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Development of granular sludge for textilewastewater treatment
Khalida Muda a,*, Azmi Aris a, Mohd Razman Salim a, Zaharah Ibrahim b, Adibah Yahya b,Mark C.M. van Loosdrecht c, Azlan Ahmad a, Mohd Zaini Nawahwi b
aDepartment of Environmental Engineering, Faculty of Civil Engineering, Universiti Teknologi Malaysia, 81310, Skudai, Johor, MalaysiabDepartment of Biological Sciences, Faculty of Biosciences and Bioengineering, Universiti Teknologi Malaysia, 81310, Skudai, Johor, MalaysiacDepartment of Biotechnology, Delft University of Technology, Julianalaan 67, 2628BC Delft, The Netherlands
a r t i c l e i n f o
Article history:
Received 18 January 2010
Received in revised form
22 April 2010
Accepted 2 May 2010
Available online 25 May 2010
Keywords:
Granulation
Granule characterization
Textile wastewater
Sequencing batch reactor
Color removal
* Corresponding author. Tel.: þ60 607 553152E-mail address: [email protected] (K. Mud
0043-1354/$ e see front matter ª 2010 Elsevdoi:10.1016/j.watres.2010.05.023
a b s t r a c t
Microbial granular sludge that is capable to treat textile wastewater in a single reactor
under intermittent anaerobic and aerobic conditions was developed in this study. The
granules were cultivated using mixed sewage and textile mill sludge in combination with
anaerobic granules collected from an anaerobic sludge blanket reactor as seed. The gran-
ules were developed in a single sequential batch reactor (SBR) system under alternating
anaerobic and aerobic condition fed with synthetic textile wastewater. The characteristics
of the microbial granular sludge were monitored throughout the study period. During this
period, the average size of the granules increased from 0.02 � 0.01 mm to 2.3 � 1.0 mm and
the average settling velocity increased from 9.9 � 0.7 m h�1 to 80 � 8 m h�1. This resulted in
an increased biomass concentration (from 2.9 � 0.8 g L�1 to 7.3 � 0.9 g L�1) and mean cell
residence time (from 1.4 days to 8.3 days). The strength of the granules, expressed as the
integrity coefficient also improved. The sequential batch reactor system demonstrated
good removal of COD and ammonia of 94% and 95%, respectively, at the end of the study.
However, only 62% of color removal was observed. The findings of this study show that
granular sludge could be developed in a single reactor with an intermittent anaero-
biceaerobic reaction phase and is capable in treating the textile wastewater.
ª 2010 Elsevier Ltd. All rights reserved.
1. Introduction granules as seeding material for aerobic granules develop-
The development of aerobic granules as a novel treatment
technology for wastewater has been extensively reported
using sequencing batch reactor (SBR) systems. The systemhas
been used to treat different types of wastewater and pollu-
tants such as dairy effluent (Arrojo et al., 2004), soybean-pro-
cessing wastewater (Su and Yu, 2005), nitrogen and
phosphorus-rich effluent (Cassidy and Belia, 2005), phenol
effluent (Carucci et al., 2009) and also municipal wastewater
(de Kreuk and van Loosdrecht, 2006). The use of anaerobic
2/81; fax: þ60 607 556615a).ier Ltd. All rights reserved
ment has recently been reported by Linlin et al. (2005).
In recent years, the ability of biodegradation for textile
dyeing wastewater and dyestuffs involving both anaerobic
and aerobic processes has been widely reported in the litera-
ture (Ong et al., 2005; Isik and Sponza, 2008; Franciscon et al.,
2009). Color removal and complete mineralization of the dyes
have been achieved through the combination of both
processes. For azo dyes, the cleavage of N]N bond, which
results in the removal of color, occurs during anaerobic stage
along with the generation of aromatic amines, a toxic
7.
.
Sampling point
DO probe
pH probe
Effluent
Influent
Influent tank
Air compressor controlled by a
timer
Effluent tank
Peristaltic pumps controlled by timers
Mass-flow controller
Fig. 1 e Schematic layout of the SBR reactor system.
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 04342
compoundwhich is detrimental to human health (van der Zee
and Villaverde, 2005). Under aerobic condition, mineralization
of the amines takes place to complete the treatment process.
Several studies have been carried out using anaerobic
granular sludge but this doesn’t lead to complete removal of
the dyes. Study conducted using granular sludge grown in
anaerobic/aerobic system for complete dye removal is
apparently missing.
Since complete dye degradation requires both anaerobic
and aerobic conditions, several studies have been focused on
treatment systems which utilized two different reactors to
fulfill both conditions (Ong et al., 2005; Moosvi and
Madamwar, 2007; Isik and Sponza, 2008). The operation of
such a system is rather complicated as the anaerobic micro-
organisms in the anaerobic tank need to be separated before
the wastewater can be pumped to the aerobic tank. To
simplify the system, a study was conducted to develop
microbial granular sludge that can survive and function in
both anaerobic and aerobic conditions and hence requires
only one reactor. Potential strict anaerobic microorganisms
can survive easily since oxygen only penetrates partially in
the granules during aerated phase of the process. The study
focused on the development of this type of granular sludge
and the effectiveness of the system in treating synthetic
textile dyeing wastewater.
2. Materials and methods
2.1. Wastewater composition
Synthetic wastewater with the following composition was
used: NH4Cl 0.16 g L�1, KH2PO4 0.23 g L�1, K2HPO4 0.58 g L�1,
CaCl2$2H20 0.07 g L�1, MgSO4$7H2O 0.09 g L�1, EDTA
0.02 g L�1and trace solution 1 mL L�1. The carbon sources used
in this experiment were glucose (0.5 g L�1), ethanol (0.125 g L�1)
and sodium acetate (0.5 g L�1). The trace elements used were
based on the composition recommended by Smolders et al.
(1995). The composition of the trace element was H3BO3
(0.15 g L�1), FeCl3$4H2O (1.5 g L�1), ZnCl2 (0.12 g L�1),MnCl2$4H2O
(0.12 g L�1), CuCl2$2H2O (0.03 g L�1), NaMoO4$2H2O (0.06 g L�1),
CoCl2$6H2O (0.15 g L�1), and KI 0.03 g L�1. Mixed dyes consisted
of Sumifix Black EXA, Sumifix Navy Blue EXF and Synozol Red
K-4B with total concentration of 50 mg L�1 was used in this
study. The mixture gave an initial COD of 1270 mg L�1; 1020
ADMI (American Dye Manufacturing Index) and average
ammonia concentration of 38 mg L�1. The pH of the synthetic
wastewater was adjusted to 7.0 � 0.5 before feeding.
2.2. Reactor set-up
The schematic representation of the reactor set-up is given in
Fig. 1. A column reactor was designed based on Wang et al.
(2004) and Zheng et al. (2005) with several modifications. The
column was designed for a working volume of 4 L with
internal diameter of 8 cm and total height of 100 cm. The
wastewater was fed into the reactor from the bottom of the
column. Air was supplied into the reactor by a fine air bubble
diffuser also from the bottom of the column. The decanting of
the wastewater took place via an outlet port located at 40 cm
height from the bottom of the reactor. The mean cell resi-
dence time (SRT) was set by the discharge of suspended solids
with the effluent.
2.3. Analytical methods
The morphological and structural observations of the
granules were carried out by using a stereo microscope
equipped with digital image processing and analyzer ((PAX-
ITv6, ARC PAX-CAM)). The microbial compositions within the
granules were observed qualitatively with scanning electronic
microscope (FESEM-Zeiss Supra 35 VPFESEM). The granules
were left to dry at room temperature prior to gold sputter
coating (Bio Rad Polaron Division SEM Coating System) with
coating current of 20 mM for 45 s. Other parameters such as
mixed liquor suspended solid (MLSS), mixed liquor volatile
suspended solid (MLVSS), COD, color and NH3 were analyzed
according to the Standard Methods (APHA, 2005).
The granules developed in the SBR column were analyzed
for their physical, chemical and biological characteristics.
Physical characteristics include settling velocity, sludge
volume index (SVI) and granular strength. The settling
velocity was determined by averaging the time taken for an
individual granule to settle at a certain height in a glass
column filled with tap water. The SVI assessment was carried
out according to the procedure described by Beun et al. (1999).
Determination of the granular strength was based on
Ghangrekar et al. (2005). Shear force on the granules was
introduced through agitation using an orbital shaker at
200 rpm for 5 min. At certain amplitude of the shear force,
parts of the granules that are not strongly attached within the
granules were detached. The quantity of the ruptured gra-
nules was separated by allowing the fractions to settle for
1 min in a 150 ml measuring cylinder. The dry weight of the
settled granules and the residual granules in the supernatant
weremeasured. The ratio of the solid in the supernatant to the
Fig. 2 e Change in biomass concentration during the
formation of granular sludge in the SBR. (C) MLSS, (,)
MLVSS, (A) Suspended solid in the effluent.
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 0 4343
total weight of the granular sludge used for granular strength
measurement was expressed in percentage as an integrity
coefficient (IC). This percentage indirectly represents the
strength of the granules. Smaller IC value indicates stronger
granule and vice versa.
The granules were analyzed chemically for their mineral
content which includes Ca2þ, Mg2þ, Naþ, Kþ and Fe2þ. Themineral content was determined using Perkin Elmer Analyst
400 Flame Atomic Absorption Spectrophotometer (FLAA). The
microbial activity of the microbial granular sludge was con-
ducted by measuring the oxygen utilization rate (OUR),
specific oxygen utilization rate (SOUR) and specific methano-
genic activity (SMA). The OUR (mg.O2 L�1 h�1) measurement
was performed according to the Standard Methods (APHA,
2005). The SMA measurements were conducted according to
Erguder and Demirer (2008) with several modifications where
a 500 mL BOD bottle seeded with FSG with final concentration
of 1e2 g VSS L�1 and basal medium (250 mL effective volume).
The bottle was flushed with N2 gas mixture for 5 min to obtain
an anaerobic condition. The bottle was then sealed with
rubber septum. Acetic acid (HAc) was fed into the serumbottle
at a concentration of 3000 mg L�1. The experiments were
conducted in a temperature controlled condition of 30 � 2 �C.The production of methane gas (CH4) was determined
according to Erguder et al. (2001). The production of methane
gas (CH4) was determined daily for 5e7 days using liquid
displacement methods containing concentrated KOH stock
solution (20 g L�1). After each gasmeasurement, the bottlewas
manually shaken. At the end of the SMA assay, the VSS
content in the bottle was measured. The SMA was calculated
as the maximum CH4 produced per gram of VSS per hour
(mL CH4/g�1 VSS h�1) (Zitomer and Shrout, 1998).
2.4. Experimental procedures
A mixture containing an equal volume of sludge from
a municipal sewage treatment plant and a textile mill
wastewater treatment plant that gave a total volume of 2 L
was used in this experiment. The sludge inoculums were
sieved with amesh of 1.0 mm to remove large debris and inert
impurities. The sludge mixture was acclimatized for two
months with 2 L of synthetic wastewater containing dye
degrading microbes. The dye degrading microbes used in this
study was based on the study conducted by Nawahwi (2009)
and Ibrahim et al. (2009). Together with the sludge mixture,
about 100 mL of anaerobic granules with size less than 1 mm
diameter were used as seed for the granulation process. The
anaerobic granules were collected from an anaerobic sludge
blanket reactor system treating paper mill industrial effluent.
The MLSS of the anaerobic granules were 3.3 g L�1.
During the start-up period, 2 L of mixed sludge and 2 L of
synthetic textile wastewater were added into the reactor
system making the final volume of 4 L with total sludge
concentration after inoculation of 5.5 g L�1. The system was
supplied with external carbon sources consist of glucose,
sodium acetate and ethanol which gave a substrate loading
rate of 2.4 kg COD m�3 d�1. The reactor was operated in
successive cycles of 6 h. Each cycle comprised of 5 min filling,
340 min reaction, 5 min settling, 5 min decanting and 5 min
idle. The reaction phase started with an anaerobic phase of
40min, followed by aerobic phase for 130min, another 40 min
of second anaerobic phase and 130 min for second aerobic
phase. The dissolved oxygen (DO) concentration remained
low during the anaerobic condition (0.2 mg L�1) and reached
saturation concentration during the aerobic phase. The
superficial air velocity during the aerobic phase was
1.6 cm s�1. The pH during the reaction process varied in the
range of 6.0e7.8 and the temperature of the reactor system
was set at 30 � 2 �C. The reactor system was operated for
a period of 66 days. Two liters of the wastewater remained in
the reactor after the decanting stage yielding a volumetric
exchange rate (VER) of 50%. 20 mL of sample from the influent
and effluent (wastewater released after decanting stage) of the
reactor system were collected for the measurement of COD,
ammonia and color removal (Fig. 1).
3. Results and discussions
3.1. Biomass profile
The change in biomass concentration (i.e. MLSS) from the
start-up until the end of the study is shown in Fig. 2. During
the first few days of the experiment, almost half of the sludge
was washed-out from the reactor causing a rapid decrease in
the biomass concentration. The MLSS reduced from initial
concentration of 5.5 g L�1 to 2.9 g L�1mainly due to the short
settling time used in the cycle (i.e. 5 min). During this initial
stage, the anaerobic granules were also observed to disinte-
grate into smaller fragmented granules and small debris
resulted from shear force caused by the aeration during the
aerobic stage. These small fragments have poor settling ability
and were washed out from the reactor. This caused an
increase of the suspended solids concentration in the effluent
as shown in Fig. 2. As the experiment continued and granules
with adapted biomass were formed, the concentration of the
biomass in the reactor increased and finally reached
7.3 � 0.9 g MLSS L�1 when the experiment was discontinued
on the 66th day. The MLVSS followed the same trend as MLSS,
Fig. 3 e The changes of dissolved oxygen and oxygen uptake rate in one complete cycle of the SBR reactor system (A) DO,
(,) OUR at day 66th of the experiment. (PI) First aerobic phase, (PII) First anaerobic phase, (PIII) Second aerobic phase, (PIV)
Second anaerobic phase.
Table 1 e The OUR levels during the aerobic reactionphase of one complete cycle.
Aerobic reactionphase
OUR (mg L�1 h�1)
1st stage (PII) 2nd stage (PIV)
Begin react 281 � 39 167 � 51
End react 14 � 2 11 � 2
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ranging from 1.9 � 0.5 g L�1 to 5.6 � 0.8 g L�1. The SRT also
increased from 1.4 days at the initial stage to 8.3 days on the
66th day, indicating the accumulation of the biomass in the
reactor.
3.2. Bioactivities of the granules
A typical DO concentration profile for one complete cycle and
the OUR profile during both of the aerobic reaction phases are
shown in Fig. 3. Stage PI and PIII show the first and second stage
of anaerobic reaction phase, respectively. At these anaerobic
stages, most of the dye degradation is expected to occur where
amines, as the byproduct,were released (Sponzaand Isik, 2005).
Stage PII and PIV represent the first and second stage of aerobic
reaction phase, respectively. Most of the substrates provided to
the reactor system were anticipated to be consumed within
a fewminutes of the first aerobic reaction phase (PII), known as
the feast period. During the feast period, the DO concentration
in the reactor was low (about 4 mg L�1). The high utilization of
DO during the feast period was also indicated by the high OUR
which was 281 mg L�1 h�1. The amines, which were produced
during anaerobic reaction phase (PI), were mineralized under
this aerobic condition (PII) as they cannot be further degraded
under anaerobic phases (Sponza and Isik, 2005).
When all the carbon sources (substrate and amines) in the
wastewater were utilized, endogenous respiration process
took place, which is referred as the famine period. The tran-
sition from the feast to famine phase was clearly observed
with the drastic increase of the dissolved oxygen and the
extreme drop of the OUR within few minutes of the aerobic
reaction phase (PII). The DO concentration immediately
increased to around 7.0 mg L�1 which was closed to the DO
saturation level. The OUR also reduced to 14 mg L�1 h�1
indicating low utilization of DO.
Since there was no addition of substrate during the second
aerobic reaction phase (PIV), the consumption of DO during
this phase was also low. This is shown by high DO level
reaching saturation value of 7.6 mg L�1. A sharp increase in
the OUR was observed at the beginning of this phase but at
a lower value than the one observed in Stage PII. Apparently,
the residual dyes which were not degraded in Stage PI and PII
were transformed into smaller molecules (e.g. amines) during
the second stage of the anaerobic phase (PIII). These smaller
molecules were further mineralized in Stage PIV which
resulted in a sharp increase in the OUR. As the concentration
of these molecules were reduced, the OUR also became lower
until it reached aminimumof 11mg L�1 h�1 Table 1 shows the
OUR value during both of the aerobic reaction phases.
The SOUR of themicrobial granular sludgewas determined
before the termination of the experiment. The SOUR was
51.1 � 6.8 mg DO g�1 VSS h�1. This value was slightly lower
than those of the aerobic granules reported by Tay et al. (2001)
which ranged from 55.9 to 69.4 mg DO g�1 VSS h�1and higher
than the coupled granules reported by Erguder and Demirer
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 0 4345
(2005) (6e47 mg DO g�1 VSS h�1). The specific methanogenic
activity (SMA) of the microbial granular sludge is lower
(10.3 mL CH4 g�1 VSS h�1) than the one reported by Erguder
and Demirer (2005) (14e42 mL CH4 g�1 VSS h�1). However,
despite the low SMA emission, it provides the evidence of the
existence of methanogens within the microbial granular
sludge. Obviously, the granular sludge offered the metha-
nogens sufficient protection from the toxic oxygen concen-
tration in the bulk liquid.
3.3. Morphology of granular sludge
A week after inoculating the reactor, visual and microscopic
observations of granule formation were made. At this stage,
the developed granules were composed mostly of loosely
clumped sludge which could easily break up into pieces if
placed under vigorous shaking. Within a week, the anaerobic
seed granules had undergone morphological changes from
spherical in shape and black in color with average diameter of
1mm into smaller grey granules. It is likely that the sulfides in
the anaerobic sludge were oxidized due to exposure to the
aerobic condition. On day 30, two different types of granules
were clearly observed in the reactor as shown in Fig. 4.
Fig. 4a shows mainly irregular shaped, yellow colored
granules (Type A) that are solely developed from the activated
sludge. In Fig. 4b, the anaerobic granules that have frag-
mented into smaller pieces have formed different sizes of
granules (Type B) that contained pieces of anaerobic granules.
The outer layer of the latter were yellow in color indicating the
Fig. 4 e Morphological development of granular sludge from an
Pictures were taken using stereo microscope with magnification
from the activated sludge. (b) Granules developed from anaerobi
experiment. (d) Microbial granular sludge at the 66 days of the
domination of aerobic or facultative microorganisms while
the darker spots within the granules indicate the presence of
anaerobic fragments originated from anaerobic granules. The
formation of Type A granules could be elucidated by the
mechanisms explained by Beun et al. (1999). The development
was initiated from the mycelial pellets that were retained in
the reactor due to high settling velocity. Thesemycelial pellets
eventually become the supportmatrix for the bacteria growth.
Bacteria that were able to attach to this matrix were retained
and suppressed the filamentous growth and became the
dominant species in the reactor.
The formation of Type B granules has been discussed by
Linlin et al. (2005). These granuleswere formed through a series
of physical and morphological changes. The anaerobic gra-
nules initially disintegrated into smaller size flocs and debris
when exposed to aeration forces in the SBR column. Some of
the granules and debris that were too small were washed-out
with the effluent while the heavier ones were retained in the
column and acted as nuclei for the formation of the aerobic
granules. This type of granules that consisted of combination of
aerobic and anaerobic portions within the granules could
increase the possibility of degradation process that requires
both aerobic and anaerobic conditions for complete degrada-
tion particularly for textile wastewater that contains azo-dyes.
Fig. 4c shows the sludge particles during the initial stage of the
experimentwith an average size of 0.02� 0.01mmwhile Fig. 4d
shows the granules at the final stage (day 66) of the experiment
with the average particle diameter size of 2.3 � 1.0 mm with
maximum size reaching up to 4 mm.
aerobic granular sludge and aerobic activated sludge.
of 6.33, Scale bar equals to 1 mm. (a) Granules developed
c granules patches. (c) Sludge particles during early stage of
experiment.
Fig. 5 e SEM microstructure observation on mature microbial granular sludge at a magnification of 10,000 K. (a) Coccoid
bacteria tightly linked to one another. (b) Cavities that appear between bacteria clumped inside the granules.
Fig. 6 e The relationship between the settling velocity of
the microbial granular sludge and the biomass
concentration retained in the reactor. (,) Settling velocity
(C) Biomass concentration.
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 04346
The microstructure of the microbial granular sludge was
examined using SEM and shown in Fig. 5. The SEM observation
of the mature granules shows the domination of non-fila-
mentous coccoid bacteria that are tightly linked and embedded
to one another and form a rounded shape on the surface of the
granule, covered with extrapolysaccarides (EPS) (Fig. 5a). The
absence of filamentous bacteria in the developed granulesmay
be due to the experimental conditions that did not favor the
growth of filamentous bacteria at such high concentration of
DO during aerobic phase (i.e. 7.0� 0.5mg L�1) and considerably
high organic loading rate (2.4 kg COD m�3 d�1) (Chudoba, 1985;
Zheng et al., 2006). Fig. 5b shows the presence of cavities
between the clumpedbacteria. These cavities are anticipated to
be responsible in allowing smooth mass transfer of substrates
or metabolite products in and out of the granules.
3.4. Physical characteristics of granular sludge
3.4.1. SizeThe shear force imposed in the development of granules in
this experiment, in terms of superficial upflow air velocity
which was 1.6 cm s�1, resulted in the development of micro-
bial granular sludge with average diameter of 2.3 � 1.0 mm.
According to Peng et al. (1999), the diameter of the developed
aerobic granules is in the range of 0.3e0.5 mm which is much
smaller as compared to anaerobic and anoxic granules that
could develop up to 2 to 3 mm. The strong shearing force
produced by mixing and aeration during the reaction phase
could prevent the development of bigger granules which can
be achieved in an anaerobic system (van Benthum et al., 1996;
Kwok et al., 1998).
3.4.2. Settling velocityThe average settling velocity of the seed sludge and seed
anaerobic granular sludge was 9.9 � 0.7 m h�1 and
42 � 8 m h�1 respectively. The average settling velocity of the
anaerobic granular seed is in accordance with those reported
by Schmidt and Ahring (1996) which is in the range of
18e100 m h�1. The average settling velocity of the granular
sludge developed in this study increased from 17.8� 2.6m h�1
to 80 � 8 m h�1 at the end of experiment. The settling velocity
obtained in this study is almost three times greater than the
settling velocity of the aerobic granules reported by Zheng
et al. (2005) (i.e. 18e31 m h�1).
The increase in settling velocity has given significant
impact on the biomass concentration in the reactor. The
relationship between settling velocity of the granules and the
concentration of the MLSS is shown in Fig. 6. Despite the short
settling time (5 min), the high settling velocity possessed by
the developed microbial granular sludge enabled the granules
to escape from being flushed out during the decanting phase.
Such conditions have caused more microbial granular sludge
to be retained in the system and resulted in the increase of
biomass concentration in the reactor. In this experiment, the
SRT value was 1.4 days during the start-up (partly low due to
wash out of inoculated sludge) and rose up to 8.3 days at the
end of experiment. As less biomass was washed-out during
the decanting period, the increase in SRT is another mani-
festation of good settling characteristics resulting from the
high settling velocity.
The SVI value has improved from 276.6mL g�1 at the initial
stage of the experiment to 69 mL g�1 at the end of the
experiment indicating the good settling properties of the
Fig. 7 e The SVI and mean cell residence time (SRT) profile.
(B) SVI, (-) SRT.
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 0 4347
granules which is favorable in wastewater treatment plant
operation. The change of the SVI value and the SRT as
a function of time are given in Fig. 7. The SVI value achieved in
this experiment is in agreement with the result reported by
McSwain et al. (2004) with SVI values of 115 � 36 mL g�1
(settling time 10 min) and 47 � 6 mL g�1 (settling time 2 min).
The higher settling velocity and lower SVI value of the mature
microbial granular sludge as compared to previous reports by
other researchers indicate that the formation of granules
seeded with anaerobic granules would develop better settling
properties of the granules. It may also be due to the specific
reaction condition of anaerobic/aerobic setting in the experi-
ment that induced the well settling of the granule.
3.4.3. Granular strengthThe granular strength of the granules was measured based on
the integrity coefficient (IC) as described by Ghangrekar et al.
(2005). The smaller the IC value, the higher the strength and
ability of the granules to remain as high structural integrity
granules during aeration phase that caused the shear force.
Fig. 8 e The change in integrity coefficient (representing
the granular strength) during the formation of granular
sludge in the anaerobic/aerobic SBR.
Fig. 8 shows the IC profile of the developed granular sludge as
a function of time. The IC reduces as the granules developed.
With an initial value of 30 � 0.3, the IC was reduced to about
9.4 � 0.5 at the termination of the experiment. A sharp
reduction of IC was observed after 40 days of the experimental
run. According to Ghangrekar et al. (2005) granules with
integrity coefficient of less than 20 were considered high
strength granules. The reduction in IC value indicates the
increase in the strength of the bond that holds the microor-
ganisms together within the developed granules.
During the early stage of the granule development, the
microbes within the granules were loosely bounded to each
other. When the microbes were loosely linked together, the
granules may contain more cavities which make the granules
less dense, as shown by low settling velocity value. As more
microbes were linked together, the granules increased in size.
Under the applied selective pressures (i.e. short settling time,
hydrodynamic shear force, feast-famine regime) within the
reactor, microbes may produce more EPS (Qin et al., 2004). As
reported by Adav et al. (2008), the EPS could contribute greatly
to the strength and the stability of anaerobic granules. When
more EPS are being produced by the microbial cells, they form
a cross-linked network and further strengthen the structural
integrity of the granules. The cavities within the granulesmay
be filled with the EPS as it is a major component of the bio-
granule matrix material in both anaerobic and aerobic gra-
nules. This caused the granules to become denser and
stronger as shown by their high settling velocity and low IC
value at the end of the experiment.
3.4.4. Mineral contentThe concentration of minerals in granular sludge, newly
developed and matured granular sludge were determined in
mg/g of dry sludge and presented in Table 2. The concentra-
tion of Naþ and Kþ are not much different in the sludge, newly
developed and matured granules. However, there is an
obvious increase in the concentration of Ca2þ andMg2þwithin
the matured granules. The concentration of Fe2þ was slightly
reduced in the matured and newly developed granules as
compared to the sludge.
Basically, an unchanged concentration of Naþ and slightly
decreasing Kþ concentration in the sludge and matured
granulesmay indicate that thesemonovalent cations may not
be involved in the granulation process. It has been reported
that high concentration of the Naþ and Kþ may cause adverse
Table 2 e Comparison of mineral content at differentstages during the development of microbial granularsludge.
Mineral contents (mg g�1 of dry sludge)
Mineral Sludge Newly developedmicrobial granularsludge (1 week)
Matured microbialgranular sludge
(10 weeks)
Ca2þ 1.53 � 0.02 2.0 � 0.6 4.65 � 0.04
Mg2þ 0.13 � 0.01 0.322 � 0.003 1.75 � 0.08
Naþ 0.22 � 0.07 0.25 � 0.04 0.24 � 0.05
Kþ 1.31 � 0.07 1.15 � 0.03 0.93 � 0.05
Fe2þ 2.32 � 0.02 1.90 � 0.04 1.98 � 0.08
Fig. 9 e Percentage removal for (a) COD, (b) ammonia and (c)
color of the SBR system. (A) Influent, (B) Percentage
removal, (-) Effluent. (ADMI) American Dye Manufacturing
Units.
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 04348
effect on the granules formation. It could cause reduction in
sludge concentration, settling velocity of the sludge, granular
strength and treatment efficiency (Ghangrekar et al., 2005).
The monovalent cation, Naþ has also been reported causing
detrimental impact on the flocculation system (Sobeck and
Higgins, 2002). Nevertheless, there are contradictory reports
related to the effect of these monovalent cations in granula-
tion development processes. Fernandez et al. (2008) reported
that the concentration of granular biomass has improved
significantly in the reactor system fedwith high concentration
of inorganic salt influent. A quick decrease of solids concen-
tration in the effluent was observed after the addition of NaCl.
The development of the granular sludge in this study
showed higher accumulation of Ca2þ and Mg2þ towards the
end of the experiment. This may indicate the involvement of
the inorganic elements in the granulation process. Based on
the divalent cation bridging theory, the presence of Ca2þ and
Mg2þ promotes equivalent floc properties (Sobeck and
Higgins, 2002). According to Ren et al. (2008), granule-rich
Ca2þ showed more rigid granular structure and higher
strength as compared to granule without Ca2þ accumulation.
3.5. Removal performance
The performance of the reactor system from start-up until the
end of granules development period based on the removal of
COD, color and ammonia is given in Fig. 9. At the initial stage
of the operation, the percentage removal for COD and
ammonia was 71% and 67% respectively (Fig. 9a and b). The
removal efficiency increased to 94% for COD and 95% for
ammonia at the end of the experiment. The increase in the
removal efficiency indicates the occurrence of high biological
activity in the reactor system. During the first month, the
removal efficiency for COD and ammonia fluctuated but the
removal became stable for the remaining period. The removal
efficiency for color was fluctuating almost throughout the
study period (Fig. 9c). The percentage of color removal was
about 25% during the start up and increased to 62% at the end
of the experiment. The average of color removal was 55%. This
low percentage of the color removalmay be due to insufficient
adaptation time. As dye substances are recalcitrant and
difficult to be degraded, more time is required to accumulate
sufficient organisms which degrade the dyes in the reactor.
The inconsistent percentage for color removal may also be
contributed by the unstable condition of the aromatic amines,
the byproduct of dye degradation which easily oxidized and
recolorwhen exposed to oxygen during the aerobic phase. The
increase of color during autoxidation of aromatic amines was
confirmed by several researchers (Cruz and Buitron, 2001;
Libra et al., 2004; Sponza and Isik, 2005).
The inconsistency of color removal may be also influenced
by the absorption of color into sludge biomass throughout the
experiment. The absorption of color into the sludge biomass
has been reported by other researchers (Otero et al., 2003;
Wang et al., 2006; Sirianuntapiboon and Srisornsak, 2007).
Fig. 10 shows the percentage removal of COD, ammonia
and color in a complete 340-min reaction phase of the SBR
system recorded on the 66th days of experiment. The profile
and the percentage removal for COD and ammonia were
almost the same while the removal of color was much lower.
After 340 min of intermittent anaerobic and aerobic modes,
about 93%, 95% and 62% of COD, ammonia, and color
respectively were removed.
During the first anaerobic phase (PI) (0e40 min), approxi-
mately 15% and 4% of COD and ammonia respectively, were
removed. In the first aerobic phase (PII) (40e170 min), about
68% of the COD was removed while 80% of the ammonia was
oxidized. Most of the organic compounds and nitrification of
ammonia were achieved during this stage. The supply of
oxygen at this stage enabled good oxidation of these
compounds (Brauer and Henning, 1986). The nitrate produced
Fig. 10 e The removal for COD, ammonia and color in one
complete cycle of the SBR system. (-) Color, (B) COD, (:)
Ammonia. (PI) First aerobic phase, (PII) First anaerobic
phase, (PIII) Second aerobic phase, (PIV) Second anaerobic
phase.
wat e r r e s e a r c h 4 4 ( 2 0 1 0 ) 4 3 4 1e4 3 5 0 4349
will be removed through the denitrification process that will
occur in the second stage of anaerobic phase. The second
anaerobic phase (PIII) (170e210 min) showed only about 5% of
COD and ammonia being removed while the remaining
(about 4%) was removed in the second aerobic phase (PIV)
(210e340 min). As for color, about 45% and 16% were removed
during anaerobic and aerobic phases respectively. The result
shows the ability of anaerobic microbes within the granule to
degrade the dye. The high percentage for color removal indi-
cated active cleavage of dye compound took place during both
of the anaerobic phases (PI and PIII).
The degradation and decolorization of dye during anaer-
obic condition has been widely reported in the literatures (dos
Santos et al., 2007). In anaerobic condition, the electrons from
the electron donor are transferred to the N]N bond of the azo
dye causing the cleavage of the bond forming aromatic
amines. The amines are then degraded under aerobic condi-
tion reducing the COD value of the wastewater. In addition to
the degradation mechanism, dye removal may also occur via
adsorption onto the biomass (Aksu, 2001 and Crini, 2006).
Amines, the colorless byproduct of anaerobic degradation of
dye compound are unstable compound that could easily be
oxidized during the presence of oxygen. These autoxidation of
the amines may form different intensity colored compound
(Cruz and Buitron, 2001; Libra et al., 2004; Sponza and Isik,
2005). This reaction may cause the reduction on the overall
percentage of color removal during aerobic condition (PII and
PIV). Based on the removal performance of the system, it has
been proven that the developed microbial granular sludge is
capable to perform the degradation process during anaerobic
and aerobic phases. This indicates the presence of aerobic,
facultative and anaerobic microorganisms in the microbial
granular sludge. According to Li and Liu (2005), when the
granules grew to a size larger than 0.5 mm, the diffusion of
oxygen into the inner part of the granules became a limitation.
This may give an indication of the presence of anaerobic
microorganisms within centre part of the microbial granular
sludge since the average size of microbial granular sludge
developed in this study was 2.3 � 1.0 mm. Aerobic microor-
ganismsmay be found at the outer layer of the granuleswhich
can easily access the oxygenmolecule andmainly responsible
for the COD removal. The facultative microorganisms may be
found in any part of the microbial granular sludge due to its
capability to live both under anaerobic and aerobic condition.
4. Conclusion
� Stablemicrobial granular sludgecouldbecultivated inasingle
SBRsystemwiththeapplicationof intermittentanaerobicand
aerobic reactionmode during the reaction phase.
� The matured granules showed the domination of non-fila-
mentous bacteria that were tightly linked and embedded to
one another and covered with EPS. The SVI value of the
biomass has decreased from 276.6 mL g�1 to 69 mL g�1at the
end the 66 days, also indicating the excellent settling
properties of the granules.
� The development of the granular sludge is positively
correlated with the accumulation of divalent cationic Ca2þ
and Mg2þ in the granules suggesting the role played by the
cations in the granulation process.
� The results indicate the viability of the single reactor system
for treating textile wastewater under intermittent anaerobic
and aerobic phase strategy.
� The OUR/SOUR and SMA analyses indicate the presence of
anaerobic and aerobic microorganisms activities in the
granular sludge which is capable to perform degradation
process both in anaerobic and aerobic conditions.
Acknowledgements
The authorswish to thank theMinistry of Science, Technology
and Innovation (MOSTI), Ministry of High Education (MOHE)
and Universiti Teknologi Malaysia for the financial supports of
this research (Grants No.: 79137, 78211 and 75221).
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