DSpace at VNU: Fouling characterization and nitrogen removal in a batch granulation membrane bioreactor

International Biodeterioration & Biodegradation 85 (2013) 491e498

Contents lists available at SciVerse ScienceDirect

International Biodeterioration & Biodegradation
journal homepage: www.elsevier.com/locate/ibiod

Fouling characterization and nitrogen removal in a batch granulation
membrane bioreactor
Bui Xuan Thanh a, *, Chettiyapan Visvanathan b, Roger Ben Aim c
a

Faculty of Environment, Ho Chi Minh City University of Technology, Building B9, 268 Ly Thuong Kiet Str., District 10, Ho Chi Minh City 70000, Viet Nam
Environmental Engineering and Management Program, School of Environment, Resources and Development, Asian Institute of Technology, P.O. Box 4,
Klong Luang, Pathumthani 12120, Thailand
c
Université de Toulouse, INSA, UPS, INP, LISBP, 135 Avenue de Rangueil, F-31077 Toulouse, France
b

a r t i c l e i n f o

a b s t r a c t

Article history:
Received 16 November 2012
Received in revised form
19 February 2013
Accepted 21 February 2013
Available online 19 March 2013

A submerged membrane bioreactor (MBR) combined with aerobic granulation reactor was investigated

for the simultaneous organic/nitrogen removal and membrane fouling control. Total nitrogen (TN)
removal was 59% (1.76 mg TN/g VSS h) in the aerobic granulation reactor. The filtration of granulation
effluent or low operating F/M condition of the MBR could extend the filtration period of up to 78 days
without any need for physical cleaning The soluble fraction was the main contributor to fouling
compared to colloids and solids. The soluble polysaccharides (sPS) had more adverse effects than that of
soluble protein (sPN). The deposition on a unit of the membrane’s surface area was 11 mg sPS/L m2 and
8 mg sPS/L m2. As a result, the BG-MBR could be an alternative treatment process for simultaneous
organic/nitrogen removal and fouling control.
Ó 2013 Elsevier Ltd. All rights reserved.

Keywords:
Aerobic granules
Membrane bioreactor
Extracellular polymeric substances
Fouling

1. Introduction
The aerobic granular sludge process has been known to have
many advantages as compared to the conventional activated sludge
operations for about a decade. The aerobic granule possesses a
compact spherical structure, excellent settling ability, dense
biomass structure, high biomass retention, ability for simultaneous
nitrification-denitrification and removal of toxic substance (Beun
et al., 2002; Carvalho et al., 2006; Thanh et al., 2008; Shi et al.,
2011). The sludge is more stable in batch reactors due to the existence of feast and famine conditions in each cycle (Beun et al.,
2002). The organic and nitrogen removal in the granulation system is high compared to that of conventional activated sludge
process (Arrojo et al., 2004; Tay et al., 2007; Thanh et al., 2009;
Lotito et al., 2012). However, the single granular sludge reactor was
not able to meet the effluent standards due to the high suspended
solids content in the effluent. The suspended solids (SS) concentrations in the effluent of the granulation reactor were high,

ranging from 75 to 250 mgVSS/L (Beun et al., 2002) and 200 to

* Corresponding author. Tel.: þ84 907866073.
E-mail addresses:
(C. Visvanathan).

(B.X.

Thanh),



0964-8305/$ e see front matter Ó 2013 Elsevier Ltd. All rights reserved.
/>
450 mgTSS/L (Arrojo et al., 2004). Thus, a post treatment such as
membrane filtration could be an add-in polishing step for complete
treatment and water reuse.
Membrane technology has been proven to be the most effective
wastewater treatment system in recent decades. The advantages
are less footprint requirements due to a high substrate loading rate,
good treated water quality which can be reused for appropriate
operations, less sludge production rate, high biomass retention, and
microbial diversity, among others (Visvanathan et al., 2000).
Membrane fouling could be due to the deposition of suspended
solids/flocs (cake/gel formation, pore blocking), colloids (Bouhabila
et al., 2001) and solutes (Shane Trussell et al., 2006; Jarusutthirak
and Amy, 2006; Miyoshi et al., 2012). Recently, it has been found
that the fouling mechanism of the submerged MBR is mainly caused
by the deposition/accumulation of soluble extracellular polymeric
substances (sEPS) on the membrane if reversible fouling (cake formation) is well controlled. The sEPS mainly comprises of soluble

polysaccharide (sPS) and soluble protein (sPN). The fouling potential
of sPS, sPN or both of them is still unclear. Both sPS and sPN were
some of the factors which influenced membrane fouling (Shane
Trussell et al., 2006; Liang et al., 2007; Miyoshi et al., 2012) where
the sPS played a major role as membrane foulant (Rosenberger et al.,
2006; Jarusutthirak and Amy, 2006; Kim and DiGiano, 2006).


492

B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498

At the moment, there exists very limited published information
related to the fouling behavior of the aerobic granular reactor
effluent. Some researchers studied the filterability of MBR seeding
with pre-cultivated aerobic granules for a short operation period.
The granules used in the MBR were taken from a batch reactor. Li
et al. (2005) reported that the permeability of the MBR seeding
with pre-cultivated granules was 50% higher than that of the conventional MBR during 16 days of operation. The author proposed
that the compact and round shaped structure of the granule might
cause less fouling due to the contact of less floc particles with the
membrane’s surface. Additionally, Tay et al. (2007) also reported
that the filterability of pre-cultivated granules was much better
than that of conventional sludge flocs. Granular sludge had a
membrane permeability loss of 1.68-fold less than conventional
sludge flocs during the constant pressure test.
In this study, a hybrid system includes a submerged MBR
following a sequencing batch airlift reactor (SBAR) to filter the
effluent. This is named as a batch granulation membrane bioreactor
(or BG-MBR). This combination was selected instead of inserting

the membrane inside the granulation reactor because granular
sludge was not stable in the continuous operation mode. It is clearly
proven that granules could be stable with the cyclic feast and
famine conditions in a batch reactor (Beun et al., 2002; Tay et al.,
2007). The advantages of this hybrid system include high organic
and nitrogen removal efficiencies and fouling control. This paper
focuses on the investigation of high loading simultaneous organic
and nitrogen removal and the fouling characteristics of the BG-MBR
system. Further, the fouling behavior of sludge fractions was also
investigated.
2. Materials and methods
2.1. Experimental setup
Fig. 1 describes the BG-MBR system including a SBAR (granulation reactor), a settler and a submerged MBR. The SBAR which
operated in batch mode consisted of four cycles of operation. Air

was supplied through a porous stone diffuser from the bottom of
the reactor. Each batch of operation consisted of four stages namely
feeding (6 min), reaction (high aeration rate: 3 h; and low aeration
rate: 48 min), settling (3 min) and withdrawal (3 min). The high
aeration rate is to achieve oxidation of organic and nitrogen compounds and granule stability. Further, it was followed by low
aeration to reduce the aeration cost and to enhance the nitrogen
removal through the denitrification process occurring inside the
core of the granule. The denitrification process might be enhanced
by limitation of oxygen diffusivity into the core of the granule. The
SRT was not controlled in this study because the suspended solids
from SBAR effluent fluctuated according to time. The second unit
was the settler. The settler is a dual purpose tank to function as both
holding and settling tank (denoted as “settler”). The effluent of
SBAR was transferred into the settler which was then fed into the
MBR in a continuous mode of operation. Settled sludge of 500 mL/

d (twice, each 250 mL) from the settler was removed periodically.
The final unit, the submerged MBR was used for the separation of
liquid and solid fractions. The remaining substrate, unsettled colloids and pin flocs could be further biologically degraded in the
MBR. All these systems were controlled automatically by programmable logic controller. Table 1 shows the operating conditions
of BG-MBR system.
2.2. Wastewater and support media
The feeding wastewater contained 260 mg TOC/L (700 mg COD/
L as glucose), 190 mg N/L of NH4Cl, 50 mg/L of KH2PO4, 30 mg/L of
CaCl2$2H2O, 12 mg/L of MgSO4$7H2O, and 4 mg/L of FeCl3
throughout the experiment. Trace elements were added at the rate
of 1 mL/L of wastewater as described by Thanh et al. (2008).
The shell carrier produced from the shell of white rose cockle
was added to act as a support for microbial granule formation.
The carrier was used to enhance the structure, round shape, and
physico-chemical characteristics of the granules. The shells were
dried, ground and sifted with sieve Nos. 70 and 100, to reach a
fraction between 150 and 212 mm. The powder obtained was

Influent valve
Manometer

Effluent valve
Level control

Permeate
pump

membrane

SBAR

MBR
Settler

Pump

Air supply
Fig. 1. Experimental set-up of BG-MBR system.


B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498
Table 1
Operating conditions of BG-MBR system.
Parameters

SBAR

Settler

MBR

Working
volume (L)
Size (length Â
diameter)

9.7

8

4

þ Down-comer:
120 cm  11.5 cm
þ Raiser: 90 cm  7 cm
0.86 Æ 0.22
0.6 Æ 0.1
þ 6 batches/d;
4 h/batch:
À Feeding: 6 min;
À Reaction: high
aeration rate (3 h)
followed by low
aeration rate (48 min);
À Settling: 3 min
À Withdrawal: 3 min
þ High aeration: 1.67
þ Low aeration: 0.08
7.3
w24
e

e

53 cm  10 cm

e
e
e

e

e
Intermittent
suction (7 min on/
3 min off)

e

0.3

6
e
e

e

e

3.4
20
PE 0.1 mm, 0.42 m2,
Mitsubishi, Japan
2.8 L/m2 h

OLR (kg TOC/m3 d)
NLR (kg N/m3 d)
Operating mode

Air velocity
(cm/s)
HRT (h)

SRT (d)
Membrane
module
Membrane
flux (net)

then washed and dried before used in the experiment. The carrier had a bulk density of 1.45 g/cm3 and a weight loss of 2% at
550  C within 20 min. Initially, the amount of the carrier added
was 200 g (20 g/L) to SBAR. Ten grams were added every month
to compensate for the loss through sampling and effluent
discharge.
2.3. Analytical methods
The supernatant samples from the settler and the MBR mixed
liquor were prepared once a day by centrifugation at 4500 rpm for
15 min (Centrifuge M/c Universal 320R, Germany). The procedure
for getting supernatant sample was described by Bouhabila et al.
(2001). Dissolved organic carbon (DOC) was determined by a TOC
analyzer (TOC-VCSN, Shimadzu, Japan). Parameters of ammonia,
nitrite and nitrate were measured according to standard methods
(APHA, 1998) (Total Nitrogen or TN ¼ NH3eN þ NO2eN þ NO3eN).
In addition, the PN and PS were analyzed by methods of Lowry et al.
(1951) and Dubois et al. (1956), respectively (EPS ¼ PS þ PN). The
UVA254 was measured by using 1-cm quartz cell by UV/Visible
Spectrophotometer (U-2001, Hitachi, Japan) where specific ultraviolet absorbance (SUVA) was calculated from the ratio of UV254
and DOC.
During the operation, sludge characterization was conducted
for samples of granular sludge, mixed effluent from SBAR and MBR
sludge. SVI and MLSS measurements were determined according
to standard methods. MLVSS of shell granular biomass was not able
to measure accurately by gravitational method according to standard methods. For this kind of shell granules, measurement of

MLVSS of gravitational method is not accurate due to the loss of
shell carriers when mixed biomass (cell and carrier) are burned at
temperature more than 450  C. Thus the authors selected the indirect measurement method which measures the total organic
carbon (TOC) of cell biomass and then converts into cell mass
based on the cell formulae. To measure MLVSS, sludge samples
were ground for 1 min with the Ultra-Turrax machine (Ika-Werk,
Germany) before homogenous sonication at 100 hz for 4 min (Ultra
Sonic processor, CP130, USA). The sample was then diluted with
milli-Q water and stirred in a volumetric flask for 10e20 min at

493

500 rpm until homogenization occured. Biomass in terms of
MLVSS was determined by measuring TOC of the homogenized
sample. Then, the value of TOC was converted to MLVSS
(multiplied by the factor 2.05) (Tijhuis et al., 1994). The particle
size distribution of samples for the MBR sludge and settler were
determined by the laser diffraction technique (MastersizerS, Malvern, UK, and detection range of 0.05e900 mm). The size of
colloidal fraction was examined by the zetasizer nanoZS after
centrifuging at 4500 rpm and 4  C for 1 min (detection range of
0.6e6000 nm). The size of the granules was measured by a digital
camera with a transparent scale located under the beaker containing the granules.
The bound EPS (bEPS) of granular sludge, the MBR sludge and
fouling layer sample were extracted using the cation exchange
resin technique (Dowex HCR-S/S, 16e50 mesh, sodium form, Dow
Chemical Company) according to Frølund et al. (1996). For granular
sludge sample, it was ground by the Ultra-Turrax equipment for
one minute before carrying out resin extraction. The extraction was
conducted with resin dosage (60 g/gVSS) and stirring speed of
600 rpm for 45 min. Then, the bEPS solution was centrifuged at

4500 rpm and 4  C for 15 min (twice).
The amount of PS deposition on the membrane was quantified with the same method adopted by Kim and DiGiano (2006).
Two fibres (about 10e30 cm) were cut off from the fouled
membrane and washed with tap water until the membrane fibre
became white/clean like cleaned (initial) membrane (removal of
entire fouling layer attached on the fibre). The fibres were cut
into small segments and immersed into a test tube containing
2 ml milli-Q water. Then the color reagent (1 ml of phenol 5%
and 5 ml of concentrated H2SO4) was added into the test tube.
This analytical procedure is similar to Dubois’ measurement
method. The PS deposition on the membrane was measured at a
wavelength of 490 nm and converted to the unit of mg PS/cm2 of
fibre.
Modified fouling index (MFI) and cake resistance was measured
by a stirred cell (AMICON 8400 USA, diameter 67 mm,
area ¼ 41.8 cm2) with a stirring speed of 500 rpm and a flat sheet
membrane with pore size of 0.22 mm under a constant transmembrane pressure of one bar. The raw experimental data (V and
t) were used to plot t/V versus V graph to get the slope (s/L2) which
represents the MFI of the sample. The MFI is defined as the gradient
of the linear region found in the cake filtration equation (Eq. (1)).
Cake resistance (1/m2) was estimated as multiplying by specific
cake resistance a (m/kg) and cake mass C (kg/m3) (Boerlage et al.,
2002).

m$a$C
m$Rm
t
¼
Vþ
V

A$TMP
2$A2 $TMP

(1)

Rc ¼ aC

(2)

In addition, the fouling potential of sludge fractions in the MBR,
including suspended solids (SS), colloids (CL) and solutes (SL), were
quantified. The separation of SS, CL and SL were prepared according to
Bouhabila et al. (2001). The mixed liquor MBR sludge sample contained SS þ CL þ SL. The sample, containing CL þ SL, was achieved by
centrifugation at 4500 rpm and 4  C for one minute. Finally, the
sample containing only SL was centrifuged twice at 4500 rpm and 4  C
for 15 min. Resistance of each fraction was calculated as follows:

Rt ¼ Rm þ Rf

(3)

Rt, Rm, Rf are total, clean membrane and fouling resistance,
respectively.


494

B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498

3. Results and discussions

3.1. Organic and nitrogen removal in the SBAR
Aerobic granules were cultured from conventional activated
sludge and with shell media. Biomass started covering the surface
of shell carriers during the first 10 days. Shell granules started
forming after 30 days of operation. The granule size gradually
increased and varied from 0.5 mm to 9.0 mm during next 80 days.
The average size of matured granule was 4.9 Æ 1.0 mm. The color of
granule changed from light yellow (initial granules) to dark yellow
(mature granules). The average settling velocity was 260 Æ 124 m/h
which was much higher than that of activated sludge (1e2 m/h).
Granules were stable during the study period. The study on nitrogen removal and fouling behavior below was conducted since
granules matured in the reactor (i.e., after 110 days).
At the stage of matured granule formed, organic matter was
quickly removed by an aerobic granulation system in each batch.
Fig. 2 shows evolution of concentrations of organic matter and nitrogen species with time in a typical batch. 91% of TOC removal was
achieved within the first 30 min and 94% removal reached during
the 90 min of aeration stage. The dissolved oxygen (DO) concentration during the high aeration stage was saturated during first 3 h
of the stage of high aeration rate (w6.5 mg/L) and then reduced to
4 mg/L during the next 48 min due to lower aeration rate. The pH of
SBAR slightly fluctuated after 2 h of operation due to simultaneous
nitrification (alkalinity consumption) and denitrification (alkalinity
production) in the outer and the inner layer of granules, respectively. The influent ammonia was converted into nitrite and nitrate
where nitrite was dominant due to partial nitrification in the SBAR.
The complete nitrification did not occur due to the free ammonia
concentration in the reactor. The free ammonia inhibited the
nitratation (Anthonisen et al., 1976). Complete nitrogen removal
(converted to nitrogen gas) was observed to occur during first

TOC (mg/L)

300

Low aeration rate

High aeration rate

250
200
150

TOC

100

DO

1
0
240

0
30

60

90

120 150
Time (min)

180

210

200
Concentration (mg/L)

NH4-N

NO2-N

NO3-N

TN

160
120
80
40
0
0

30

60

90

120 150
Time (min)

7
6
5
4
3
2

pH

50
0

9
8

180

210

240

Fig. 2. TOC, DO, pH and nitrogen species profile of SBAR in a batch.

DO (mg/L); pH

350

30 min in which the bulk liquid was rich in organic and nitrite. This
reveals that the denitrification process could be achieved in the

granules when organic substrate is available. The simultaneous
nitrification-denitrification (SND) could take place in both low and
high aeration stages depending on the availability of organic substrate due to the structure of the granule. The DO in the outer layer
of the granule was almost as high as the bulk liquid while that of
the inner core was very low due to the limitation of oxygen transfer
from the bulk liquid to the core of granules. This phenomenon
allowed the denitrification process to occur inside the core of the
granule. As reported by Tijhuis et al. (1994), the anoxic/anaerobic
condition could be achieved at the depth of 300 mm below the
granule surface. In this study, the average size of the granules was
about (4.7 Æ 1.4 mm) whose radius was almost much greater than
the diffusion depth of oxygen inside the core of granule. This could
lead to the anoxic condition in the core. Generally, the special
spherical structure of granules favors the SND phenomena to occur
in the single aerobic granulation reactor even with bulk DO concentration higher than 4 mg/L.
Fig. 2 shows that the ammonia nitrogen was completely
oxidized into nitrite nitrogen during the first three hours of operation in which the removal rate was 0.015 mg N/gVSS h (0.18 mg N/
L h). The nitrite production rate was 0.013 mg N/g VSS h (0.16 mg N/
L h) where it was converted from free ammonia with time. Conversion of nitrite nitrogen into nitrate nitrogen was not significant
in SBAR due to the existence of free ammonia in the reactor. Previous research reported that free ammonia inhibition threshold
was 0.1e4.0 mg/L for nitrobacter which plays a role in the oxidation
of nitrite (Yang et al., 2004). In this process, the dynamic balance of
nitrogen species occurred between ammonia consumption and
nitrite production. The concentration of TN did not change drastically during the last 3.5 h which could conclude that the SND
reached its maximum efficiency at this operating condition. The
SND only occurs during the first duration where the organic substrate is available. The removal efficiency of TN becomes insignificant since available organic matter is limited in the SBAR. Some
other research also found that the nitrite-oxidizing bacteria (NOB)
are not favorable in granular sludge. Li et al. (2013) reported that
compared to sludge flocs sludge granulation with selective sludge
discharge help halt ammonia oxidation to the level of partial

nitrification rather than complete nitrification. This is also
confirmed by the molecular analysis that aerobic granulation
resulted in ammonia-oxidizing bacteria (AOB) enrichment and
reduction of nitrite-oxidizing bacteria (NOB). In addition, Shi et al.
(2011) also postulated that a fairly large proportion of AOB was
close to the granule surface but NOB were rarely found. The granules had excellent partial nitrification ability due to inhibition of
free ammonia (FA) and limited DO diffusion within granules.
The data set at steady state during 44 days was used for nitrogen
balance. The result shows that the TN removal is 59% which includes 12% TN removal by biological assimilation and 47% by the
SND process in the SBAR. The overall denitrification rate is
22.2 mg N/L/h (1.76 mg N/g VSS h) under aerobic condition without
external substrate addition in the reactor. In practice, complete
nitrogen removal could be fully achieved in the availability of
electron donor or lower ammonia concentration in the feed. The
organic matter removal and SND in the SBAR indicates that there
was co-existence of heterotrophic, nitrifying and denitrifying
population in the structure of aerobic granules. Nitrogen removal
could take place even in aerobic granulation reactor (DO greater
than 4 mg/L). The results indicate that the complicated anaerobic/
anoxic/aerobic system could be integrated in a single aerobic
granulation reactor.
Table 2 summarizes the treatment performance of BG-MBR
system. most of the organic matter was removed in the SBAR


B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498
Table 2
Treatment performance of BG-MBR.
Parameters

SBAR

Settlera

MBRb

DOC removal (%)
Ammonia removal (%)
Overall TN removal (%)
SVI (mL/g)
F/M (dÀ1)
MLVSS (mg/L)
Settling velocity (m/h)
Average particle size
(% volume)

97.3
99.9
59
25 (Æ5)
0.18 (Æ0.05)
12,600 (Æ240)
260 (Æ125)
4.9 (Æ0.2) mm

e
e
19
e
e

35 (Æ15)
<10
62.3 (Æ1.3) mm

61.7
e
3.3
93 (Æ26)
e
2200 (Æ600)
<10
108.6 (Æ1.5)mm

a
b

Removal efficiency of settler is calculated during a batch (4 h).
removal efficiency of MBR is calculated between settler and permeate.

concentration (mg/L)

NH4-N
NO3-N

NO2-N
TN

150
100
50

0
-50
Influent

settler

MBR
supernatant

permeate

Fig. 3. Nitrogen species change in the BG-MBR

(DOC removal >97%). The VSS concentration in the SBAR effluent
and in the settler supernatant was 239 Æ 42 mg/L and 35 Æ 15 mg/L,
respectively. The concentration of suspended solids in granulation
effluent was as high as other research results. The SBAR effluent
contained certain amount of pin flocs, which could not be settled by
conventional gravitational settling. This made the supernatant
turbid. This is one of disadvantages of the granular sludge technology. Approximately, 85% of suspended solids settled in the
settler before fed into the MBR. The remaining soluble substrate
(mainly nitrite and organic residue) and unsettled pin flocs were
further aerobically treated in the MBR. The SBAR effluent and the
MBR permeate were rich in nitrite (75 Æ 18 mg/L) and nitrate
(77 Æ 12 mg/L), respectively. Hence, the MBR plays the roles of posttreatment as polishing and complete nitrification (Fig. 3).
The SRT of SBAR was approximately 24 days which was calculated based on the ratio of sludge in the reactor over the sludge
wastage. Particularly for the granular sludge reactor, the actual SRT
was much higher than the calculated one. The SRT of granular

25

(volume)
(number)

30
%

%

15

There are two modes of measurement for particle size distribution (PSD), namely volume distribution and number distribution.
In terms of volume distribution, the particle size of the settler and
the MBR mixed liquor was 98 mm and 158 mm, respectively. However in terms of number distribution it was 0.53 mm and 0.20 mm
(Fig. 4). For the light scattering technique, the volume distribution
did not provide the representative size of majority of bio-particles
because there was a large distribution range in the sludge samples. The volume of all small particles only made up of small volume percentage. Therefore, the number distribution mode could
reflect more accurately the actual size of the measured samples.
The colloidal size measurement confirms that the nanosize of the
MBR sludge was 262 nm (0.26 mm), almost similar to the result
achieved from the mixed sludge sample (0.20 mm) (Fig. 4). For the
colloidal size measurement, the number and volume distribution
were rather identical because the centrifugation step removed all
the large particles and made the two distribution curves become
narrow and comparable.
The MBR sludge sample showed wider distribution and a
smaller size than the settler sample (number distribution). Again,
this indicates that the sludge flocs were disintegrated and/or
deflocculated in the MBR due to the endogenous condition of the
MBR. The shear stress of aeration, again, could break the linkage of

floc structure and produce pin flocs, debris and soluble EPS. The
destructuration was certainly due to erosion strengths or to ruptures of the network of polysaccharides fibrils which was the
support of the different compounds and particularly of the cells.
Wisniewski and Grasmick (1998) found that flocs decrease in the
settleable fraction and consequently, an increase in the nonsettleable one. This is in line with this result in which the particle
size was reduced. The deflocculation makes the particle size
smaller which corresponds to the increase of smaller sludge particles in the MBR compared to the settler.
In the fouling sense, the particle size of the MBR sludge was
larger than the pores of the membrane, thus the particles had less
possibility to infiltrate into the membrane pores. If the fouling is
caused by the suspended solids, it is a reversible fouling which
could be eliminated by a physical cleaning technique.

40

MBR (volume)
Settler (volume)
MBR (number)
Settler (number)

20

sludge reactor was different from that of the conventional activated
sludge reactor. At the steady state, the washed-out sludge was just
the newly grown cells or pin flocs generated from the biological
assimilation and granule detachment. The granules, containing old
biomass, were retained in the reactor until disintegration. This
made the actual SRT is longer than the calculated one in the
granular sludge reactor. Granules were retained in the reactor since
they had excellent settling ability compared to flocs. Therefore, the

slow-growing microorganisms could exist to perform the SND
process and to degrade the refractory.
3.2. Particle size distribution of the MBR sludge

250
200

495

20

10
10

5
0
0.01

0

0.1

1

µm

10

100

1000

0.1

1

10
100
1000
nanosize (nm)

10000

Fig. 4. Typical particle size distributions of settler sample and MBR sludge (left); Particle size of colloidal fraction in MBR sludge (4500 rpm, 1 min) (right).


496

B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498

sPN

Accumulation

sPS/sEPS ratio
1.0
Deposition

15

0.9

10

0.8

5

0.7

0

0.6
settler

MBR

UVA
0.20

Fraction/sludge type SS

Accumulation

DOC
16
14

Deposition

12
10

0.10
0.05

8
6
4

0.00

2
0

UVA (1/cm)

0.15

MBR

MBR filtering
granulation
effluent
MBR sludge
MBR sludge
(solute
separation)

CL

SL

Remark

2 12 86 No backwash,
HF, PE

Reference
This study

24 50 26 Backwash, HF
Bouhabila et al., 2001
23 25 52 Backwash,
Wisniewski and
ceramic membrane Grasmick, 1996

Note: HF: Hollow fibre, PE: Polyethylene.

Permeate
SUVA

settler

Table 4
Comparison of fouling potential of sludge fractions (%).

sPS/sEPS ratio

sPS

DOC (mg/L); SUVA (L/mg/m)

concentration (mg/L)

20

permeate

Fig. 5. Soluble characteristics (settler, MBR supernatant, permeate).

3.3. Fouling behavior in the MBR
The MBR was operated under endogenous conditions with a low
incoming substrate and a low biomass concentration. The bPN
(bounded PN) is usually much higher than bPS (bounded PS) in the
MBR sludge. The bPS/bPN ratio in Massé et al. (2006) was 0.25e0.5
and in Le Clech et al. (2006) was 1/3. In this study it is 0.7. Massé
et al. (2006) noticed that protein compounds are more easily
degradable than polysaccharides. The reason for the low sPN concentration in the bulk liquid was due to the quick sPN consumable
rate compared to the bounded one. While the sPS was not readily
degradable as sPN, sPS was accumulated in the MBR with time. In
addition, the deflocculation phenomena occurred in the MBR due
to its low operating F/M conditions. This causes the production of
EPS brought about by the release of bridging polymers from the
flocs structure (Wisniewski and Grasmick, 1998).
Fig. 5 illustrates the fouling behavior of three different sludge
fractions, namely suspended solids (as SS), colloids (as CL) and
solute (as SL). The fouling potential in term of MFI of SS, CL and SL
fractions occupied 12, 39 and 49% of the total fouling potential of
MBR sludge, respectively. The resistance of SS, CL and SL fractions

was 0.01 Â 1012 mÀ1, 0.33 Â 1012 mÀ1 and 2.38 Â 1012 mÀ1 (inferred

Table 3
Fouling potential of sludge fractions in MBR.
Fractions of sludge

SS-CL-SL

CL-SL

SL

MFI20 (103 s/L2)
a*C (1/m2)
a (m/kg)
Rt (mÀ1)
Rm (mÀ1)
Rf ¼ RteRm (mÀ1)

86.7 (0.986)a
3.02*1014
1.37*1014
2.83*1012
1.12*1011
2.72*1012

76 (0.999)a
2.65*1014
NA
2.82*1012

1.12*1011
2.71*1012

42.4 (0.996)a
1.48*1014
NA
2.49*1012
1.12*1011
2.38*1012

NA: not applicable.
a
The numbers in the brackets is R2 of the linear segments of the time to volume
profile.

from Table 3) which makes up 2, 12 and 86%, respectively. This
could support the notion that the suspended solids and colloidal
fractions do not strongly influence flux decline. The soluble fraction
is the main fouling contributor among the sludge fractions in the
case of granulation effluent. The comparison of fouling potential of
sludge fractions with other studies is presented in Table 4.
In addition, it was observed that the formation of the cake layer
took a long time (70 days) to form on the membrane surface even
without backwash application. The membrane was fouled on day
78 with the fouling rate as low as 0.027 kPa/d. There was no
complete cake layer formation on the membrane surface during
operation. The white color (original) of membrane fibres could still
be seen at most area of fibres. This observation is very different
from the case of conventional submerged MBR. Sludge cake fully
covers all the surface of membrane fibres as reported by Khan and

Visvanathan (2008). It appears that the low F/M operating conditions or low organic loading rate could prolong the filtration period.
This study is in line with results of some researchers (Barker and
Stuckey, 1999; Rosenberger et al., 2006; Shane Trussell et al., 2006).
Fig. 5 shows an increment of sPS in the MBR supernatant and
then a slight reduction in the permeate. On the other hand, the
trend of sPN shows a slight increment with sPS while sPN does not
increase much in the MBR supernatant and is negligible in the
permeate. As observed, sPS is always much higher than sPN in the
filtration system (the ratio of sPS/sEPS ¼ 0.72e0.98). The concentrations of sPS in the settler, the MBR supernatant and permeate
were 7.2 Æ 1.1 mg/L, 14.9 Æ 2.6 mg/L and 10.3 Æ 2.2 mg/L, respectively, during the membrane fouling cycle (78 days); while
sPN concentrations were 2.9 Æ 2.1 mg/L, 3.5 Æ 1.3 mg/L and
0.2 Æ 0.2 mg/L.
The sPS in the settler could be the byproduct of substrate
metabolism generated from the granular sludge activity. The increase of sPS in the MBR supernatant as compared to the settler
could be caused by two reasons, namely rejection by membrane
(Liang et al., 2007) and deflocculation. Barker and Stuckey (1999)
postulated that the bound EPS is hydrolyzed to soluble EPS which
is called biomass associated product (BAP). In addition, the DOC of
the settler, the MBR supernatant and permeate were 6.0 Æ 3.1 mg/L,
9.8 Æ 5.2 mg/L and 2.9 Æ 1.9 mg/L, respectively, during the duration
of the operation. The concentrations of sPN in the settler and in the
MBR supernatant were similar during operation. This can be
explained by the biodegradation of protein compounds in the MBR.
Table 5
Bound EPS of fouling layer, MBR sludge and granule.
Bound EPS

bPS
bPN
bEPS

bPS/bPN
(mgPS/gVSS) (mgPN/gVSS) (mgEPS/gVSS)

bEPS of fouling
10.5 (Æ0.4)
layer (n ¼ 2)
bEPS of MBR
18.4 (Æ7.7)
sludge (n ¼ 6)
bEPS of granule (n ¼ 7) 10.7 (Æ1.4)
Note: n is number of measurements.

19.9 (Æ1.9)

30.4

0.5

39.9 (Æ11.5)

58.3

0.7

17.0 (Æ2.4)

27.7

0.6

B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498

497

Table 6
Comparison of polysaccharides deposition on membrane.
Items

This study

Miyoshi et al., 2012

Kim and DiGiano, 2006

Cho and Fane, 2002

PS deposition (mg/cm2)
Cumulative filtrate volume per
unit of membrane area (L/m2)
Operating flux (L/m2/h)
Membrane material/pore size
Module configuration
Filtration

20 Æ 1
524 (78 days)

2.29e4.02
e

3e10
1000e3600

20e70 (200e700 mg/m2)
NA

2.8
MF 0.1 mm
Submerged HF
Granulation effluent

33.3
MF 0.4 m
Pilot-scale MBR
Domestic wastewater

80e20
MF 0.22 m
Flat sheet
UASB effluent

Backwashing

No

No

50
UF 150 kDa

Pressurized filtration HF
Secondary effluent
(pretreated by sand filter & MF 150 mm)
Yes

The concentrations of sPS and sPN decreases in permeate
compared to the MBR supernatant proclaims that they were both
trapped on the surface and inside of the pores of the membrane. The
amount of sPS and sPN adsorbed in the membrane was about 31%
and 94%, repsectively, compared to that of the MBR supernatant.
This indicates that the sPS in the MBR supernatant was partially
deposited while the sPN was retained completely on the membrane
surface. The deposition loading on a unit of the membrane surface
was calculated by the loss of concentration after passing through the
membrane which was at 11 mg sPS/L m2 and 8 mg sPS/L m2. The
difference of deposition percentage of soluble macromolecules (sPS
and sPN) shows that they possess different characteristics. The
partial deposition of sPS on the membrane could be hypothesized
that there were two main fractions of sPS existing in the MBR supernatant (large and small molecules relative to membrane pore
size). The large ones were deposited on the membrane and the
smaller ones were passed through it (Thanh et al., 2010).
Although the DOC in the MBR supernatant increased compared
to that in the settler, the UVA254 and SUVA showed decreasing
trends. This indicates that there is a reduction of double bond
substances (such as humic-like materials, sPN) which are prone to
absorb UV light (Jarusutthirak and Amy, 2006). This correlates with
the majority of the sPS present in the MBR which is usually a long
chain of macromolecules with less double bond linkages. The DOC
reduction in the permeate, after passing through the membrane,
again confirms that the DOC (i.e., mainly sPS and sPN) sludge was

deposited on the membrane’s surface. UVA254 of permeate reduction means that the double bond compounds (mostly protein) were
trapped on the membrane. The passage through the membrane
could be mainly the small MW organic matters such as low molecular weight PS portions and humic-like materials.

3.4. Bound EPS of fouling layer and EPS deposition on membrane
The bEPS of fouling layer was extracted to understand its characteristics, fouling behavior and to compare it with that of the MBR
sludge and granular sludge. The bEPS of the fouling layer was
similar to that of granular sludge and approximately half of the
MBR sludge (Table 5). This result implies that the biomass in the
fouling layers on the membrane starts to experience lysis due to the
dense biomass concentration and limitation of substrate transfer
from the bulk liquid. The sludge particles may not contribute
significantly to fouling propensity when it is moving as a bulk liquid
as mentioned but when it attaches to the membrane as a fouling
layer; it could enhance the reversible fouling. This is the reason for
the TMP “jump” observed in this study similar to another lab-scale
submerged MBR (Zhang et al., 2006).
Table 6 presents the comparison data of polysaccharides deposition on the membrane and inside the pores for various operating
modes of the MBR. It appears that the deposition of the sPS on the
MF membrane is high because the sPS can penetrate into and

e

adsorb on the surface and pores of membrane. This result shows
the evidence of sPS deposited inside the pores of the membrane.
4. Conclusions
The research presented here focused on the simultaneous
organic/nitrogen removal and fouling behavior of a batch granulation membrane bioreactor. The following conclusions are drawn:
 The simultaneous nitrification and denitrification could be
achieved by a single granular sludge reactor even at aerobic

conditions. The total nitrogen removal is 59% or 22 mg TN/L h
(1.76 mg TN/g VSS h) at OLR of 0.86 kg TOC/m3 d.
 Submerged MBR coupling with the granular sludge reactor
(BG-MBR) extends the filtration duration up to 78 days without
any physical cleaning techniques (slow fouling rate of
0.027 kPa/d).
 Soluble extracellular polymeric substances are also the main
fouling causes in the BG-MBR system that is similar to the case
of conventional MBR. Polysaccharides and proteins are both
deposited on membrane pores and surfaces where polysaccharides are found to be the major deposition factor. The
deposition loading on the membrane is 11 mg/L m2 and 8 mg/
L m2 for soluble polysaccharides and soluble protein, respectively. The amount of polysaccharides deposited on membrane
fibres after 78 days of filtration is 20 mg/cm2.
References
Anthonisen, A.C., Loehr, R.C., Prakasam, T.B.S., Srinath, E.G., 1976. Inhibition of
nitrification by ammonia and nitrous acid. Journal of the Water Pollution
Control Federation 48, 835e852.
APHA, 1998. Standard Methods for Examination of Water and Wastewater, twentieth ed. USA, APHA-AWWA-Washington, DC.
Arrojo, B., Mosquera-Corral, A., Garrido, J.M., Mendez, R., 2004. Aerobic granulation
with industrial wastewater in sequencing batch reactors. Water Research 38,
3389e3399.
Barker, D.J., Stuckey, D.C., 1999. A review of soluble microbial product (SMP) in
wastewater treatment systems. Water Research 33, 3063e3082.
Beun, J.J., van Loosdrecht, M.C.M., Heijnen, J.J., 2002. Aerobic granulation in a
sequencing batch airlift reactor. Water Research 36, 702e712.
Boerlage, S.F.E., Kennedy, M.D., Dickson, M.R., El-Hodali, D.E.Y., Schippers, J.C., 2002.
The modified fouling index using ultrafiltration membranes (MFI-UF): characterization, filtration mechanisms and proposed reference membrane. Journal of
Membrane Science 197, 1e21.
Bouhabila, E.H., Ben Aim, R., Buisson, H., 2001. Fouling characterization in membrane bioreactors. Separation and Purification Technology 22e23, 123e132.
Carvalho, G., Meyer, R.L., Yuan, Z., Keller, J., 2006. Differential distribution of

ammonia- and nitrite-oxidising bacteria in flocs and granules from a nitrifying/
denitrifying sequencing batch reactor. Enzyme and Microbial Technology 39,
1392e1398.
Cho, B.D., Fane, A.G., 2002. Fouling transient in normally sub-critical flux operation
of a membrane bioreactor. Journal of Membrane Science 209, 391e403.
Dubois, M., Gilles, K.A., Hamilton, J.K., Rebers, P.A., Smith, F., 1956. Calorimetric
method for determination of sugars and related substances. Analytical Chemistry 28, 350e356.
FrØlund, B., Palmgren, R., Keiding, K., Nielsen, P.H., 1996. Extraction of extracellular
polymer from activated sludge using a cation exchange resin. Water Research
30, 1749e1758.


498

B.X. Thanh et al. / International Biodeterioration & Biodegradation 85 (2013) 491e498

Jarusutthirak, C., Amy, G., 2006. Role of soluble microbial products (SMP) in
membrane fouling and flux decline. Environmental Science and Technology 40,
969e974.
Khan, S.J., Visvanathan, C., 2008. Influence of mechanical mixing intensity on a
biofilm structure and permeability in a membrane bioreactor. Desalination 231,
253e267.
Kim, J., DiGiano, F.A., 2006. A two-fibre, bench-scale test of ultrafiltration (UF) for
investigation of fouling rate and characteristics. Journal of Membrane Science
271, 196e204.
Le Clech, P., Chen, V., Fane, T.A.G., 2006. Fouling in membrane bioreactors used in
wastewater treatment. Journal of Membrane Science 284, 17e53.
Li, A.J., Li, X.Y., Yu, H.Q., 2013. Aerobic sludge granulation facilitated by activated
carbon for partial nitrification treatment of ammonia-rich wastewater. Chemical Engineering Journal 218, 253e259.
Li, X.F., Gao, F.S., Hua, Z.Z., Du, G.C., Chen, J., 2005. Treatment of synthetic wastewater by a novel MBR with granular sludge developed for controlling membrane fouling. Separation and Purification Technology 46, 19e25.

Liang, S., Liu, C., Song, L., 2007. Soluble microbial products in membrane bioreactor
operation: behaviors, characteristics and fouling potential. Water Research 41,
95e101.
Lotito, A.M., Fratino, U., Mancini, A., Bergna, G., Iaconi, C.D., 2012. Effective aerobic
granular sludge treatment of a real dyeing and textile wastewater. International
Biodeterioration & Biodegradation 69, 62e68.
Lowry, O.H., Resebrough, N.J., Farr, A.L., Randall, R.J., 1951. Protein measurement
with the folin phenol reagent. Journal of Biological Chemistry 193, 265e275.
Massé, A., Spérandio, M., Cabassud, C., 2006. Comparision of sludge characteristics and performance of a submerged membrane bioreactor and an activated sludge process at high solids retention time. Water Research 40,
2405e2415.
Miyoshi, T., Aizawa, T., Kimura, K., Watanabe, Y., 2012. Identification of proteins
involved in membrane fouling in membrane bioreactors. International Biodeterioration & Biodegradation 75, 15e22.
Rosenberger, S., Laabs, C., Lesjean, B., Gnirss, R., Amy, G., Jekel, M., Schrotter, J.-C.,
2006. Impact of colloidal and soluble organic material on membrane performance in membrane bioreactors for municipal wastewater treatment. Water
Research 40, 710e720.

Shane Trussell, R., Merlo, R.P., Hermanowicz, S.W., Jenkins, D., 2006. The effect of
organic loading on the process performance and membrane fouling in a submerged membrane bioreactor treating municipal wastewater. Water Research
40, 2675e2683.
Shi, Y.J., Wang, X.H., Yu, H.B., Xie, H.J., Teng, S.X., Sun, X.F., Tian, B.H., Wang, S.G., 2011.
Aerobic granulation for nitrogen removal via nitrite in a sequencing batch reactor
and the emission of nitrous oxide. Bioresource Technology 102, 2536e2541.
Tay, J.-H., Yang, P., Zhuang, W.Q., Tay, S.T.L., Pan, Z.H., 2007. Reactor performance and
membrane filtration in aerobic granular sludge membrane bioreactor. Journal of
Membrane Science 304, 24e32.
Thanh, B.X., Sperandio, M., Guigui, C., Ben Aim, R., Wan, J.F., Visvanathan, C., 2010.
Coupling sequencing batch airlift reactor (SBAR) and membrane filtration: influence of nitrate removal on sludge characteristics, effluent quality and
filterability. Desalination 250, 850e854.
Thanh, B.X., Visvanathan, C., Ben Aim, R., 2009. Characterization of aerobic granular
sludge at various organic loading rates. Process Biochemistry 44, 242e245.

Thanh, B.X., Visvanathan, C., Sperandio, M., Ben Aim, R., 2008. Fouling characterization in aerobic granulation coupled baffled membrane separation unit.
Journal of Membrane Science 318, 334e339.
Tijhuis, L., Van Loosdrecht, M.C.M., Heijnen, J.J., 1994. Formation and growth of
heterotrophic aerobic biofilms on small suspended particles in airlift reactors.
Biotechnology and Bioengineering 44, 595e608.
Visvanathan, C., Ben Aim, R., Parameshwaran, K., 2000. Critical review in environmental science and technology: membrane separation bioreactors for wastewater treatment. Environmental Science and Technology 30, 1e48.
Wisniewski, C., Grasmick, A., 1996. Membrane bioreactor for treatment of polluted
effluents: biological performance and advantages. Med Fac Landbouw Univ
Gent 61/4b, 2017e2024.
Wisniewski, C., Grasmick, A., 1998. Floc size distribution in a membrane bioreactor
and consequences for membrane fouling. Colloids and Surfaces A: Physicochemical and Engineering Aspects 138, 403e411.
Yang, S.F., Tay, J.-H., Liu, Y., 2004. Inhibition of free ammonia to the formation of
aerobic granules. Biochemical Engineering Journal 17, 41e48.
Zhang, J., Chua, H.C., Zhou, J., Fane, A.G., 2006. Factors affecting the membrane
performance in submerged membrane bioreactors. Journal of Membrane Science 284, 54e66.