tem

p.3

0�C

Furu

kaw

aet

al.

(2003)

Upfl

ow

colu

mn

bed

0.8

910

-3

Synth

etic

Munic

ipal

WW

TP

20–550/2

0–460

58.5

26.0

1.6

Appro

x.

60

Choic

eof

11

dif

fere

nt

WW

TP

sas

inocu

lum

.pH

7.0

–7.5

,T

emp.

37�C

.

Tsu

shim

aet

al.

(2007a)

Upfl

ow

colu

mn

bed

0.5

910

-3

Synth

etic

(mim

ickin

gse

a

wat

er)

Mar

ine

sedim

ent

21–100/6

.4–100

0.4

0.3

n.d

.A

ppro

x.

80

Ca.

Sca

lindua

was

enri

ched

.

Tem

p.

30

�CK

awag

osh

iet

al.

(2009)

Upfl

ow

gra

nula

rbed

1.2

59

10

-3

Synth

etic

Anam

mox

bio

mas

s100–300/

100–300

Appro

x.

32.0

17.4

–60

±7

pH

8,

Tem

p.

37

±1

�CC

ho

etal

.(2

010

)

A. Terada et al.

123

Ta

ble

2co

nti

nu

ed

Pro

cess

confi

gura

tion-

obje

ctiv

e

Rea

ctor

type*

Rea

ctor

volu

me

(m3)

Was

tew

ater

type

Inocu

lum

Infl

uen

tN

H4

?/

NO

2-

conce

ntr

atio

n

(g-N

/m3)

Volu

met

ric

N-l

oad

ing

(kg-N

/m3/

day

)

Volu

met

ric

N-r

emoval

rate

(kg-N

/m3/d

ay)

Max

.sp

ecifi

c

anam

mox

acti

vit

y(g

-N/g

-

VS

S/d

ay)

N-r

emoval

effi

cien

cy(%

)

Rem

ark

Ref

eren

ces

Upfl

ow

gra

nula

rbed

3.0

59

10

-3

Synth

etic

Riv

erse

dim

ent

200/2

00

0.8

–0.5

0A

ppro

x.

80

pH

8.0

–8.5

.T

emp.

30

±5�C

Quan

etal

.(2

008)

Upfl

ow

gra

nula

rbed

1.1

910

-3

Synth

etic

Anam

mox

gra

nula

r

sludge

200–420/

220–450

45.2

4A

ppro

x.

60

1.9

283

pH

6.8

,T

emp.

35

±1�C

Tan

get

al.

(2010

)

MS

BR

5.0

910

-3

Synth

etic

Anam

mox

bio

mas

s14–366/1

5–370

0.7

30.7

10.3

5–0.4

5–

Use

of

ult

rafi

ltra

tion

mem

bra

ne

(pore

size

of

0.0

4l

m),

pH

8(i

nfl

uen

t),

Tem

p.

35

�C

Tri

go

etal

.(2

006)

MB

R15

910

-3

Synth

etic

Anam

mox

gra

nule

s1680/1

680

1.6

8–

–70–90

12

day

SR

T.

Use

of

mem

bra

ne

wit

ha

pore

size

of

0.1

lm

.pH

7.1

–7.5

,

Tem

p.

38

�C.

van

der

Sta

ret

al.

(2008)

Entr

apped

gel

1.0

910

-3

Synth

etic

Anam

mox

sludge

63–154/7

1–193

3.7

3.4

0.9

(pro

tein

bas

edw

eight)

–P

oly

ethyle

ne

gly

col

was

use

d

asa

gel

.pH

8.4

–8.6

,T

emp.

36�C

Isak

aet

al.

(2007a)

Entr

apped

gel

0.5

910

-3

Synth

etic

Anam

mox

sludge

64–148/7

1–191

66.2

(32�C

)

0.3

6(6

.3�C

)

––

Poly

ethyle

ne

gly

col

was

use

d

asa

gel

.T

emp.

6–36�C

Isak

aet

al.

(2008)

Par

tial

nit

rita

tion

and

anam

mox

SB

R1.6

Dig

este

r

effl

uen

t

Exce

sssl

udge

from

aW

WT

P

657

±56/

0.4

±0.7

–2.4

0.3

90

pH

7.5

2±

0.0

7,

Tem

p.

30�C

Fux

etal

.(2

002

)

MB

R1.5

910

-3

Rej

ecte

d

wat

er

from

dig

este

d

sludge

Bio

film

sam

ple

from

RB

Cfo

r

anam

mox

862

±113/0

0.6

5–1.1

00.5

5(t

wo

MB

Rs

for

par

tial

nit

rita

tion

and

anam

mox)

–82

Use

of

mic

rofi

ltra

tion

wit

ha

pore

size

of

0.6

lm

.pH

8.0

±0.2

,T

emp.

20–30

�C

Wyff

els

etal

.(2

004)

Upfl

ow

gra

nula

rbed

70

Munic

ipal

WW

TP

Nit

rify

ing

bio

mas

s–

10

Appro

x.

4–

Full

-sca

lepla

nt.

Rott

erdam

in

the

Net

her

lands.

van

der

Sta

ret

al.

(2007)

Sim

ult

aneo

us

Nit

rita

tion–

anam

mox

(SN

A)

SB

R2

910

-3

Synth

etic

Anam

mox

(80%

)131/0

0.1

30.0

64

0.1

5–

CA

NO

Npro

cess

.pH

7.8

,

Tem

p.

30

�CS

liek

ers

etal

.(2

002)

SB

R1.8

79

10

-3

Synth

etic

OL

AN

Dbio

mas

s

from

anR

BC

200/0

–0.4

50.1

33

–pH

7.4

–7.8

,T

emp.

35

±2�C

,D

O0.4

–1.1

g/

m3

Vla

emin

cket

al.

(2008)

SB

R1400

WW

TP

Nit

rita

tion–

anam

mox

bio

mas

s

650/0

.20.4

4–

0.1

31

[90

Dis

solv

edorg

anic

carb

on

in

infl

uen

tw

as80

g/m

3.

N2O

inves

tigat

ion.

pH

7.1

±0.2

;T

emp.

30

±3

Joss

etal

.(2

009

)

Gas

-lif

t1.8

910

-3

Synth

etic

Anam

mox

(80%

)?

nit

rifi

ers

1545/0

–1.5

–40.2

pH

7.5

Sli

eker

set

al.

(2003)

RB

C44

910

-3

Synth

etic

OL

AN

Dbio

mas

s

from

anR

BC

840/0

1.1

89

1.0

58

0.1

989

±5

OL

AN

D.

Oth

erco

nsp

icuous

Pla

nct

om

ycet

ale

sw

as

obse

rved

.pH

7.8

5±

0.1

5,

Tem

p.

29.1

±1.8

�C

Pynae

rtet

al.

(2003

)

Review of anammox detection, presence, and application

123

Ta

ble

2co

nti

nu

ed

Pro

cess

confi

gura

tion-

obje

ctiv

e

Rea

ctor

type*

Rea

ctor

volu

me

(m3)

Was

tew

ater

type

Inocu

lum

Infl

uen

tN

H4?

/

NO

2-

conce

ntr

atio

n

(g-N

/m3)

Volu

met

ric

N-l

oad

ing

(kg-N

/m3/

day

)

Volu

met

ric

N-r

emoval

rate

(kg-N

/

m3/d

ay)

Max

.sp

ecifi

c

anam

mox

acti

vit

y(g

-N/

g-V

SS

/day

)

N-r

emoval

effi

cien

cy(%

)

Rem

ark

Ref

eren

ces

RB

C50

910

-3

Synth

etic

?dig

este

r

effl

uen

t

Nit

rify

ing

bio

mas

s,

foll

ow

edby

anae

robic

gra

nula

rsl

udge

880/0

.15

–1.7

95

n.d

.–

pH

7.7

8±

0.2

6,

22.4

±1.9

�CP

ynae

rtet

al.

(2004

)

RB

C2.8

910

-3

Synth

etic

?B

lack

wat

er

OL

AN

Dbio

mas

s

from

anR

BC

1000/0

(CO

D

600

g/m

3)

–1.2

98

(Synth

etic

)/

0.7

15

(Bla

ck)

0.0

68

76

OL

AN

D.

pH

7.2

–7.7

,T

emp.

25.8

±0.4

�C,

DO

0.7

–1.1

g/m

3

Vla

emin

cket

al.

(2009)

FF

BR

8.1

910

-3

Synth

etic

?dig

este

r

effl

uen

t

Nit

rify

ing

bio

mas

s,

foll

ow

edby

anae

robic

gra

nula

rsl

udge

880/0

.15

–0.5

83

0.4

34

–T

emp.

22.4

±1.9

�CP

ynae

rtet

al.

(2004

)

SN

AP

5.4

39

10

-3

Synth

etic

Nit

rify

ing

bio

mas

s

10–100/0

0.4

8–

–60–80

Fil

led

wit

hhydro

phil

icnet

-

type

acry

lfi

ber

mat

eria

l.

pH

7.5

–7.7

,T

emp.

35�C

,

DO

2–3

g/m

3

Furu

kaw

aet

al.

(2006)

DE

MO

N�

500

Rej

ect

wat

erfr

om

dig

este

dsl

udge

––

0.6

–0.1

1–0.3

7

(CO

Dbas

e)

–S

BR

-type.

pH

contr

ol.

DO

0.3

g/m

3.

Str

asse

inA

ust

ria

Inner

ebner

etal

.(2

007

)

MA

BR

49

10

-3

Synth

etic

Anam

mox

bio

mas

s

200/0

0.8

70.7

7–

89

pH

7.9

,T

emp.

35�C

Gong

etal

.(2

007

)

MA

BR

2.4

910

-3

Synth

etic

Nit

rify

ing

bio

mas

s,

foll

ow

edby

anam

mox

bio

mas

s

530–780/0

1.1

30.7

8–

69

Seq

uen

tial

aera

tion,

Tem

p.

32�C

Pel

lice

r-N

acher

etal

.

(2010)

*A

bbre

via

tions

are

list

edbel

ow

:C

AN

ON

com

ple

teau

totr

ophic

nit

rogen

rem

oval

over

nit

rite

,D

EM

ON

�dea

mm

onifi

cati

on,

FB

Rfl

uid

ized

bed

reac

tor,

FF

BR

fixed

film

bio

reac

tor,

n.d

.not

det

erm

ined

,M

AB

Rm

embra

ne-

aera

ted

bio

film

reac

tor,

MB

Rm

embra

ne

bio

reac

tor,

MSB

Rm

embra

ne

sequen

cing

bat

chre

acto

r,R

BC

rota

ting

bio

logic

alco

nta

ctor,

SB

Rse

quen

cing

bat

chre

acto

r,SN

Asi

mult

aneo

us

nit

rita

tion

and

anam

mox,

SN

AP

single

-sta

ge

nit

rogen

rem

oval

usi

ng

anam

mox

and

par

tial

nit

rita

tion,

UA

SB

upfl

ow

anae

robic

sludge

bla

nket

,W

WT

Pw

aste

wat

ertr

eatm

ent

pla

nt,

–not

men

tioned

ina

man

usc

ript

A. Terada et al.

123

their detailed physiology would pave the way for new

paradigm of anammox research.

An extremely high nitrogen loading of more than 10 kg-

N/m3/day increases the likelihood of accumulating mono-

species anammox bacterium: e.g., a Ca. Brocadia-like

species with a 16S rRNA gene sequence similarity of

95.7% was enriched in an anammox cell aggregate known

as a granule with more than 80% dominance (Cho et al.

2010). These reports clearly indicate that the anammox

genera have unusual characteristics, not only utilizing both

ammonium and nitrite but also having a unique property

that is not outcompeted by the other anammox genera. This

type of enrichment strategy would ultimately be used for

genome analysis without isolating each bacterium.

The competition between Ca. Brocadia and Ca. Ku-

enenia, both of which are commonly found in engineered

systems for wastewater treatment, has been argued. The

affinity constants for nitrite are reported to be 2.8–42 mg-

N/L for Ca. Kuenenia (van der Star et al. 2008) and

\70 mg-N/L for Ca. Brocadia (Strous et al. 1999b). It has

been reported that Ca. Brocadia are less likely to grow at

high nitrite concentrations because of toxicity (Gaul et al.

2005), but recent publications disagree with that observa-

tion, with Ca. Brocadia detected even at high nitrite con-

centrations (Kindaichi et al. 2007; van der Star et al. 2007;

Cho et al. 2010). Given the population shift from a Ca.

Brocadia-dominant inoculum composition to Ca. Kuenenia

dominance in a membrane bioreactor (MBR) with a 16 day

SRT, van der Star et al. (2008) hypothesized that Ca.

Kuenenia is an affinity (K) strategist while Ca. Brocadia is

a growth rate (r) strategist, based on their observation of

MBR performance and FISH results (van der Star et al.

2008). The physiological properties of these two genera are

still being discussed, but the elucidation of such properties

may allow for control of the anammox community and

improvements in anammox process performance and

determine biokinetic parameters for the implementation of

a more accurate mathematical model for the prediction of

process performance.

In series partial nitritation and anammox processes

A prerequisite of the anammox process for partial nitrita-

tion and anammox (otherwise known as complete auto-

trophic nitrogen removal) is to create an environment

where the anammox bioreactor receives wastewater with

ammonium and nitrite concentrations appropriate to the

stoichiometry for the anammox reaction as shown in Eq. 1.

Thus, the ammonium has to be partly oxidized to nitrite

with an ammonium:nitrite ratio of 1:1.32 for the sub-

sequent anammox reaction (Strous et al. 1998). Hence, the

preliminary step prior to the anammox reaction, nitritation,

needs to occur in a stable process.

The single reactor system for high ammonium removal

over nitrite (SHARON) process is one possible process for

achieving the nitritation (Hellinga et al. 1998). This process

was originally developed for conventional nitrification and

heterotrophic denitrification via nitrite but has been applied

for complete autotrophic nitrogen removal. The SHARON

process takes advantage of the high-temperature of the

sludge digester after anaerobic digestion, allowing a higher

specific growth rate for AOB than NOB (Mulder et al.

2001). Because the growth rate of anammox bacteria

reaches at maximum at temperatures around 30–40�C

(Strous et al. 1999b; Egli et al. 2001; Dosta et al. 2008;

Isaka et al. 2008), the high-temperature would be particu-

larly beneficial for the anammox reaction. Generally,

investigations into the anammox reaction found in engi-

neered systems for wastewater treatment have been per-

formed at moderate temperatures of around 30–43�C as

shown in Table 2. Anammox bacteria in nature are typi-

cally detected when the temperature is well below 30�C

(Dalsgaard and Thamdrup 2002; Rysgaard et al. 2004), so

there is no apparent limit to the application of anammox at

approximately 20�C (Isaka et al. 2007b, 2008; Pathak et al.

2007). A minimum temperature of 18�C is reportedly a

threshold for the success or failure of a stable anammox

reaction (Dosta et al. 2008). Temperatures lower than 18�C

have been tested over short periods: for example, anammox

activity was reported at 11�C (Egli et al. 2001) and even at

6�C with anammox bacteria entrapped in a gel matrix

(Isaka et al. 2008). However, the nitrogen conversion rate

dramatically decreased with a decrease in temperature:

from 6.0 kg-N/m3/day at 32�C to 0.45 kg-N/m3/day at

6.3�C (Isaka et al. 2008); from 0.6 kg-N/m3/day at 30�C to

0.3 kg-N/m3/day at 15�C (Dosta et al. 2008). The anam-

mox activity loss caused by temperature decrease is

potentially mitigated by control of influent nitrite concen-

tration (Isaka et al. 2007b) or acclimatization of anammox

bacteria to low-temperature conditions (Dosta et al. 2008).

Strategy of the adaptation to and the long-term stability at

lower temperatures warrants future study.

Various types of reactors have been developed along

with the studies of anammox physiology and phylogeny

as shown in Table 2. These include fluidized bed reactors

(FBRs) (Mulder et al. 1995; Strous et al. 1997), gas-lift

reactors (Sliekers et al. 2003), upflow anaerobic sludge

blanket (UASB) reactors (Imajo et al. 2004; Tang et al.

2010), upflow column bed reactors (Fujii et al. 2002;

Kindaichi et al. 2007; Tsushima et al. 2007b), SBRs

(Strous et al. 1998; Dapena-Mora et al. 2004b, c),

membrane sequencing batch reactors (MSBRs) (Trigo

et al. 2006), and RBCs (Koch et al. 2000; Egli et al.

2001).

Given the extremely slow growth rate of anammox

bacteria, ranging from 2 to 17 days (normalized for the

Review of anammox detection, presence, and application

123

different methods used to estimate the rate) (Isaka et al.

2006; Tsushima et al. 2007a; van der Star et al. 2007,

2008), the immobilization of anammox bacteria is impor-

tant for engineering purposes to avoid washout. There are

many ways to immobilize anammox bacteria including in

an entrapped gel (Isaka et al. 2007a), a polyester non-

woven sheet as a fixed bed (Fujii et al. 2002), and sand

particles as a fixed bed (Strous et al. 1997). The immobi-

lization of anammox bacteria ensures an extremely long

SRT, and this is a good strategy to retain anammox bacteria

in a bioreactor, where the anammox bacteria are reportedly

grown in an aggregate form. However, the periodic with-

drawal of anammox biomass enhances the disaggregation

of the biomass while wastewater is permeated through a

membrane. This disaggregation ultimately eliminates the

diffusion limitation of ammonium and nitrite, ensuring

high enrichment of the anammox bacteria (van der Star

et al. 2008).

The inoculum source and composition differ according

to research (Table 2). From Table 2, the inoculum com-

position does not seem to be a determining factor for

triggering the success or failure of the anammox reaction,

but is a factor for determining the startup time. For

instance, Tsushima et al. (2007b) quantified the anammox

16S rRNA gene copy number from 11 inocula by qPCR

and have successfully reduced the anammox emergence

time to 37 days by initiating an inoculum with the highest

anammox density. This indicates the significance of inoc-

ulum selection in terms of startup time reduction. Once a

full-scale reactor has been operated for some time, the

dissemination of the anammox system should be acceler-

ated, as the inoculum from the full-scale plant can be used

as the inoculum for another plant. In addition to the engi-

neering perspective, it is also of ecological interest to

observe the effect of inoculum composition on the resultant

anammox community, stoichiometry, and performance.

Theoretically, an infinite SRT is reportedly a driver for

determining microbial populations in a bioreactor, and this

is an inherent property of a biofilm system (Terada et al.

2010). It has been recognized that, due to the non-linear

dynamics of complicated microbial communities, the

functional redundancy of bioreactors positively correlates

with the SRT (Curtis et al. 2003). It has also been reported

that different inocula lead to distinctive anammox com-

munities after long-term anammox reactor operation (Date

et al. 2009). Future studies should address the relationship

between the resultant microbial community and anammox

performance, determining whether a diversified anammox

community is useful or not, e.g., when there are fluctua-

tions in the nitrogen loadings/concentrations or changes in

environmental factors.

Since anammox reactors have been shown to be suc-

cessful at the laboratory scale, the operation of a full-size

plant with a volume of 70 m3 has commenced in Rotter-

dam, The Netherlands. The reactor aims to treat highly

concentrated ammonium effluent from the sludge digester

at the Dokhaven–Sluisjesdijk WWTP (Population Equiva-

lent 470,000, nitrogen loading 830 kg/day) (van der Star

et al. 2007). The reactor has been in full operation since

2006, and the volumetric nitrogen removal rate has reached

8–10 kg-N/m3/day, twice as high as its design capacity

(van der Star et al. 2007; Kuenen 2008). As mentioned

above, part of the anammox biomass from a reactor can be

transplanted for the startup of another full-scale WWTP.

The rapid dissemination of anammox biomass would be

expected in the near future.

Simultaneous nitritation–anammox processes

in biofilms and granules

Nitritation and anammox processes can be coupled in a

single reactor. In this situation, biofilms or aggregates

provide a redox-stratification where oxygen is either pres-

ent or absent, ensuring appropriate environments for both

nitritation and anammox (Terada et al. 2007). Thus, the

concept of a one-step autotrophic nitrogen removal process

from NH4? has been realized in processes known as

complete autotrophic nitrogen removal over nitrite

(CANON) (Sliekers et al. 2003, 2004; Third et al. 2005),

oxygen-limited autotrophic nitrification–denitrification

(OLAND) (Kuai and Verstraete 1998; Pynaert et al. 2003,

2004; Vlaeminck et al. 2009), single-stage nitrogen

removal using anammox and partial nitritation (Furukawa

et al. 2006) and aerobic deammonification (Hippen et al.

1997; Innerebner et al. 2007; Wett 2007; Wett et al. 2010).

The original definitions of OLAND and aerobic deammo-

nification assumed that AOB, especially Nitrosomonas

species, were able to denitrify, but these definitions were

corrected after the discovery of anammox bacteria (Helmer

et al. 2001; Pynaert et al. 2003). Thus, each process is

basically the same and is referred to hereafter as simulta-

neous nitritation and anammox (SNA). The spatiotemporal

control of a biofilm, i.e., control of the oxygen concentra-

tion, can be achieved by growing SNA granules in an SBR

system (Sliekers et al. 2002; Vlaeminck et al. 2008).

The application of SNA for wastewater treatment is

potentially beneficial as the SNA process has a small foot-

print and avoids the accumulation of nitrite in biofilms or

granules, which can inhibit the growth of anammox bac-

teria (van de Graaf et al. 1995; Kimura et al. 2010). Full-

scale SBR-type SNA systems with maximum volumes of

500 and 1400 m3 at nitrogen loading rates of 340 kg-N/day

and 700 kg-N/day have been operated in Strasse, Austria

(Hippen et al. 1997; Innerebner et al. 2007; Wett 2007;

Wett et al. 2010) and in Zurich, Switzerland (Joss et al.

2009), respectively, treating reject water from digested

A. Terada et al.

123

sludge at high nitrogen removal efficiency of above 85%. A

laboratory-scale RBC for SNA has demonstrated high

nitrogen removal performance for the treatment of black

water with a volumetric nitrogen removal rate of 0.715 kg-

N/m3/day (Vlaeminck et al. 2009). As can be seen in

Table 2, the SNA process appears to be suitable for treat-

ment of inorganic or low organic carbon/nitrogen ratio

wastewater.

The engineering challenge in a single-stage SNA pro-

cess is to control the redox-stratification in the biofilm or

granule (Terada et al. 2007; Vazquez-Padin et al. 2010).

Sufficient nitrogen removal in a single-stage SNA requires

fastidious micro-environmental control, as oxygen over-

load and subsequent nitrite accumulation are harmful to

anammox activity (van de Graaf et al. 1996; Hao et al.

2002; Dapena-Mora et al. 2007). Thus, complicated mon-

itoring is required for a single-stage SND, including control

of dissolved oxygen in the bulk liquid, nitrite, and biofilm

thickness (Nielsen et al. 2005; Vazquez-Padin et al. 2010).

Changes in the air flow rate, using a controller for dissolved

oxygen concentration in the bulk liquid, potentially leads to

significant shear of the bacterial aggregate, compromising

biofilm rigidity (Tsuneda et al. 2004). Therefore, careful

control of the air flow rate and choice of an appropriate

aerator are necessary when scaling up a reactor.

The use of a gas-permeable membrane may be a

potential solution for oxygen control. The membrane is

basically used as a carrier for the biofilm formation and as

an oxygen-supplying material (Syron and Casey 2008).

The oxygen supply by a gas-permeable membrane provides

bubbleless aeration, allowing independent control of both

ammonium and oxygen loadings. This biofilm has a

counter-diffusion geometry where oxygen is supplied from

the bottom of the biofilm while ammonium is supplied

from the top of the biofilm. Preliminary modeling studies

have indicated the controllability of redox-stratification in a

biofilm on a gas-permeable membrane. Thus, a membrane-

aerated biofilm (MAB) can theoretically operate under a

wider range of ammonium and oxygen loadings to achieve

sufficient nitrogen removal efficiencies than a conventional

biofilm where oxygen and ammonium are supplied from

the same direction (Terada et al. 2007; Lackner et al.

2008). The performance of MAB reactors (MABRs) have

been investigated, showing the feasibility of the MABRs

for SNA and the comparable SNA performance with con-

ventional biofilm or granule systems (Gong et al. 2007;

Pellicer-Nacher et al. 2010). The performance of MABRs

is governed by the membrane surface area (Brindle et al.

1998), so the optimization of the membrane surface area

and the fabrication of a membrane module configuration

are challenges for obtaining the highest removal perfor-

mance of SNA and long-term reactor stability (Syron and

Casey 2008).

Outlook for anammox research and concluding

remarks

Anammox research has been one of hottest topics in the

field of microbial ecology, microbiology and biotechnol-

ogy. Since the discovery of the anammox reaction and

anammox bacteria, the research has dramatically acceler-

ated. Much of this is due to the value of multidisciplinary

collaborations. Intensive microbiological research has

identified the genomic blueprint and biosynthesis of

anammox bacteria, which has provided clues to elucidate

the behavior of anammox bacteria in natural environments

and to harness anammox bacteria in engineered systems.

Conversely, the techniques for enriching specialized

anammox microorganisms have opened the door to studies

on environmental genomes.

The use of the state-of-the-art molecular techniques has

allowed investigations into zones where anammox reac-

tions takes place and into anammox bacterial density/

diversity. Anammox reactions have been detected all over

the world. However, there are some areas where anammox

reaction is not necessarily predominant in marine envi-

ronment (Ward et al. 2009) and soils (Humbert et al. 2010),

which illustrates that not all natural environments neces-

sarily harbor a niche where anammox reactions are pre-

valent over denitrification. Ward et al. (2009) suggested

that a critical factor is a difference in the magnitude and

timing of the supply of organic matter to an OMZ, which

has been discussed elsewhere (Voss and Montoya 2009).

Future studies should aim to elucidate a factor that deter-

mines the predominance of the denitrification or anammox

reactions, allowing for a more complete understanding of

the nitrogen cycle in oceans and other natural environ-

ments. In addition, recent molecular evidence of anammox

bacteria under thermophilic and hyperthermophilic condi-

tions (Byrne et al. 2009; Jaeschke et al. 2009a; Li et al.

2010a) potentially allows better comprehension of the

nitrogen cycle in natural environments and application of

these bacteria to engineered systems.

An anammox-based process for treating high-strength

nitrogenous wastewater has already been matured,

achieving a volumetric nitrogen removal rate of more than

25 kg-N/m3/day in an upflow column bed reactor (Tsu-

shima et al. 2007b) and 45 kg-N/m3/day in a UASB (Tang

et al. 2010) when suitable conditions are met, e.g., at

temperature of 35�C. As described in Chapter on Biore-

actor systems involving enrichment of anammox bacteria,

anammox activities drastically decrease at lower tempera-

ture (Strous et al. 1999b; Egli et al. 2001; Isaka et al. 2008;

Dosta et al. 2008). Hence, a startup strategy should be

developed in order to reduce time for anammox emergence

and to stabilize anammox performance at temperature

below 25�C. As proposed by Dosta et al. (2008), the

Review of anammox detection, presence, and application

123

adaptation of anammox bacteria enriched under mesophilic

conditions may be a promising strategy. The use of

anammox bacteria present in psychrophilic conditions as

inoculum could be an alternative counterpart. These

investigations should be a future challenge to accelerate

dissemination of the anammox process.

The anammox reaction is not mediated through nitrous

oxide (N2O), but nitrite accumulation through nitrification

potentially emits large amounts of N2O, known to be a

highly potent greenhouse gas (Kampschreur et al. 2009).

The amount of N2O emissions from the partial nitritation–

anammox pathway is slightly higher than that through the

conventional nitrification–denitrification pathway, which

offsets the reduced emissions of carbon dioxide (Joss et al.

2009). As high nitrite concentrations from limited oxy-

genation is a major cause of N2O emissions through the

nitrification process (Kampschreur et al. 2009), a single-

stage SNA process, e.g., CANON, OLAND, SNAP,

MABR, and deammonification, could be promising for

applications in wastewater treatment considering the

greenhouse gas emissions. Reportedly, an MABR system

successfully reduces nitrous oxide emission 100-fold lower

than two-stage nitritation and anammox reactors (Pellicer-

Nacher et al. 2010), supporting the promise of the SNA

process. However, recent reports indicate that the SNA

process encounters difficulty to maintain stable partial

nitrification (Terada et al. 2010; Cho et al. 2011). There-

fore, the stabilization of redox-stratification in a biofilm or

granule is contingent upon the replacement of two-stage

nitritation and anammox with a single-stage SNA. In

summary, an overall estimation of the energy consumption

and impact on global warming may be required when

designing the most appropriate nitrogen removal process in

the near future.

Acknowledgments This work was supported by the Environment

Research and Technology Development Fund (RF-1002) of the

Ministry of the Environment, Japan and in part by Grants-in-Aid for

Scientific Research (C) (22510005) and Grant-in-Aid for Young

Scientists (B) (22710072) from the Ministry of Education, Culture,

Sports, Science, and Technology of Japan.

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