Abstract
We investigated ammonia-oxidizing bacteria in activated sludge collected from 12 sewage treatment systems, whose ammonia
removal and treatment processes diﬀered, during three diﬀerent seasons. We used real-time PCR quantiﬁcation to reveal total bacterial numbers and total ammonia oxidizer numbers, and used speciﬁc PCR followed by denaturing gel gradient electrophoresis,
cloning, and sequencing of 16S rRNA genes to analyze ammonia-oxidizing bacterial communities. Total bacterial numbers and total
ammonia oxidizer numbers were in the range of 1.6 · 1012–2.4 · 1013 and 1.0 · 109–9.2 · 1010 cells lÀ1, respectively. Seasonal variation was observed in the total ammonia oxidizer numbers, but not in the ammonia-oxidizing bacterial communities. Members of
the Nitrosomonas oligotropha cluster were found in all samples, and most sequences within this cluster grouped within two of the
four sequence types identiﬁed. Members of the clusters of Nitrosomonas europaea–Nitrosococcus mobilis, Nitrosomonas cryotolerans,
and unknown Nitrosomonas, occurred solely in one anaerobic/anoxic/aerobic (A2O) system. Members of the Nitrosomonas communis cluster occurred almost exclusively in association with A2O and anaerobic/aerobic systems. Solid residence time mainly inﬂuenced the total numbers of ammonia-oxidizing bacteria, whereas dissolved oxygen concentration primarily aﬀected the
ammonia-oxidizing activity per ammonia oxidizer cell.
Ó 2005 Published by Elsevier B.V. on behalf of the Federation of European Microbiological Societies.
Keywords: Activated sludge; Ammonia-oxidizing bacteria; Real-time PCR; Sewage treatment; 16S rRNA gene sequences

1. Introduction
Nitriﬁcation, the two-step process by which ammonia
is oxidized to nitrate via nitrite, plays a key role in the
biological removal of nitrogen in wastewater treatment
systems. The process involves two phylogenetically
unrelated groups of obligately chemolithotrophic bacteria. Ammonia-oxidizing bacteria (AOB) ﬁrst oxidize
ammonia to nitrite, and subsequently nitrite-oxidizing
*

Corresponding author. Tel.: +81 3 5841 7784; fax: +81 3 5841
8538.
E-mail address: tawan@env.t.u-tokyo.ac.jp (T. Limpiyakorn).

bacteria oxidize nitrite to nitrate by. Because of the slow
growth rate of AOB, their high sensitivity to many environmental factors, and their inability to outcompete heterotrophs, ammonia oxidation is a rate-limiting step of
nitrogen removal in wastewater treatment systems [1].
For this reason, a better understanding of the ecology
and microbiology of ammonia-oxidizing bacteria in
wastewater treatment systems is necessary to enhance
treatment performance and control.
Recently developed molecular tools include sequence
analysis of the 16S rRNA and amoA genes to reveal
AOB populations and communities in various environments. In combination with clone libraries or denaturing

0168-6496/$22.00 Ó 2005 Published by Elsevier B.V. on behalf of the Federation of European Microbiological Societies.
doi:10.1016/j.femsec.2005.03.017

206

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

gradient gel electrophoresis (DGGE), the application of
speciﬁc PCR ampliﬁcation [2–4] provides clariﬁcation of
the ammonia oxidizing community in detail. Implementation of ﬂuorescence in situ hybridization (FISH) [1,5–
9] makes it possible to analyze complex communities of
ammonia-oxidizing bacteria and estimate their numbers.
More recently, PCR-based quantiﬁcation techniques allow precise enumeration of AOB populations in the
environments [10–12].
The distribution patterns of distinct AOB species in
the environments reﬂect the physiological properties of
AOB isolates observed in the laboratory [13]. Among
these, ammonia seems to be the most important factor
for the inclusion of distinct AOB species, while other
factors such as salinity are also reported to inﬂuence
their appearance in the environments [6,13–15]. In
general, members of the Nitrosospira spp. or/and the
Nitrosomonas oligotropha clusters are the dominant
ammonia-oxidizing bacteria in the environments low in
ammonia, whereas members of the Nitrosomonas europaea–Nitrosococcus mobilis cluster comprise the majority
of AOB in the environments that are rich in ammonia.
Although a number of studies have assessed the ecology and microbiology of ammonia-oxidizing bacteria in
wastewater treatment systems [1,5–9,11,12,15–18], our
understanding of those in the sewage treatment systems
is still uncertain. The eﬀects of other factors than ammonia on individual ammonia oxidizer species in sewage
treatment systems are poorly understood. AOB found
in these low-ammonia systems are often cited as being
the same bacteria. Some individual species are overlooked by representing their characters by those of the
only few common members of the groups. However,
their characteristics may diﬀer, and they may be inﬂuenced by distinct factors in the systems.
Because of these concerns, we investigated ammoniaoxidizing bacteria in activated sludge collected from 12
sewage treatment systems. We used real-time PCR quantiﬁcation to reveal total bacterial numbers and total
AOB numbers and we used speciﬁc PCR ampliﬁcation
followed by DGGE, cloning, and sequencing of 16S
rRNA genes to identify members of AOB communities.
We focused on the eﬀects of inﬂuent characteristics,
treatment processes, system operation, and seasonal variation on the total AOB numbers and AOB communities.

2. Materials and methods
2.1. Samples of sewage activated sludge and description of
sewage treatment systems
Activated sludge samples were collected from the aeration tanks of 12 sewage treatment systems. These systems are in use in eight sewage treatment plants in
Tokyo, which are run by the Bureau of Sewerage, Tokyo

Metropolitan Government, Japan. The 12 systems differed in ammonia removal and were operated with different treatment processes: anaerobic/anoxic/aerobic
(A2O); anaerobic/aerobic (AO); and conventional activated sludge (AS) processes. Samples were collected
from the 12 systems during three diﬀerent seasons: summer (August 2001); autumn (November 2001); and winter (February 2002). Mixed-liquor suspended solids
(MLSS) concentrations were determined on the day of
sampling. The sludge from approximately 2 mg of
MLSS was transferred into a 1.7-ml Eppendorf tube
and centrifuged at 14,000g for 10 min. The supernatant
was removed, and the pellet was kept at À20 °C until
analysis.
Details of the treatment processes, inﬂuent and eﬄuent characteristics, removal eﬃciencies, and operational
parameters of the 12 systems are listed in Table 1. Systems B1, B2, and B3, systems F1 and F2, and systems
G1 and G2 were located in plants B, F, and G, respectively. Plant B received sewage from a single sewer line,
and the sewage was split among systems B1, B2, and B3
for treatment. In contrast, multiple sewer lines entered
plants F and G; as a result, the various systems in both
plants received diﬀerent sewage. However, the characteristics of the inﬂuents were expected to be similar because the areas from which the sewages were collected
were near each other. The treatment processes of the
12 systems varied: systems A and B1 are A2O processes;
systems B2, C, D, and E are AO processes; and systems
B3, F1, F2, G1, G2, and H are AS processes.
Biological oxygen demand (BOD) in the inﬂuents
ranged from 34 to 141 mg lÀ1, while ammonium concentrations were between 12 and 30 mg N lÀ1. The characteristics of the inﬂuents did not vary notably among
the systems, except for system A. This system was associated with inﬂuent ammonia concentrations of
26–30 mg N lÀ1 and chloride concentrations that were
double those of other systems. These diﬀerences arose
because system A received sewage mostly from commercial areas without rainwater, whereas the other systems
served household areas and received combined sewage.
In addition, the location of system A, an artiﬁcial island
in the sea, might tend to increase the chloride concentration in the inﬂuent of this system.
BOD removal eﬃciencies were excellent (P95%) in
all systems; however, ammonia removal eﬃciencies differed among them. Completed ammonia removal was
achieved in systems A, B1, B2, D, F2, and G1. Ammonia removal was poor in systems G2 and H, possibly because of insuﬃcient oxygen. Ammonia concentrations in
the eﬄuents varied according to the diﬀerence of ammonia removal among the systems. Nitrite concentrations
in the eﬄuents were less than 2 mg N lÀ1 and pH were
maintained between 6.2 and 7.4 in all systems.
Temperature in the 12 systems ranged from 14 to
22 °C in winter to 27 to 31 °C in summer. No marked

Parameters

System
A

Treatment process
Aeration tank volume
Volumetric ﬂow to aeration tank
BOD in inﬂuenta
NH4-N in inﬂuenta
ClÀin inﬂuenta
NH4-N in eﬄuenta
NO2-N in eﬄuenta
BOD removala
NHþ -N removala
4
Volumetric NHþ -N removala
4
HRTb
b
SRT
MLSSb
MLDOb

Unit

B1

B2

B3

C

D

E

F1

F2

G1

G2

H

m3
m3 mÀ3 dÀ1
mg lÀ1
mg lÀ1
mg lÀ1
mg lÀ1
mg lÀ1
%
%
g mÀ3 dÀ1
h
Day
mg lÀ1
mg lÀ1

A2O
6400
1.8–2.1
91–139
26–30
140–180
0
0
98–99
100
42–44
24.9
16.5
2122
4.1

A2O
7450
2.8–3.4
53–141
13–16
54–80
0
0
98–99
98–100
33–34
16.1
14.2
1629
1.8

AO
11,175
2.3–2.8
53–141
13–16
54–80
0–1
0
97–99
96–99
26–30
12.7
8.6
1422
2.2

AS
44,700
2.4–3.0
53–141
13–16
54–80
0–2
0–1
97–99
87–97
24–31
9.2
7.2
1256
2.3

AO
10,230
2.2–4.6
80–133
20–24
63–80
3–12
1
95–96
40–90
18–50
7.3
4.6
1085
3.4

AO
11,760
2.2–2.5
64–132
14–18
40–58
0
0
99–100
100
28–33
14.0
12.6
1249
5.3

AO
31,050
3.3
65–100
12–18
80–90
1–2
0
98–99
85–97
27–44
9.8
7.9
1844
5.8

AS
15,520
2.6–3.0
67–90
15–18
–c
0–6
0–1
96–99
69–98
32–34
8.5
6.3
1339
6.5

AS
37,840
4.1–4.9
63–111
15–18
–c
0–1
0
97–98
97–100
54–60
5.4
11.7
1698
6.7

AS
97,200
3.5–4.1
74–103
18–19
89–94
0
0
97–99
100
49–60
6.3
5.7
1159
6.5

AS
268,800
3.0–3.8
83–94
18–21
89–98
12–19
0–2
95–97
0–40
0–13
7.3
4.5
1009
1.8

AS
202,320
2.6–2.9
34–73
14–17
52–72
4–16
0–1
95–97
7–89
0–20
9.0
7.1
1088
1.6

HRT, hydraulic retention time; SRT, solid residence time; MLSS, mixed liquor suspended solid; MLDO, mixed liquor dissolved oxygen.
a
All inﬂuent characteristic, eﬄuent characteristic, and removal eﬃciency values were analyzed from one-day grab samples collected on the day close to the day of sludge collection.
b
All operational parameters were the averages of the three months in which sludge was collected.
c
Data not available.

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

Table 1
Treatment processes, inﬂuent- and eﬄuent characteristics, removal eﬃciencies, and operational parameters of 12 sewage treatment systems; information provided by Tokyo Metropolitan
Government

207

208

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

seasonal variations in inﬂuent characteristics, removal
eﬃciencies, or operational parameters were observed in
all systems throughout this study. However, solid residence time (SRT) and dissolved oxygen (DO) concentration in system H varied during the studied seasons, and
these variations will be discussed later.
2.2. DNA extraction
DNA was extracted directly from samples using FastDNA SPIN kits for soil (Bio 101, Vista, CA, USA) with
a small modiﬁcation at the initial step: 1 ml of sodium
phosphate buﬀer solution was added to and mixed with
the sample, and then the tube was sonicated for 30 s on
ice. The remaining steps followed the manufacturerÕs
instructions. The product from DNA extraction was
veriﬁed by electrophoresis in 1% agarose (TaKaRa
LO3, Tokyo, Japan).
To minimize the variation in DNA extraction, the
templates used for real-time PCR quantiﬁcation were
prepared from the mixture of DNA, which was extracted in triplicate for a sample.
2.3. Real-time PCR quantiﬁcation of total bacteria and
total AOB
Real-time PCR quantiﬁcation of 16S rRNA genes of
total bacteria was performed using the primers 1055f
and 1392r and the Taq Man probe 16Staq1115 as previously described [12]. The oligonucleotide sequences of
the primers and the Taq Man probe are shown in Table
2. The PCR mixture was prepared in a total volume of
25 ll using the Taq Man Universal PCR Master Mix
kit (PE Applied Biosystems), 15 pmol of forward primer
1055f, 15 pmol of reverse primer 1392r, 6.25 pmol of
Taq Man probe 16Staq1115, and standard DNA or extracted DNA from samples. The standard DNA, which
was the pT7Blue T-Vector (Novagen, Darmstadt, Germany) possessing 16S rRNA gene of N. europaea, was
prepared ranging from 5 · 101 to 5 · 107 copies. The extracted DNA from a sample was prepared for three dif-

ferent 10-fold dilutions, and each of the dilutions was
real-time PCR quantiﬁed in duplicate. PCR ampliﬁcation was performed in an ABI Prism SDS 7000 instrument (PE Applied Biosystems) under conditions of
3 min at 50 °C and 10 min at 95 °C followed by 45 cycles
of 30 s at 95 °C, 1 min at 50 °C, and 20 s at 72 °C. To
calculate the cell numbers of bacteria from the quantiﬁed numbers of 16S rRNA gene, it is assumed that the
average 16S rRNA gene copies per genome of bacterial
cell is 3.6 based on the average 16S rRNA gene copies
found in cultured bacteria [19].
Real-time PCR quantiﬁcation of 16S rRNA genes of
total AOB was performed using the primers CTO 189f
and RT1r and the Taq Man probe TMP1 as described
earlier [10]. The oligonucleotide sequences of the primers
and the Taq Man probe are shown in Table 2. The PCR
mixture was prepared in a total volume of 25 ll using
the Taq Man Universal PCR Master Mix kit, 7.5 pmol
of a 2:1 ratio of forward primers CTO 189fA/B and
CTO 189fC, 7.5 pmol of reverse primer RT1r,
3.125 pmol of Taq Man probe TMP1, and standard
DNA or extracted DNA from samples. The standard
DNA, which was the pT7Blue T-Vector possessing 16S
rRNA gene of N. europaea, was prepared in a range of
4 · 101 to 4 · 107 copies. The extracted DNA from a
sample was prepared for three diﬀerent 10-fold dilutions, and each of the dilutions was real-time PCR quantiﬁed in duplicate. PCR ampliﬁcation was performed in
an ABI Prism SDS 7000 instrument under conditions of
2 min at 50 °C and 10 min at 95 °C followed by 40 cycles
of 30 s at 95 °C and 1 min at 60 °C. To calculate the cell
numbers of ammonia-oxidizing bacteria from the quantiﬁed numbers of 16S rRNA genes, it is assumed that
AOB possess one copy of 16S rRNA gene per genome
[20].
2.4. PCR, DGGE, cloning, sequencing, and phylogenetic
analysis
We used the primers CTO189f and CTO654r [2] to
amplify 465-bp fragments of 16S rRNA genes of AOB

Table 2
Primers and probes used in this study
Primer or probe

Nucleotide sequence (5 0 –3 0 )

Reference

Primers
1055f
1392r
CTO 189A/Bf
CTO 189Cf
RT1r
CTO 654r

ATGGCTGTCGTCAGCT
ACGGGCGGTGTGTAC
GGAGRAAAGCAGGGGATCG
GGAGGAAAGTAGGGGATCG
CGTCCTCTCAGACCARCTACTG
CTAGCYTTGTAGTTTCAAACGC

[12]
[12]
[2]
[2]
[10]
[2]

Probesa
16STaq1115 (5 0 -FAM and 3 0 -TAMRA)
TMP1 (5 0 -FAM and 3 0 -TAMRA)

CAACGAGCGCAACCC
CAACTAGCTAATCAGRCATCRGCCGCTC

[12]
[10]

a

FAM, 6-carboxyﬂuorescein; TAMRA, 6-carboxy-tetramethylrhodamine.

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

structed using the ARB program package. We added
our 397-bp sequences into the distance tree, which was
previously constructed based on comparison of 1000bp sequences of all AOB, which are available in the
SSU rRNA database, and some related non-AOB,
which were used as outgroup sequences. Additionally,
our 397-bp sequences and 397-bp sequences of described
AOB species [22] and some related non-AOB were calculated based on maximum parsimony, maximum likelihood, and distance analyses using the external
software provided in the ARB program package (Phylip
DNAPARS, AxML, and Phylip Distance Method,
respectively).
2.5. Accession numbers for nucleotide sequences
The partial sequences of 16S rRNA genes obtained in
this study were submitted to the DDBJ database under
accession numbers AB176856–AB176884.

3. Results
3.1. Total bacterial numbers and total AOB numbers
We used real-time PCR quantiﬁcation to reveal total
bacterial numbers and total AOB numbers in samples of
sewage activated sludge. Fig. 1 shows total bacterial
numbers in the aeration tanks of the 12 systems. Total
bacterial numbers were in a range of 1.6 · 1012–
2.4 · 1013 cells lÀ1. Fig. 2 shows total AOB numbers,
percentage of total AOB in total bacterial populations,
and ammonia-oxidizing activities per cell of AOB in
the aeration tanks of the 12 systems. Total AOB numbers were between 1.0 · 109 and 9.2 · 1010 cells lÀ1,
and can be accounted for 0.01–2.8% of total bacterial
populations. Ammonia-oxidizing activities per cell were
calculated from the total AOB numbers and volumetric
ammonia removal in the aeration tanks. The ammoniaoxidizing activities per cell were ranging from 0 to
49.6 fmol cellÀ1 hÀ1.
10
1.0E+114
Total bacteria numbers
Cells/ l
(cells l-1)

belonging to Betaproteobacteria. The oligonucleotide
sequences of the primers are shown in Table 2. The extracted DNA was PCR-ampliﬁed using the primer set
(the forward primer had a GC clamp) for 35 cycles in
a 50-ll reaction volume. DNA eluted from bands excised from DGGE gels and colonies picked directly after
cloning were ampliﬁed for 20–25 cycles using the primer
set lacking the GC clamp in a 50-ll reaction volume.
The PCR mixture was prepared using AmpliTaq Gold
DNA polymerase (PE Applied Biosystems, CA, USA)
following the manufacturerÕs instructions with 1 pmol
of each primer. PCR ampliﬁcation was performed in a
T3 thermocycler (Biometra, Gottingen, Germany) under
the conditions of 5 min at 95 °C followed by 35 cycles of
30 s at 92 °C, 30 s at 57 °C, and 45 s at 72 °C (+1 s each
cycle), followed by 5-min ﬁnal extension at 72 °C.
DGGE was performed according to the modiﬁcation
of a described method [21]. We used 8% polyacrylamide
gels, and the urea–formamide denaturant gradient was
35–50%. Gels were run on the D Code system (BioRad Laboratories, Hercules, CA, USA) for 16 h at
60 °C and 75 V. After electrophoresis, the gels were
stained with Vistra Green (Amersham Pharmacia Biotech, Tokyo, Japan) and visualized with a ﬂuorescent
image scanner (Fluorimager 595, Molecular Dynamics,
Sunnyvale, CA, USA). Prominent bands were excised
and dissolved in 30 ll sterilized water. DNA was recovered from the gel by freeze–thawing three times.
Each target fragment of DNA recovered from the
DGGE gel was puriﬁed by cloning with pT7Blue T-Vector and DH5-a competent cells (Toyobo, Tokyo, Japan)
according to the manufacturerÕs instructions. Positive
clone colonies were ampliﬁed directly using the primer
set with the additional GC-clamp, and the products
again were subjected to DGGE to check their migration.
The target DNA fragments were then excised and reampliﬁed. Before sequencing, small DNA fragments and
excess primers were removed from the PCR products
by Microcon spin columns (Millipore, Tokyo, Japan).
The puriﬁcation procedure followed the manufacturerÕs
instructions.
Sequencing reactions were run according to the
manufacturerÕs instructions with 20 ng of the PCR
product, the ABI Big Dye Terminator kit version 3.1
(PE Applied Biosystems), and the CTO primer set.
After excess primers and dye terminators from the
products of sequencing reaction were removed using
Centri-sep spin columns (PE Applied Biosystems), the
products were analyzed in an ABI 310 DNA sequencer
(PE Applied Biosystems).
The completed 397-bp from 465-bp analyzed sequences were aligned with sequences from the SSU
rRNA database (Antwerm, Belgium) using the ARB
program package (Department of Microbiology, Technische Universitat Munchen, Munich, Germany;
[http://www.arb-home.de]). Phylogenetic tree was con-

209

10
1.0E+113

12

10
1.0E+1

10
1.0E+111
A B1 B2 B3 C

D E F1 F2 G1 G2
System

H

Fig. 1. Total bacterial numbers in aeration tanks of 12 sewage
treatment systems: s, summer; n, autumn; n, winter.

210

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

(a)

3.2. Analysis of samples of sewage activated sludge by
speciﬁc PCR ampliﬁcation followed by DGGE, cloning,
and sequencing of 16S rRNA genes

12

Total AOB numbers
Celell/sll-1)
(c ls

10
1.0E+12

1011
1010
9

10
1.0E+09
8

10
1.0E+08

D

4

o

5

% Total AOB in total

D E F1 F2 G1 G2 H
System

A B1 B2 B3 C

(b)

D E F1 F2 G1 G2 H
System

A B1 B2 B3 C

t o tbacterOB ipotpulation t e i a
al A ial n o t al b ac
p o p u l at i o n

A B1 B2 B3 C

3
2
1
0

Ammonia-oxidizing
activifms l(/ mel lceh r hr–1)
tie o f c ol / ll–1

(c)

100

10

1
E

F1 F2 G1 G2 H

0
System
Fig. 2. Total AOB numbers (a), proportion of total AOB within total
bacterial populations (b), and ammonia-oxidizing activities per AOB
cell (c) in aeration tanks of 12 sewage treatment systems: s, summer;
n, autumn; n, winter.

We used speciﬁc PCR ampliﬁcation followed by
DGGE, cloning, and sequencing of 16S rRNA genes
to reveal the structure of AOB communities in the samples of sewage activated sludge. Analysis of duplicate
samples by DGGE showed that the band patterns obtained were reproducible (data not shown). Fig. 3 shows
the image of a DGGE gel of the PCR-ampliﬁed products of samples taken from the 12 systems in autumn
(November 2001). On the right side of the ﬁgure, the
locations of the bands excised are labelled with the
names of the AOB clusters identiﬁed. Because the locations of the bands excised diﬀered for system A, which
led to diﬀerent sequences and thus identiﬁcation of different AOB clusters, the names of the AOB clusters
are labelled separately on the left side of the ﬁgure. After
the PCR-ampliﬁed products of all samples had been
electrophoresed in DGGE gels, a total of 81 representative bands were selected for sequencing. To prevent the
confusion with the bands selected directly from the original gel images, all bands were excised from the gels,
reampliﬁed, and analyzed on the new gels to clarify an
individual band position before selection for sequencing.
We constructed phylogenetic trees using several treeing methods. Fig. 4 shows a phylogenetic tree based on
16S rRNA genes of AOB belonging to Betaproteobacteria. All the tree methods resulted in the same grouping of
AOB sequences analyzed. In addition, the grouping of
AOB clusters in the trees remained the same except for
the ‘‘unknown-Nitrosomonas’’ cluster, the positions of
which varied with the method used. The unknown-Nitrosomonas cluster can be included into N. oligotropha
cluster or can be recognized as another independent cluster within Nitrosomonas spp. The sequences of many of
the bands analyzed reveal their relation to the N. oligotropha cluster (Nitrosomonas cluster 6a). Because many

Fig. 3. DGGE images of PCR-ampliﬁed products of samples taken from 12 sewage treatment systems in autumn (November 2001). Marks and
numbers indicate excised bands from which sequences were determined. AOB genus abbreviations are N. for Nitrosomonas and Nc. for
Nitrosococcus.

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

211

Fig. 4. Phylogenetic tree showing 16S rRNA genes of AOB belonging to Betaproteobacteria with addition of 397-bp sequences from our study into
the distance tree that was previously based on comparison of 1000-bp sequences of described AOB [22] using the ARB program package. AOB genus
abbreviations are N. for Nitrosomonas, Nc. for Nitrosococcus, and Ns. for Nitrosospira. Codes of the DGGE bands of this study are shown in bold;
the ﬁrst character indicates the sample name, the second character denotes the season (S, summer, A, autumn, W, winter, respectively), followed by
the band number. AOB clusters are depicted on the right side of the tree.

sequences that are aﬃliated to this cluster are distributed
in various environments and their properties somewhat
diﬀer, for the purpose of further discussion, the sequences identiﬁed as associated with this cluster were
allocated into four sequence types based on the grouping
of the AOB sequences with several treeing methods. In
the previous study [14], members of the N. oligotropha
cluster that originated from enriched brackish and freshwater were separated into three sequence types. Because
the sequences found in this study diﬀered from those of

the previous one with respect to their source, the classiﬁcation system of the previous study is not applicable for
this one. In our classiﬁcation scheme, N. ureae (another
representative of cluster 6a) are not associated with a
particular sequence type because they were not closely related to any other sequence found in this study. The ability of the primers CTO189f and CTO654r to amplify all
clusters of AOB belonging to Betaproteobacteria was
conﬁrmed when members of all AOB clusters, except
those of Nitrosomonas marina and the N. sp. Nm143,

212

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

Fig. 5. DGGE images of the PCR-ampliﬁed products of samples taken from sewage treatment systems during three seasons to illustrate the seasonal
variation in the AOB communities: (a) system A, (b) system B1, (c) system C, (d) system E, (e) system F2, and (f) system G1. The character above
each lane (S, A, or W) stands for season (summer, autumn, or winter respectively). Marks and numbers indicate the excised bands from which
sequences were determined.

were recovered from the samples. Most of the AOB sequences analyzed showed 96–99% identity to the AOB
sequences available in the database. However, a few sequences, which were closely related to N. communis, all
sequences of the unknown-Nitrosomonas cluster, and
some type 6a-3 sequences of the N. oligotropha cluster,
showed slightly lower homology (95%).
3.3. Eﬀect of seasonal variation on AOB communities
PCR-ampliﬁed products of samples taken during different seasons were analyzed on the same denaturing gel
to highlight the eﬀects of seasonal changes on the communities of ammonia-oxidizing bacteria (Fig. 5). Most
systems (except B3, E, F2, and G2) exhibited similar
band patterns for the three seasons. Moreover, the sequence analysis showed that the AOB communities of
most systems were nearly identical regardless of the season (Table 3); there were no seasonal variations in the
common AOB, and only particular AOB exhibited the
seasonal variation in some systems.
3.4. Distribution of AOB in sewage treatment systems
Table 3 lists and summarizes the related AOB sequences found in each sample. Most of the bands analyzed were related to Nitrosomonas spp. rather than
Nitrosospira spp. The highest number of bands analyzed grouped within the N. oligotropha cluster. Members of this cluster were found in all samples. Within
this cluster, sequence types 6a-1 and 6a-4 were found
in most samples. Only system A contained members
of the N. europaea–Nc. mobilis cluster and the unknown-Nitrosomonas cluster. Members of the N. cryotolerans cluster were present almost solely in system A
and appeared only during some seasons in systems D,
F2, and G1. Further, systems A, B1, B2, B3, C, D,
and E contained members of the N. communis cluster;
interestingly, all of these systems (except B3) were
operated as either A2O or AO processes. It is not possible to discern a trend in the relationship between the

sewage activated sludge system and the presence of
the Nitrosospira cluster, which was recovered from
only one sample.

4. Discussion
The total bacterial numbers in a range of 1.6 · 1012–
2.4 · 1013 cells lÀ1 in the 12 systems (Fig. 1) fairly corresponded to the numbers mentioned earlier in a municipal wastewater treatment system: 4.55 · 1012–
1.04 · 1013 cells lÀ1 by dot blot hybridization [11];
4.3 ± 2.0 · 1011 cells lÀ1 by real-time PCR quantiﬁcation based on 16S rRNA genes of total bacteria (the
primers 1055f and 1392r and the Taq Man probe
16STaq1115) [12].
The total AOB numbers between 1.0 · 109 and
9.2 · 1010 cells lÀ1 in the 12 systems can be accounted
for 0.01–2.8% of total bacterial populations (Fig. 2(a)
and (b)). Checking the primer and probe matches shows
that the reverse primer RT1r and the Taq Man probe
TMP1 completely target most AOB sequences analyzed.
This conﬁrms that most AOB sequences analyzed were
included in the real-time PCR quantiﬁcation using this
primer set. The total numbers of ammonia-oxidizing
bacteria in the 12 systems were compatible to the numbers reported previously in a municipal wastewater
treatment system: 1.2 ± 0.9 · 1010 cells lÀ1 (2.9% of total
bacterial populations) by real-time PCR quantiﬁcation
based on 16S rRNA genes of total AOB (the primers
CTO 189f and RT1r and the Taq Man probe TMP1)
and 7.5 ± 6.0 · 109 cells lÀ1 (1.7% of total bacterial populations) by real-time PCR quantiﬁcation based on
amoA genes of the N. oligotropha [12]. The total numbers of ammonia-oxidizing bacteria varied among the
seasons studied; the largest population size occurred
during winter, followed by autumn and summer
(Fig. 2(a)). However, a proportion of total AOB within
the total bacterial populations increased (0.5–3%) during autumn, whereas in summer and winter it remained
the same (0.01–1.1%) (Fig. 2(b)). The increased

+
+
+

+
+

S

+
+
+
+

+
+

+

+

+

+

+
+
+
+
+

+

+

+
+
+
+
+
+

+

+

+
+
+
+
+
+
+
+

S, summer; A, autumn; W, winter.

+
+
+
+
+
+
+
+
+
+
+

+

+

+
+
+
+

+
+
+
+

+

+

+

+

+

+

+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+
+

+
+

+

+

S
W
A
S
S
A

W
S

A

W

S

A

W

S

A

W

S

A

W

S

A

W

E
D
C
B3
B2
B1
A

System
AOB cluster

Table 3
Summary of related sequences of ammonia-oxidizing bacteria in samples of 12 sewage treatment systems

Nitrosospira sp.
Nitrosomonas communis
Nitrosomonas europaea–
Nitrosococus mobilis
Nitrosomonas cryotolerans
Unknown Nitrosomonas
Nitrosomonas oligotropha,
sequence type 6a-1
Nitrosomonas oligotropha,
sequence type 6a-2
Nitrosomonas oligotropha,
sequence type 6a-3
Nitrosomonas oligotropha,
sequence type 6a-4

F1

A

+

W

S

+

+

A

+

+

+

+
+
+

+

A
S
W

G1
F2

+

+
+

+

S
W

G2

A

W

H

A

+

W

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

213

abundance during the autumn was possibly caused by
the decrease in the ratio between BOD and NH4-N concentrations in the inﬂuents during this season (data not
shown).
The ammonia-oxidizing activities per AOB cell ranging from 0 to 49.6 fmol cellÀ1 hÀ1 in the 12 systems
(Fig. 2(c)) were consistent with the numbers reported before: 7.7 and 12.4 fmol cellÀ1 hÀ1 in a municipal wastewater treatment system estimated from real-time PCR
quantiﬁcation of 16S rRNA genes of total AOB (primers CTO 189f and RT1r and the Taq Man probe
TMP1) and amoA genes of N. oligotropha, respectively
[12]; and 9–23 fmol cellÀ1 hÀ1 for AOB isolates [23].
The ammonia-oxidizing activities per cell varied during
the seasons studied. The highest ammonia-oxidizing
activities per cell occurred during summer, followed by
those during autumn and winter. Moreover, the total
AOB numbers and the ammonia-oxidizing activities
per AOB cell depended largely on temperature variation
and altered with every temperature level: 27–31 °C during summer, 19–26 °C during autumn, and 14–22 °C
during winter.
Although the total numbers of ammonia-oxidizing
bacteria ﬂuctuated over seasons, no marked seasonal
variation in the composition of AOB communities was
observed. The AOB communities were fairly stable in
all systems during the six months of study. No seasonal
variations in the common ammonia-oxidizing bacteria
were observed. Only particular ammonia oxidizers
exhibited seasonal variation in some systems. We therefore state that AOB communities change little over temperatures ranging from 14 to 31 °C.
Sequence analysis of 16S rRNA genes revealed that
most of the electrophoretic bands were closely related
to Nitrosomonas spp., not to Nitrosospira spp. This ﬁnding corresponds to several previous studies [5–
9,15,16,24], but is in contrast to few studies in which
mainly Nitrosospira spp. were detected [25,26]. Nonetheless, based on the amoA approach, it is notable that
Nitrosospira species are not important ammonia-oxidizing bacteria in full-scale wastewater treatment systems
analyzed so far [27].
We recovered members of the N. oligotropha cluster
from all samples, and they comprised the majority of
the bands analyzed. Although the forward primer
CTO189f has a single mismatch with some members of
the N. oligotropha cluster, the sequences closely related
to this cluster could easily be recovered from our samples. Sequences of the N. oligotropha cluster are often
recovered from oligotrophic environments, including
freshwater sediment [28], wastewater treatment systems
receiving low-ammonia inﬂuents [9,11,12,16], and
chloraminated drinking water distribution systems
[29,30]. The study of 36 isolates of the N. oligotropha
cluster, most of which were from oligotrophic freshwater, revealed very low Ks values, ranging from 1.9 to

214

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

4.2 lM free ammonia [13]. Ammonium concentrations
in the inﬂuents of every system in this study were
<2 mM (28 mg N lÀ1) (Table 1), and in the systems
where nitriﬁcation was achieved, ammonium concentrations in the aeration tanks were under the detection limit. Thus, volumetric ammonia removals could be
estimated as less than 60 g mÀ3 dÀ1 in all systems. This
low ammonia level may be the key factor promoting
the predominance of the N. oligotropha cluster in most
systems in this study.
Sequence types 6a-1 and 6a-4 from the N. oligotropha
cluster were distributed in most systems in most seasons
studied. Sequence group 6a-1 is comprised of the sequences from continuous culture enrichment at
growth-limiting ammonia concentration (5 lM; bands
Ch4E2 and Ch6E3) [28], whereas sequence group 6a-4
is comprised of the sequences recovered from an aeration tank of a laboratory-scale single sludge nitriﬁcation–denitriﬁcation systems receiving a 0.8-mM
ammonia in the inﬂuent (clones DNB_Y20 and
DNB_Y39) [16], fresh water sediment (clone 6a3), and
drinking water reservoirs (clone HBO-2). In the past,
ecological and physiological diﬀerences among ammonia-oxidizing bacteria within the N. oligotropha cluster
have been overlooked because of representing their
characteristics with only one or two common members
of the group. Recently, the characteristics of AOB within
this cluster were suggested to somewhat diﬀer depending
on sequence type [14]. As our study shows, only two sequence types within this cluster were widely and abundantly distributed in most systems. Factors such as
ammonia aﬃnity, oxygen aﬃnity, sensitivity to salt
and/or nitrite may diﬀer among sequence types. For
example, salinity is known to aﬀect the AOB communities in oligotrophic environments [14]. The eﬀect of each
environmental factor on AOB within this cluster needs
further study to be able to better understand the implications of the presence of various sequence types.
We recovered members of the N. europaea–Nc. mobilis cluster only from system A. The outstanding physiological property of the N. europaea–Nc. mobilis cluster is
the low aﬃnity to free ammonia (Ks > 30 lM) [13], a
characteristic consistent with this clusterÕs preference
for eutrophic environments. Although system A contained higher ammonium concentrations in the inﬂuent
than other systems (Table 1), these ammonia concentrations were not signiﬁcantly diﬀerent from those of other
systems ($2 mM). In addition, ammonium concentrations in the aeration tank of system A were all the time
below the detection limit, and volumetric ammonium
removals of this system were in the same range as those
of other systems. Therefore, ammonia is not responsible
for the appearance of the N. europaea–Nc. mobilis cluster in system A. The study on 27 isolates of N. europaea
and seven isolates of Nc. mobilis revealed halotolerant to
moderately halophilic characteristic of this cluster [13].

Earlier molecular-level investigations supported this by
showing the predominance of N. europaea cluster in
alkaline, high-salinity environments [31] and in an
industrial wastewater treatment systems treating saline
wastewater [6]. The double chloride concentration in
the inﬂuent of system A compared to other systems
may thus explain the presence of the N. europaea–Nc.
mobilis cluster in this system.
Besides those of the N. europaea–Nc. mobilis cluster,
also the members of the N. cryotolerans cluster and of
the unknown-Nitrosomonas cluster were present almost
solely in system A. Assuming the diﬀerence in halotolerance or halosensitivity among distinct AOB species,
the high chloride concentration in the inﬂuent of this
system may have caused its AOB community diﬀer
from those in other systems. N. cryotolerans has been
reported as obligately halophilic bacterium [13]. However, the salt tolerance of the unknown-Nitrosomonas
cluster is unknown because ammonia-oxidizing bacteria aﬃliated to this cluster have not been isolated
and characterized yet. The most closely related AOB
to the unknown-Nitrosomonas cluster is N. sp. JL21
(accession number AB000700) with only 95% similarity
in the partial 16S rRNA gene sequence. Due to the
short sequence length, it is impossible at this moment
to clarify whether the unknown-Nitrosomonas group
belongs to N. oligotropha cluster or represents the
putative novel AOB cluster.
Members of the N. communis cluster were found in
all A2O and AO systems; but they were mostly absent
from AS systems. Thus far, only two 16S rRNA gene
sequences aﬃliated with this cluster (clones GaN50304
and GaN3a) [9] have been detected in the environments. This may have been caused by the fact that
many of the commonly used primers contain mismatches to members of the N. communis cluster [9].
The reverse primer CTO654r does not target all members of the N. communis cluster recognized so far.
However, using this primer set, many members of
the N. communis cluster could be recovered from
our samples. This diﬀerence is possibly because the sequences of our N. communis at the positions corresponding to the primer region match to the reverse
primer and thus are diﬀerent from those of all recognized AOB. The members of the N. communis cluster
obtained from our samples were closely related to
those recovered earlier (clones GaN50304 and GaN3a)
from the top 100 lm of the bioﬁlm surface of a
sequencing batch reactor, which operated with diﬀerent phases of aeration and non-aeration for the purposes of nitrogen and phosphate removal. Similarly
as this earlier study, we found that the members
of the N. communis cluster were present almost
exclusively in A2O and AO systems, which were operated with diﬀerent phases of aeration and nonaeration. These AOB may thus have an advantage

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

in environments with ﬂuctuating oxygen levels or during the absence of oxygen. These results reﬂect the
distinct ability of ammonia-oxidizing bacteria to survive or even prosper under special conditions. Along
this line, oscillation of oxygen concentrations may be
another factor supporting growth or survival of particular AOB.
Total AOB numbers and ammonia-oxidizing activities per AOB cell are discussed to clarify the reasons
underlying the reduced ammonia removal in some systems (G1, G2, and H). Various aspects of system operation were found to aﬀect the total AOB numbers and
the ammonia-oxidizing activities per cell of AOB.
G1 and G2 are AS systems that received similarstrength wastewater with the same volumetric ﬂows (Table 1). Both systems were operated with the similar SRT,
but the amounts of airﬂow to wastewater were diﬀerent
(data not shown). The airﬂow to wastewater was suﬃcient in system G1, but not in system G2. As a result,
DO concentrations in the aeration tank of system G1
were in a range of 5–8 mg lÀ1, whereas those of system
G2 were around 2 mg lÀ1 (Table 4). Thus, only system
G1, but not system G2, achieved ammonia removal.
Although ammonia removal in both systems was diﬀerent, this value did not reﬂect the total AOB numbers; in
contrast, it did reﬂect the ammonia-oxidizing activities.
This result suggested that oxygen supply and DO con-

215

centrations did not inﬂuence the total numbers of
ammonia-oxidizing bacteria. On the other hand, they
did aﬀect the ammonia-oxidizing activities per cell.
In system H, ammonia removal eﬃciencies and volumetric ammonia removal ﬂuctuated during the seasons studied (Table 5). Nitriﬁcation was achieved
only during autumn, not during summer or winter.
Although no change in the inﬂuent characteristics
was noted, operation of the system varied during the
season studied. Proper operation was provided only
during the autumn, whereas insuﬃcient oxygen was
supplied during the summer, and SRT was too short
during the winter. The proper operation led to the
highest total numbers of ammonia-oxidizing bacteria
during the autumn. Longer SRT in summer seemingly
supported higher total AOB numbers during this season. However, neither SRT nor total AOB numbers related to the volumetric ammonia removal and the
ammonia-oxidizing activities per cell. The ammoniaoxidizing activities were dependent on DO concentrations. We can conclude that SRT mainly inﬂuenced
the total AOB numbers, whereas the DO concentration
primarily aﬀected the ammonia-oxidizing activity per
cell. Indeed, other studies support our observation
about the eﬀect of SRT on the total AOB numbers
[32] and the eﬀect of oxygen concentrations on the
ammonia-oxidizing activity [14,33].

Table 4
Ammonia removal, total AOB numbers, ammonia-oxidizing activities per AOB cell, and DO concentrations in aeration tanks of systems G1 and G2
Parameters

Unit

G1

G2

Summer

Autumn

Winter

Summer

Autumn

Winter

NHþ -N removala
4
Volumetric NHþ -N removala
4
Total AOB number
NHþ -N oxidizing activity
4
MLDOb

%
g mÀ3 dÀ1
cells lÀ1
fmol cellÀ1 hÀ1
mg lÀ1

100
49
2.9 · 109
50
5.2

100
60
2.9 · 1010
6
6.6

100
53
6.7 · 1010
2
7.7

40
13
1.1 · 1010
3
1.6

0
0
2.8 · 1010
0
1.0

11
0
3.7 · 109
0
2.0

MLSS, mixed liquor suspended solid; MLDO, mixed liquor dissolved oxygen.
a
All removal eﬃciency values were analyzed from one-day grab samples collected on the day close to the day of sludge collection.
b
All operational parameters were the averages of the month in which sludge was collected.

Table 5
Ammonia removal, total AOB numbers, ammonia-oxidizing activities per AOB cell, and operational parameters of system H
Parameter

Unit

Season
Summer

NHþ -N
4

a

removal
Volumetric NHþ -N removala
4
Total AOB number
b
MLSS
SRTb
NHþ -N oxidizing activity
4
MLDOb

%
g mÀ3 dÀ1
cells lÀ1
mg lÀ1
Day
fmol cellÀ1 hÀ1
mg lÀ1

Autumn

Winter

7
0
3.4 · 109
1270
8.2
0
0.6

89
20
1.2 · 1010
1125
6.0
4
1.1

30
4
1.0 · 109
870
3.8
12
3.2

MLSS, mixed liquor suspended solid; SRT, solid residence time; MLDO, mixed-liquor dissolved oxygen.
a
All removal eﬃciency values were analyzed from one-day grab samples collected on the day close to the day of sludge collection.
b
All operational parameters were the averages of the month in which sludge was collected.

216

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

5. Conclusions
In this study, we mainly discussed the ammoniaoxidizing bacteria found in activated sludge of sewage
treatment systems. We noted the eﬀects of inﬂuent characteristics, treatment processes, system operation, and
seasonal variation on the total AOB numbers and the
AOB communities. We showed that the ammoniaoxidizing bacteria observed in these low-ammonia systems diﬀer and are inﬂuenced by distinct environmental
factors. However, the discussion primarily reﬂected
qualitative results because accurate numbers of particular AOB were unavailable. Further studies are necessary
to develop quantitative techniques for particular AOB,
especially for those within the N. oligotropha cluster
and the N. communis cluster, and to better clarify the
roles of these ammonia oxidizers in low-ammonia
environments.

Acknowledgements
We are grateful to the Tokyo Metropolitan Government for providing the samples and data from the sewage treatment systems.

References
[1] Wagner, M., Rath, G., Amann, R., Koops, H.P. and Schleifer,
K.H. (1995) In situ identiﬁcation of ammonia-oxidizing bacteria.
Syst. Appl. Microbiol. 18, 251–264.
[2] Kowalchuck, G.A., Stephen, J.R., De Boer, W., Prosser, J.I.,
Embley, T.M. and Woldendorp, J.M. (1997) Analysis of ammonia-oxidizing bacteria of the b subdivision of the class Proteobacteria in costal sand dunes by denaturing gradient gel
electrophoresis and sequencing of PCR-ampliﬁed 16S ribosomal
DNA fragments. Appl. Environ. Microbiol. 63 (4), 1489–1497.
[3] Rotthauwe, J.H., Witzel, K.P. and Liesack, W. (1997) The
ammonia monooxygenase structure gene amoA as a function
marker: molecular ﬁne-scale analysis of natural ammonia-oxidizing populations. Appl. Environ. Microbiol. 63 (12), 4704–4712.
[4] Nicolaisen, M.H. and Ramsing, N.B. (2002) Denaturing gradient
gel electrophoresis (DGGE) approaches to study the diversity of
ammonia-oxidizing bacteria. J. Microbiol. Meth. 50, 189–203.
[5] Mobarry, B.K., Wagner, M., Urbain, V., Rittmann, B.E. and
Stahl, D.A. (1996) Phylogenetic probes for analyzing abundance
and spatial organization of nitrifying bacteria. Appl. Environ.
Microbiol. 62 (6), 2156–2162.
[6] Juretschko, S., Timmermann, G., Schmid, M., Schleifer, K.H.,
Pommerening-Roser, A., Koops, H.P. and Wagner, M. (1998)
Combined molecular and conventional analyses of nitrifying
bacterium diversity in activated sludge: Nitrosococcus mobilis and
Nitrospira-like bacteria as dominant populations. Appl. Environ.
Microbiol. 64 (8), 3042–3051.
[7] Wagner, M., Noguera, D.R., Juretschko, S., Rath, G., Koops,
H.P. and Schleifer, K.H. (1998) Combining ﬂuorescent in situ
hybridization (FISH) with cultivation and mathematical modeling
to study population structure and function of ammonia-oxidizing
bacteria in activated sludge. Water Sci. Technol. 37 (4–5), 441–
449.

[8] Okabe, S., Satoh, H. and Watanabe, Y. (1999) In situ analysis of
nitrifying bioﬁlms as determined by in situ hybridization and the
use of microelectrodes. Appl. Environ. Microbiol. 65 (7), 3182–
3191.
[9] Gieseke, A., Purkhold, U., Wagner, M., Amann, R. and
Schramm, A. (2001) Community structure and activity dynamics
of nitrifying bacteria in a phosphate-removing bioﬁlm. Appl.
Environ. Microbiol. 67 (3), 1351–1362.
[10] Hermansson, A. and Lindgren, P.E. (2001) Quantiﬁcation of
ammonia-oxidizing bacteria in Arable soil by real-time PCR.
Appl. Environ. Microbiol. 67 (2), 972–976.
[11] Dionisi, H.M., Layton, A.C., Harms, G., Gregory, I.R., Robinson, K.G. and Sayler, G.S. (2002) Quantiﬁcation of Nitrosomonas
oligotropha-like ammonia-oxidizing bacteria and Nitrospira spp.
from full-scale wastewater treatment plants by competitive PCR.
Appl. Environ. Microbiol. 68 (1), 245–253.
[12] Harms, G., Layton, A.C., Dionisi, H.M., Gregory, I.R., Garrett,
V.M., Hawkins, S.A., Robinson, K.G. and Sayler, G.S. (2003)
Real-time PCR quantiﬁcation of nitrifying bacteria in a municipal
wastewater treatment plant. Environ. Sci. Technol. 37, 343–351.
[13] Koops, H.P. and Pommerening-Roser, A. (2001) Distribution and
ecophysiology of the nitrifying bacteria emphasizing cultured
species. FEMS Microbiol. Ecol. 37, 1–9.
[14] Bollmann, A. and Laanbroek, H.J. (2002) Inﬂuence of oxygen
partial pressure and salinity on the community composition of
ammonia-oxidizing bacteria in the Schelde estuary. Aquat.
Microb. Ecol. 28, 239–247.
[15] Rowan, A.K., Snape, J.R., Fearnside, D., Barer, M.R., Curtis,
T.P. and Head, I.M. (2003) Composition and diversity of
ammonia-oxidising bacterial communities in wastewater treatment reactors of diﬀerent design treating identical wastewater.
FEMS Microbiol. Ecol. 43 (2), 195–206.
[16] Ballinger, S.J., Head, I.M., Curtis, T.P. and Godley, A.R. (1998)
Molecular microbial ecology of nitriﬁcation in an activated sludge
process treating reﬁnery wastewater. Water Sci. Technol. 37 (4–5),
105–108.
[17] Logemann, S., Schantl, J., Bijvank, S., van Loosdrecht, M.,
Kuenen, J.G. and Jetten, M. (1998) Molecular microbial diversity
in a nitrifying reactor system without sludge retention. FEMS
Microbiol. Ecol. 27, 239–249.
[18] Pynaert, K., Smets, B.F., Wyﬀels, S., Beheydt, D., Siciliano, S.D.
and Verstraete, W. (2003) Characterization of an autotrophic
nitrogen-removing bioﬁlm from a highly loaded lab-scale rotating
biological contactor. Appl. Environ. Microbiol. 69 (6), 3626–
3635.
[19] Klappenbach, J.A., Saxman, P.R., Cole, J.R. and Schmidt, T.M.
(2001) rrndb: The ribosomal RNA operon copy number database.
Nucl. Acids Res. 29, 181–184.
[20] Aakra, A., Utaker, J.B. and Nes, I.F. (1999) RFLP of rRNA
genes and sequencing of the 16S–23S rDNA intergenic spacer
region of ammonia-oxidizing bacteria: a phylogenetic approach.
Int. J. Sys. Bacteriol. 49, 123–130.
[21] Kurisu, F., Satoh, H., Mino, T. and Matsuo, T. (2002) Microbial
community analysis of thermophilic contact oxidation process by
using ribosomal RNA approaches and the quinone proﬁle
method. Water Res. 36, 429–438.
[22] Koops, H.P., Purkhold, U., Pommerening-Roser, A., Timmermann, G. and Wagner, M. (2003) The lithoautotrophic ammoniaoxidizing bacteria In: The Prokaryotes: an Evoluting Electronic
Resource for the Microbiological Community (Dworkin, M.
et al., Eds.). Springer Verlag, New York.
[23] Belser, L.W. and Schmidt, E.L. (1980) Growth and oxidation
kinetics of three genera of ammonia oxidizing nitriﬁers. FEMS
Microbiol. Lett. 7, 213–216.
[24] Purkhold, U., Pommerening-Roser, A., Juretschko, S., Schmid,
M.C., Koops, H.P. and Wagner, M. (2000) Phylogeny of all
recognized species of ammonia oxidizers based on comparative

T. Limpiyakorn et al. / FEMS Microbiology Ecology 54 (2005) 205–217

[25]

[26]

[27]

[28]

[29]

16S rRNA and amoA sequence analysis: implications for molecular diversity surveys. Appl. Environ. Microbiol. 66 (12), 5368–
5382.
Schramm, A., De Beer, D., Wagner, M. and Amann, R. (1998)
Identiﬁcation and activities in situ of Nitrosospira and Nitrospira
spp. as dominant population in nitrifying ﬂuidized bed reactor.
Appl. Environ. Microbiol. 64 (9), 3480–3485.
You, S.J., Hsu, C.L., Chuang, S.H. and Ouyang, C.F. (2003)
Nitriﬁcation eﬃciency and nitrifying bacteria abundance in
combined AS-RBC and A2O systems. Water Res. 37, 2281–2290.
Wagner, M. and Loy, A. (2002) Bacterial community composition
and function in sewage treatment systems. Curr. Opin. Biotechnol. 13, 218–227.
Bollmann, A. and Laanbroek, H.J. (2001) Continuous culture
enrichments of ammonia-oxidizing bacteria at low ammonium
concentrations. FEMS Microbiol. Ecol. 37, 211–221.
Regan, J.M., Harrington, G.W. and Noguera, D.R. (2002)
Ammonia- and nitrite-oxidizing bacteria communities in a pilot-

[30]

[31]

[32]

[33]

217

scale chloraminated drinking water distribution system. Appl.
Environ. Microbiol. 68 (1), 73–81.
Regan, J.M., Harrington, G.W., Baribeau, H., De Leon, R. and
Noguera, D.R. (2003) Diversity of nitrifying bacteria in full-scale
chloraminated distribution systems. Water Res. 37, 197–205.
Ward, B.B., Martino, D.P., Diaz, M.C. and Joye, S.B. (2000)
Analysis of ammonia-oxidizing bacteria from hypersaline Mono
Lake, California, on the basis of 16S rRNA sequences. Appl.
Environ. Microbiol. 66 (7), 2873–2881.
Urbain, V., Mobarry, B., de Silva, V., Stahl, D.A., Rittmann,
B.E. and Manem, J. (1998) Integration of performance, molecular
biology and modeling to describe the activated sludge process.
Water Sci. Technol. 37 (3), 223–229.
Kowalchuk, G.A., Bodelier, P.L.E., Heilig, H.J., Stephen, J.R.
and Laanbroek, H.J. (1998) Community analysis of ammoniaoxidizing bacteria, in relation to oxygen availability in soils and
root-oxygenated sediments, using PCR, DGGE and oligonucleotide probe hybridization. FEMS Microbiol. Ecol. 27, 339–350.


