37

“Coming together is a beginning, keeping together is progress, and

working together is success.”

____ Henry Ford

CHAPTER-3 Aerobic Granulation

38

3.1 Cell Immobilization

Major wastewater treatment facilities based on biological process facing the

problem of biomass loss further they have limitation like;

They cannot sustain high organic load

Unable to deal with any fluctuation in organic load

Biomass retention

Surplus biomass production

Large area requirement \low volumetric conversion capacity (Loosdrecht et

al., 2002)

Disposal of sludge produced in one of the major problems facing by the

wastewater treatment process especially based on aerobic biodegradation. Sludge can

be spread on soil or landfill. For the localities near the shores the sludge is usually

dumped into ocean however in near future all kind of ocean dumping should be

stopped. The use of sludge as fertilizer on crop lands in a very regular practice but the

presence of toxic chemicals in the sludge is of growing concern. There are practices

of incineration of sludge but here also there in risk of contamination of air by the

chemical present in sludge. In order to develop a process that can overcome the

limitations of traditional biological wastewater treatment process many strategies

were proposed by wastewater engineers across the globe. Cell immobilization

technology is one of them. It is a strategy that has been proposed to overcome the

substrate inhibition difficulties associated with high-strength wastewater and to

enhance the biomass retention. It can withstand higher organic load and fluctuations

in organic load (Keweloh et al., 1989). Because of the dense packing of cells in case

of immobilised cell systems, high volumetric reaction rates are achieved. In bioreactor

operation, the use of cell biomass immobilized on surfaces or on particulate carriers

ensures efficient biomass retention, minimizing cell loss through washout. This also

minimizes the biomass discharge in downstream treated wastewater discharge for

further processing (Venugopalan et al., 2005).

Microorganisms present in an immobilized system are very much resistant (10

to 100 times) to high concentration of toxic compounds like, biocides, antibiotics or

39

other chemicals, when compared to their freely suspended form (Venugopalan et al.,

2005). This enables the immobilized cells to tide over unfavorable environmental

conditions in a very efficient manner. One of many advantages of Immobilized

systems over other suspended process is that it can be repeatedly re-used.

The cells can be engineered ones, naturally occurring strains or a combination of

both. Similarly, the immobilized systems can be artificial or natural. In artificial

systems, the cells are entrapped in suitable gel matrices (e.g. calcium alginate

hydrogels), where they retain most of their viability and physiological activity. On the

other hand, spontaneous adsorption of microbial cells to surfaces of carrier material

results in colonization of the support material and leads to the formation of what is

known as biofilms. In a biofilm, the micro-organisms are entrapped within a matrix of

extra-cellular polymeric substances (EPS) secreted by themselves (Venugopalan et

al., 2005). Formation of surface-associated biofilms is a universal survival strategy

adopted by bacteria. Some of the most commonly used immobilization techniques

include use of cross-linked gelatin, porous ceramic beads and alginate or agarose

beads.

3.2 Biogranulation

Bacteria also have the ability to attach to one another and form self-

immobilized granules. When immobilized with respect to each other by co-

aggregation or auto-aggregation or by other means, it is called as “Biogranulation”.

These granules are dense, compact aggregates and consist of a consortium of different

microorganisms held together in a common polymer matrix (Fig-3.1). Granule

development is mediated by auto aggregating or co-aggregating abilities of the

various interacting bacteria. Auto-aggregation is one type of cell-to-cell interaction

for one genetically identical strain. Co- aggregation is defined as cell-to-cell

adherence between genetically distinct bacterial partners. Microbial co-aggregation

was first recognized between bacteria isolated from human dental plaque (Gibbons,

1970). Granular sludge was first time reported in 1976 in a pilot plant at CSM sugar

factory in Breda, Netherlands. Granular sludge is a dense microbial community that

typically includes millions of organisms per gram of biomass. None of the individual

species in these micro ecosystems is capable of completely degrading the influent

40

wastes. Complete degradation of industrial waste involves complex interactions

between the resident species. Thus, granular sludge reactors are desirable in

wastewater biological treatment processes because a very high number of organisms

can be maintained in the bioreactor. This in turn implies that contaminant

transformation is rapid and highly concentrated; therefore, large volumes of waste can

be treated in compact bioreactors. In granular sludge reactors, the large size and

relatively high density of individual granules causes them to settle rapidly, which

simplifies the separation of treated effluent from the biomass. Granular sludge has

proved capable of treating high-strength wastewater contaminated with soluble

organic pollutants.

A (Our results) B (Ivanov et al., 2006)

Figure 3.1 Aerobic granules

Biogranulation involves cell-to-cell interactions that include biological,

physical and chemical phenomena. These granules are dense microbial consortia

packed with different bacterial species and typically contain millions of organisms per

gram of biomass. These bacteria perform different roles in degrading the complex

industrial wastes. Compared to the conventional activated sludge, biogranules have a

regular, dense, and strong structure and good settling properties. They enable high

biomass retention and withstand high-strength wastewater and shock loadings.

Granulation in low turbulence system is reported acidifying bacteria, nitrifying

41

bacteria (De beer et al,. 1993) and denitrifying bacteria (Van der hoek, 1988).

Biogranulation can be classified as aerobic and anaerobic granulation.

3.2.1 Anaerobic Granulation

Formation of anaerobic granules has been extensively studied and is probably

best recognized in the upflow anaerobic sludge blanket (UASB) reactor (Alves et al.,

2000). Many wastewater treatment plants already apply anaerobic granulation

technology. The feasibility and efficiency of UASB reactors and their various

modifications (e.g., the internal circulation (IC) reactor) for removing biodegradable

organic matter from municipal and industrial wastewater have been reported (Lettinga

et al., 1980; Fang and Chui, 1993; Schmidt and Ahring, 1996). The anaerobic

granulation technology has some drawbacks. These include the need for a long start-

up period, a relatively high operation temperature and unsuitability for low-strength

organic wastewater. In addition, anaerobic granulation technology is not suitable for

the removal of nutrients (N and P) from wastewater. In order to overcome those

weaknesses, research has been devoted to the development of aerobic granulation

technology.

3.2.2 Aerobic Granulation

Aerobic granulation is first time reported by Mishima and Nakamura in 1991

in a continuous up-flow sludge blanket reactor. Aerobic granulation represents a

relatively new form of cell immobilization that has attracted recent research attention

(Tay et al. 2001; 2002). Aerobically grown microbial granules are self immobilized

aggregates of bacteria cultivated in sequencing batch reactors (SBRs) (Morgenroth et

al., 1997; Beun et al., 1999; Peng et al., 19999; Etterer and wilderer, 2001; Tay et al.,

2001a; Liu and Tay, 2002). Aerobic granules like anaerobic counterpart have a strong

compact microbial structure, good settling ability and high biomass retention, with the

ability to handle high organic loading rates (Zheng et al., 2006).

Aerobic granules are self-immobilized microbial consortia cultivated

aerobically mostly consist of mixed species of bacteria trapped in an extracellular

polymeric matrix secreted by themselves (Venogopalan et al., 2005). The growth of

aerobic granules is sometimes regarded as a special case of biofilm development (Liu

42

and Tay, 2002; Yang et al., 2004a). In fact, microbial granulation is quite fundamental

in biology and cell aggregation can be defined as the gathering together of cells to

form a fairly stable, contiguous, multicellular association under physiological

conditions (Calleja, 1984). Each aerobic granule is an enormous metropolis of

microbes containing millions of individual bacteria. Almost all aerobic granules have

been cultivated in sequencing batch reactors (SBRs). The SBR system is a modified

design of the conventional activated sludge process and has been widely used in

municipal and industrial wastewater treatment.

3.3 Sequential Batch Reactor

The sequencing batch reactor (SBR) is a fill-and draw activated sludge system

for wastewater treatment (Fig-3.2). In this system, wastewater is added to a single

“batch” reactor, treated to remove undesirable components, and then discharged.

Equalization, aeration, and clarification can all be achieved using a single batch

reactor (Metcalf and Eddy, 2003).

Figure 3.2 Sequencing batch reactor (SBR) design principle (USEPA, 2000)

The reactor volume of SBRs varies with time but remain constant with

traditional continuous flow systems (Venugopalan et al., 2005). Enforcement of

controlled short-term, non steady conditions in SBR may favor induction of enzymes

to degrade biorefractory compounds (Tomei et al., 2004). The unit processes of the

43

SBR and conventional activated sludge systems are the same. A 1983 U.S. EPA

report summarized this by stating;

“The SBR is no more than an activated sludge system

which operates in time rather than in space.”

The difference between the two technologies is that the SBR performs

equalization, biological treatment, and secondary clarification in a single tank using a

timed control sequence. In a conventional activated sludge system, these unit

processes would be accomplished by using separate tanks. SBR consist of column

shape vessel (Fig-3.3) as these systems have a relatively small footprint, they are

useful for areas where the available land is limited. In addition, cycles within the

system can be easily modified for nutrient removal in the future, if it becomes

necessary. This makes SBRs extremely flexible to adapt to regulatory changes for

effluent parameters such as nutrient removal. SBRs are also very cost effective if

treatment beyond biological treatment is required, such as filtration.

Figure 3.3 Schematic diagram of laboratory scale SBR

Temperature

Electrode

DO Electrode

PH Electrode

Air Diffuser

Effluent

Discharge Port

Effluent Inlet

Port

Solenoid valve

Timer

44

3.3.1 SBR Design and Operation

For any wastewater treatment plant design, the first step is to determine the

anticipated influent characteristics of the wastewater and end use of treated effluent.

Based on these parameters, and other site specification such as temperature, elevation

above sea levels key, total dissolve solids concentration, design parameters are

selected for the system. An example of these parameters for a wastewater system

loading is listed in Tab-3.1. Once the key design parameters are determined, the

number of cycles per day, number of basins, decants volume, reactor size, and

detention times can be calculated.

Municipal Industrial

Food to Mass 0.15-0.4/day 0.15-0.6/day

Treatment Cycle Duration 4.0 hours 4.0-24 hours

MLSS 2000-2500 mg L-1

2000-4000 mg L-1

Hydraulic Retention Time 6-14 hours varies

Source: AquaSBR Design Manual (1995)

Table 3.1 Key Design Parameters for a Conventional Load

Influent wastewater is added to the reactor during the Fill step. The following

three variations are used for the Fill step and any or all of them may be used

depending on the operating strategy: static fill, mixed fill, and aerated fill. During

static fill, influent wastewater is added to the biomass already present in the SBR. Yu

et al. (1996) in his study compared the performance of an SBR with unaerated fill and

an SBR with aerated fill for the treatment of synthetic phenolic wastewater, and to try

to select an appropriate fill mode. In this study, phenol was selected as an object

substrate. The wastewater tested was formed through adding phenol to domestic

sewage. At low influent phenol concentrations (e.g. 400 mg L-1

), the SBR with

unaerated fill performed better than the SBR with aerated fill, in which there was a

tendency for filamentous bacteria to develop. However, when the influent phenol

concentration was high (e.g. 800 mg L-1

), phenol accumulated during the fill period

in the SBR with unaerated fill became inhibitory to micro-organisms.

45

Mixed fill is classified by mixing influent organics with the biomass, which

initiates biological reactions. During mixed fill, bacteria biologically degrade the

organics and use residual oxygen or alternative electron acceptors, such as nitrate-

nitrogen. In this environment, denitrification may occur under these anoxic

conditions. An anoxic condition is defined as an environment in which oxygen is not

present and nitrate-nitrogen is used by the microorganisms as the electron acceptor. In

a conventional biological nutrient removal (BNR) activated sludge system, mixed fill

is comparable to the anoxic zone which is used for denitrification. Anaerobic

conditions can also be achieved during the mixed fill phase. After the microorganisms

use the nitrate-nitrogen, sulfate becomes the electron acceptor. Anaerobic conditions

are characterized by the lack of oxygen and sulfate as the electron acceptor.

Aerated Fill is classified by aerating the contents of the reactor to begin the

aerobic reactions completed in the React step. Aerated Fill can reduce the aeration

time required in the React step. The biological reactions are completed in the React

step, in which mixed react and aerated react modes are available. During aerated

react, the aerobic reactions initialized during aerated fill are completed and

nitrification can be achieved. If the mixed react mode is selected, anoxic conditions

can be attained to achieve denitrification. Anaerobic conditions can also be achieved

in the mixed react mode for phosphorus removal.

Settle is typically provided under quiescent conditions in the SBR. In some

cases, gentle mixing during the initial stages of settling may result in a clearer effluent

and a more concentrated settled sludge. In an SBR, there are no influent or effluent

currents to interfere with the settling process as in a conventional activated sludge

system. The Idle step occurs between the Draw and the Fill steps, during which

treated effluent is removed and influent wastewater is added. The length of the Idle

step varies depending on the influent flow rate and the operating strategy.

Equalization is achieved during this step if variable idle times are used. Mixing to

condition the biomass and sludge wasting can also be performed during the Idle step,

depending on the operating strategy.

SBR offer an attractive alternative to conventional biological wastewater

treatment process, mainly because of their simple, flexible operation and cost

effectiveness for industries discharging small amount of wastewater (Mangat and

46

elefsiniotis, 1999; chiavola et al., 2004). Some advantages and disadvantages of SBRs

are listed below:

3.3.2 Advantages

Equalization, primary clarification (in most cases), biological treatment, and

secondary clarification can be achieved in a single reactor vessel.

Operating flexibility and control.

Minimal footprint.

Potential capital cost savings by eliminating clarifiers and other equipment.

3.3.3 Disadvantages

A higher level of sophistication is required (compared to conventional systems),

especially for larger systems, of timing units and controls.

Higher level of maintenance (compared to conventional systems) associated with

more sophisticated controls, automated switches, and automated valves.

Potential of discharging floating or settled sludge during the draw or decant phase

with some SBR configurations.

Potential plugging of aeration devices during selected operating cycles, depending

on the aeration system used by the manufacturer.

3.4 Mechanisms of Aerobic Granulation

Successful cultivation of aerobic granules demands a series of conditions to be

fulfilled, although different mechanisms have been proposed (Beun et al., 2000; Liu

and Tay, 2002) but formation of aerobic granules was not yet understood because

bacteria indeed would prefer a dispersed rather than aggregated state. Hence, there

should be an initiating force that can bring bacteria together and, further, make them

aggregate.

Based on his research finding Beun et al. (1999) proposed a mechanism of

Aerobic granulation. According to his proposed mechanism growth of the filamentous

fungi is dominant after the reactor seeding, and these filamentous fungi under the

hydrodynamics shear condition can easily form compact quick settling pellets while

others washed out. They further proposed that these pellets because of possible

47

nutrient and oxygen limitation undergo lysis and produce bacterial colonies which

further grow in size to form granules (Fig-3.4).

Liu and Tay (2002) have also proposed a 4 step model for the cultivation of

aerobic granules. Aerobic granulation starts with Physical movement to initiate one to

one contact of bacteria. Hydrodynamics, diffusion mass transfer, gravity,

thermodynamic effects, and cell mobility are the factors which are suppose to be

responsible for physical movement. Physical movement is followed by stabilization of

the aggregates. Cell aggregates matured through production of extracellular polymer,

growth of cellular clusters, metabolic change, environment-induced genetic effects

that facilitate the cell–cell interaction and result in a highly organized microbial

structure. In final step shaping of the steady state three-dimensional structure of

microbial aggregate by hydrodynamic shear forces (Chisti, 1999a). According to

Hailei et al. (2006) formation of aerobic granules consists of following five stages:

microbes’ multiplication phase, floc appearance phase, floc cohesion phase, mature

floc phase and aerobic granule phase.

Figure 3.4 Proposed mechanism of aerobic granulation by Beun et al., (2000)

48

Recent research clearly demonstrated that this driving force behind aerobic

granulation could be the cell hydrophobicity (Liu et al., 2003). In fact, it is well

known that the physico-chemical properties of the cell surface have a profound effect

on the formation of biofilms and aerobic granules (Zita and Hermanson, 1997; Kos et

al., 2003). When bacteria becomes more hydrophobic, an increased cell-to-cell

adhesion is observed i.e. the cell surface hydrophobicity might contribute to the

ability for cells to aggregate (Kos et al., 2003). Bacteria are not likely to aggregate

naturally because of the repulsive electrostatic forces and hydration interactions

among them.

It has been observed that aerobic granules were successfully cultivated in the

SBRs operated at a settling time of < 15 min, while only bioflocs appeared in the

reactor run at the settling time of 20 min. The shorter settling time was seen to

significantly improve the production of cell polysaccharide. A feature of the SBR is

cyclic operation and the settling time acts as hydraulic selection pressure on the

microorganisms. Selection pressure can be used to induce microbial changes that

favor the formation of aerobic granules. Although mechanisms and models for aerobic

granulation have been described, they do not provide a complete picture of the

granulation process. Intercellular communication and multicell coordination are

known to contribute to the organization of bacteria into spatial structures. Quorum

sensing has been shown to be one example of social behavior in bacteria, as signal

exchange among individual cells allows the entire population to choose an optimal

way of interacting with the environment. The cellular automaton model shows that

biofilm structure is determined by localized substrate concentration. A cell can

determine its position in a concentration gradient of an extracellular signal factor and

uses this to modify its development. Research on cell–cell communication confirms

that cell–cell signaling is effective in developing aerobic granules and organizing the

spatial distribution of the bacteria in the granules. Quorum sensing effects in aerobic

granules need to be further examined. Liu et al. (2004) demonstrate that aerobic

granulation process is driven by number of parameters like substrate composition,

organic loading rate, hydrodynamic shear force, settling time, hydraulic retention

time, reactor configuration dissolved oxygen etc. These selection pressure works as

triggers for aerobic granulation.

49

3.5 Factors Affecting Aerobic Granulation

For the successful cultivation of aerobic granules, a series of conditions have

to be met (Tay et al., 2003a; Zhu and Wilderer, 2003) same as in the case of cells in a

culture where, a series of factors are responsible for its successful aggregation (Liu

and Tay 2004)

3.5.1 Substrate Composition

Cultivation of Aerobic granules have been reported with a wide variety of

substrates including glucose (Fig-3.5a), acetate (Fig-3.5b), ethanol, phenol, and

synthetic wastewater (Beun et al., 1999; Peng et al., 1999; Tay et al., 2001a; Moy et

al., 2002; Jiang et al., 2002; Schwarzenbeck et al., 2003). Aerobic granules have been

also cultivated with nitrifying bacteria and an inorganic carbon source (Tay et al.,

2002b; Tsuneda et al., 2003). However, granule microstructure and species diversity

appear to be related to the type of carbon source (Beun et al., 2002, Zheng et al.,

2005). Fig-3.6a shows that glucose-fed aerobic granules have exhibited a filamentous

structure (Chudoba, 1985), while acetate-fed aerobic granules have had a non-

filamentous and very compact bacterial structure in which a rod like species

predominated (Fig-3.6b). In recent years use of aerobic granular sludge for domestic

sewage treatment has been studied (De Kreuk and van Loosdrecht, 2006). These

nitrifying granules showed excellent nitrification ability. More recently, aerobic

granules were also successfully developed in laboratory-scale SBR for treating toxic

substrate such as chlorophenols and nitrophenols (Jiang et al., 2004; Tomei et al.,

2005; Yi et al., 2006).

Figure 3.5 Aerobic Granules, glucose-fed (a) and acetate-fed (b)

(Tay et al., 2001a)

50

Figure 3.6 Microstructures of aerobic granules, glucose-fed (a) and acetate-fed (b)

(Tay et al., 2001a)

3.5.2 Organic Loading Rate

The importance of organic loading rate in case of anaerobic granules has been

studied. In comparison to aerobic granules, high organic loading rates facilitate the

formation of anaerobic granules. Cultivation of aerobic granules is reported over a

very wide range of organic loading rates from 2.5 to 15 kg COD m-3

day-1

(Moy et al.,

2002; Liu et al., 2003a) (Fig-3.7). It appears that aerobic granulation is not affected by

the organic loading rate but Tay et al. (2004) reported that it was difficult to form

granules with organic loading rates lower than 2 kg COD m−3

day−1

. The physical

characteristics of aerobic granules depend on the organic loading rate. The mean size

of aerobic granules increased from with the increase of the organic loading (Liu et al.,

2003a) same as in case of anaerobic granulation (Grotenhuis et al., 1991). The effect

of organic loading rate on the morphology of aerobic granules in terms of roundness

was found to be insignificant, while the aerobic granules developed at different

organic loading rates exhibited comparable dry biomass density, specific gravity, and

sludge volume index (SVI), the physical strength of aerobic granules decreased with

the increase of organic loading rate (Liu et al., 2003a; Tay et al., 2001a). (Liu et al.,

2003c) explained that an increased in organic loading rate can enhance the biomass

growth rate but reduces the strength of the three-dimensional structure of the

microbial community. Wang et al. (2007) verified that high organic loading rate

would be favorable for granule stability.

51

(A) (B) (C)

Figure 3.7 Morphology of granules (Moy et al., 2002)

(A) 6 kg COD m-3

day-1 (

B) 15 kg COD m-3

day-1

(C) 9kg COD m-3

day-1

3.5.3 Hydrodynamic Shear Force

Hydrodynamic shear force plays a pivotal role in successful cultivation of

aerobic granules. Literature review suggested that high shears force have a positive

impact on microorganism aggregation and granule stability (Khan et al., 2009; chisti,

1999a; Shin et al., 1992; Tay et al., 2001a). Higher shear forces in the SBR resulted

in more dense granules with a smaller diameter (De Kreuk and van Loosdrecht, 2006)

It is reported by Tay et al. (2001a) that cultivation of aerobic granules could be

feasible at a superficial upflow air velocity of 1.2 cm sec-1

or more. High

hydrodynamic shear force favors more regular, rounder, and compact aerobic granules

(Tay et al, 2001a). High shear force also favors granular density and strength (Tay et

al., 2003c). In nutshell shear force in one of the deciding factor for the aerobic granule

cultivation. However, how the high shear force help in aerobic cultivation is still

unclear.

The relation between the production of extracellular polysaccharides and the

shear force is a proven fact. EPS is believed as a structural backbone of the aerobic

granules. It is reported that extracellular polysaccharides can mediate both cohesion

and adhesion of cells and the stability of aerobic granules is very much dependent on

the production of EPS (Tay et al., 2001a; Tay et al., 2001c). Secretion of EPS by the

bacteria is triggered by a high shear force. In case of biofilm, shear force-induced

production of extracellular polysaccharides has been reported earlier (Ohashi and

Harada, 1994). Hence, the EPS production is closely related to the compact and

stronger structure of aerobic granules (Beun et al., 1999; Chsti, 1999a).

52

1 liters min-1

No granules

3 liters min-1

Stable granules but small in size (Hailei et al., 2006)

2 liters min-1

Large but overgrown filamentous leading to SBR failure

3.5.4 Settling Period

SBR operates in a cycle consist of different steps. At the end of every cycle,

after aeration the biomass is settled before the effluent is withdrawn. The settling time

is a major selection pressure regarding high settling sludge. Etterer and Wilderer

(2001) reported that during the start-up period a short settling time washed out the

biomass with low settling ability. A short settling time favors the growth of bacteria

with high settling ability. Settling velocity or settling time affected the granule

characteristics and ratio between granule and flocs in reactor. The production of EPS

is triggered at short settling times (Adav et al., 2009a). But at the same time a short

settling time could result into washout of the biomass and caused the system failure.

Hence, selection of an optimal settling time is very important in SBR operation. In

most of the reported cases aerobic granules settle within 1 min, resulting into a clear

effluent discharge (Tay et al., 2001a). The immobilized biomass in the reactor ensures

a faster and more efficient removal of organic pollutants in wastewater. Cultivation of

aerobic granules has been reported only in the SBR operated at a settling time of 5-10

min (Qin et al., 2004; Tay et al., 2001a). In SBR always a short settling time has been

kept to promote granulation process (Tay et al., 2001b; Beun et al., 2002; McSwain et

al., 2004; Thanh, 2005; Chen et al., 2008).

3.5.5 Hydraulic Retention Time

It is defined as the volume of the effluent discharge divided by the working

volume of the reactor. HRT plays a significant role in selection of biomass with good

settling ability. A short HRT select the high settling velocity biomass and light

dispersed sludge is washed out. But a very short HRT could result into the excess

sludge loss which leads to the failure of the granulation process. Hence an optimum

HRT should be selected for the healthy cultivation of the aerobic granules (Fang and

Yu, 2000 and 2001). Pan et al. (2004) have cultivated granules on different HRT they

successfully demonstrated the relation between granular size and HRT (Fig-3.8).

53

Figure 3.8 Microscope images of granules cultivated on different HRT

a- 2, b-6, c-12 and d-24 hrs (Pan et al., 2004)

3.5.6 Aerobic Starvation

The SBR cycle consist of feeding, aeration, settling, and discharging of treated

effluent. Because of different steps the microorganisms present in reactor are subject

to a periodic starvation during the course of SBR operation. This periodic starvation

affects the cell hydrophobicity of the microorganism, which triggered the aerobic

granulation (Tay et al., 2001a; Li et al., 2006b). It has been found that cell surface

hydrophobicity was proportionally related to the starvation time in SBR (Bossier and

Verstraete, 1996; Tay et al., 2001a). Recently McSwain et al. (2003) have developed

operation strategy to enhance aerobic granulation is by intermittent feeding. The

aeration period of the SBR operation consists of two phases. In first phase the

substrate is degraded to a minimum and in second phase substrate are no longer

available leads to starvation (Liu and Tay, 2008). Tay et al. (2001) reported that the

long periodical starvation phase played a role for microbial granulation in the SBR.

Such changes contribute to microbial ability to aggregate. It has been reported that

under starvation bacteria become more hydrophobic which leads to compactness of

granules but a large starvation period weakened the granules stability (Wang et al.,

2005). Very short starvation time facilitates the pre-anoxic feast period positively

54

affected granulation process through different mechanisms (Wan et al., 2009; Pijuan

et al. 2009). However, according to Liu and Tay (2008) starvation period is not the

key factor in aerobic granulation. Further, they also reported that a growth of fluffy

granules with poor settling. The formation of aerobic granule is initiated by starvation

and cooperated by shear force. It makes bacteria more hydrophobic which promote

the granulation from flocs (Tay et al., 2001a; Li et al., 2006b). The aggregation is

regarded as a strategy of cells against starvation. It was reported that bacteria become

more hydrophobic which facilitates adhesion or aggregation under starvation

conditions (Bossier and Verstraete., 1996; Tay et al., 2001a).

3.5.7 Presence of Calcium Ion in Feed

It has been reported that addition of Ca2+

and Mg2+

accelerated the aerobic

granulation process (Liu et al. 2010). With addition of Ca2+

(100 mg L-1

) the

formation of aerobic granules took 16 days compared to 32 days in the culture without

Ca2+

added. The Ca2+

augmented aerobic granules also showed better settling and

strength characteristics and had higher polysaccharides contents (Jiang et al., 2003). It

has been proposed that Ca2+

binds to negatively charged groups present on bacterial

surfaces and extracellular polysaccharides molecules and thus acts as a bridge to

promote bacterial aggregation. Polysaccharides play an important role in maintaining

the structural integrity of biofilms and microbial aggregates, such as aerobic granules,

as they are known to form a strong and sticky non-deformable polymeric gel-like

matrix.

3.5.8 Dissolved oxygen, pH and Temperature

pH of the medium is also a decisive factor during aerobic granulation and

degradation. A slight alkaline pH (around 7.5) is necessary for proper aerobic

granulation and granulation cease to occur at pH>8.5 as reported by (Hailei et al.,

2006). Furthermore, a slight low pH favours fungi dominating granules whereas a

slight alkaline pH favours bacterial dominating granules (Yang et al., 2008). Most

fungi prefer a low-pH medium, which is nevertheless unfavourable to the growth of

bacteria. A high alkalinity in HCO3¯ can prevent a drop in pH and inhibit fungal

growth. These factors apparently imply that, by controlling the alkalinity in the

55

influent, a strategy of species selection can be developed for the granulation of

aerobic sludge with different microbial communities, morphological evolutions and

structural properties (Williams and Reyes, 2006).

Temperature can also influence the performance of aerobic granulation to

some extent. Hailei et al. (2006) reported that too high (41°C) or too low temperature

(26 °C) can lead to decrease in biomass which is necessary for granule formation.

However, influence of temperature on biogranulation is not very significant in the

range 29-38 °C. Song et al. (2009) cultivated aerobic granules in sequencing batch

airlift reactors (SBAR) at 25, 30, and 35 °C, respectively and analysed the microbial

community structures of the granules using scanning electron microscope (SEM) and

polymerase chain reaction-denaturing gradient gel electrophoresis (PCR-DGGE).and

reported that 30 °C is optimum temperature for matured granule cultivation, where the

granules had a more compact structure, better settling ability and higher bioactivity.

Hence for optimum purposes, the temperature of the SBR system should be kept in

between 28-35 °C for proper functioning of microorganism.

Dissolve oxygen concentration is one of the factors influencing the

biodegradation of a substance under aerobic conditions. 1 mgL-1

is normally the

minimum requirement for aerobic biodegradation. Aerobic granules have been

reported at DO concentration as low as 0.7 to 1.0 mg L-1

and more than 2 mg L-1

. So

far it appears that DO concentration plays an insignificant role in the formation of

aerobic granules.

3.5.9 Seed Sludge

Aerobic granules have been cultivated on seed sludge taken from conventional

activated sludge processes (Beun et al., 1999; Tay et al., 2001; Jang et al., 2003;

Arrojo et al., 2004; Wang et al., 2004; Qin et al., 2004; Schwarzenbeck et al., 2004;

Thanh, 2005). In case of anaerobic granulation, characteristics of the seed sludge

influence the formation and properties of anaerobic granules which is not reported in

case of aerobic granules. A series of granules are reported on different seed sludge

Therefore, types of seed sludge do not play a role in cultivating aerobic granules.

56

3.5.10 Reactor Configuration

Reactor configuration has an impact on the flow pattern of liquid and

microbial aggregates in the reactor (Beun et al., 1999; Liu and Tay, 2002). Batch

operating reactors are preferable over continuous systems for granulation process due

to the existence of periodic feast-famine stages and high gradient of substrate

concentration in case of batch reactor (Beun et al., 2002). Although aerobic granules

are also reported in continuous reactor such as Biofilm Airlift Suspension Reactor

(Tijhuis et al., 1994). Mostly aerobic granules were produced in column-type upflow

reactors. The flow in column shape reactors can create a relatively homogenous

circular flow moving upword along the reactor’s axis and microbial aggregates are

constantly subject to a hydraulic attrition. The circular flow apparently forces the

microbial aggregates to adapt a regular granular shape that has a minimum surface

free energy.

A high ratio of reactor height to diameter (H/D) can ensure a longer circular

flow trajectory which in turn provides a more effective hydraulic attrition to microbial

aggregates in a column-type upflow reactor. For practical applications, the SBR

should have a high H/D ratio to improve selection of granules by the difference in

settling velocity (Beun et al., 1999). A high H/D ratio and the absence of an external

settler result in a reactor with a small footprint. Although the effect of H/D on aerobic

granulation is neglected by few researchers (Kong et al., 2010).

3.5.11 Inhibition to Aerobic Granulation

Yang et al. (2004a and b) investigated the inhibitory effect of free ammonia on

aerobic granulation in a SBR fed with acetate as the sole carbon source. It was found

that the Aerobic granules formed only when the free ammonia concentration was less

than 23.5 mg L-1

and nitrification was completely inhibited at a free ammonia

concentration of >10 mg L-1

. High free ammonia concentration suppresses

hydrophobicity and polysaccharide production which results into the failure of aerobic

granulation process. The cell hydrophobicity decreased from 70.6% to 40.6% with the

increase of the free ammonia concentration from 2.5 mg/L to 39.6 mg/L. The PS/PN

ratio decreased from 2.8 to 0.55 when free ammonia concentration increases from 2.5

to 39.6 mg/L and no granules were observed in the reactors (Yang et al., 2004).It has

57

been demonstrated that free ammonia could hinder the formation of aerobic granules

by inhibiting the energy metabolism of microorganisms Yang et al. (2004a, b).

3.6 Applications of Aerobic Granulation Technology

Aerobic granulation technology is used for the variety of wastewater ranging

from nutrient to recalcitrant organic compounds (Jiang et al., 2002; Moy et al., 2002;

Tay et al., 2002e; Lin et al., 2003). The performance of any biological wastewater

treatment depends on the reactor configuration, the active biomass concentration, the

biodegradation rates and the feeding rates of the pollutants. Process efficiency of

large-scale treatment plants can be improved by using aerobic granular sludge in ways

that allow high conversion rates and efficient biomass separation to minimize the

reactor volume. Because of its unique property it can be used many places with a high

efficiency, low initial capital investment and less running and maintenance cost. The

aerobic granules cultivated in SBR are very easy to incorporate with any current

wastewater treatment facility set-up

3.6.1 Recalcitrant Organic Wastewater Treatment

The use of Aerobic granule is not only for the substances such as glucose,

acetate (Tay et al., 2001a), peptone, etc. but also recalcitrant such as phenol (khan et

al., 2009; Usmani et al., 2007; Jiang et al., 2004; Tay et al., 2004 & 2005),

Chlorophenol (khan et al., 2011; usmani et al., 2010) pyridine (Adav et al., 2007c), p-

nitrophenol (khan et al., 2010; Yi et al., 2006), 2,4- dichlorophenol (Wang et al.,

2007b), methyl t-butyl ether (Zhang et al., 2008) and nitrilotriacetic acid (Nancharaiah

et al., 2006).

Degradation of phenol is reported by Tay et al. (2004). He found that the

granules degraded phenol at a specific degradation rate of 1 g g-1

VSS day-1

at 500 mg

L-1

of phenol or at a reduced rate of 0.53 g g-1

VSS day-1

at 1900 mg L-1

of phenol. At

the same time it is reported that aerobic granules could remove phenol at 1.18 g g-

1VSS day

-1 (Adav et al. 2007). Yi et al. (2006) reported that the specific degradation

rate of p-nitrophenol (PNP) increased with corresponding increase in PNP

concentration up to 40.1 mg L-1

and declined with any further increase in PNP

concentration. Aerobic granules can efficiently degrade pyridine over initial

58

concentrations of 200– 2500 mg L-1

(Adav et al., 2007c). The specific degradation

rate of pyridine has been found to be 73.0 and 66.8 mg pyridine g-1

VSS day-1

at 250

and 500 mg L-1

of pyridine, respectively.

3.6.2 Organic and Nitrogen Removal

Scientist investigated the simultaneous removal of organics and nitrogen by

aerobic granules. Heterotrophic, nitrifying, and denitrifying populations were shown

to successfully coexist in microbial granules. Simultaneous nitrification and

denitrification process produce nitrogen gas in the system (Beun et al., 2002; Arrojo

et al., 2004; McSwain et al., 2004; Cassidy and Belia, 2005; Schwarzenbeck et al.,

2005; Mosquera-Corral et al., 2005; Qin and Liu, 2006). There are some other options

in order to enhance complete nitrogen removal; one of them is alternative

anoxic/aerobic cycle (Arrojo et al., 2004; Cassidy and Belia, 2005). The use of

aerobic/anoxic cycle can also enhance nitrogen removal (Yang et al., 2003; Qin and

Liu, 2006). Complete nitrogen removal is achieved by apply low DO concentration of

0.5 mg L-1

(Yang et al., 2003) or by adding external organic substrate (Qin and Liu,

2006). Dissolved oxygen (DO) concentration had a pronounced effect on the

efficiency of denitrification by microbial granules and a certain level of mixing was

necessary for ensuring sufficiency of mass transfer between the liquid and granules

during denitrification. Similar phenomena were reported by other researchers (Beun et

al., 2001). In nutshell, it appears that nitrogen removal can happen effectively in

granule even aerobic condition exists in system and depends on granule structures,

size, oxygen concentration, external carbon source and microbial population.

3.6.3 Phosphorous Removal

The discharge limit of phosphorus in wastewater is as low as 0.5 to 2.0 mg L-1

.

The enhanced biological phosphorus removal (EBPR) based on suspended biomass

process removes P without the use of any chemical but at the same time most EBPR

processes require large reactor volumes.

The use of aerobic granules to overcome the problems associated with the

conventional phosphorus removal is frequently reported. The Phosphorus removal

efficiency is high (more than 70%) in aerobic granular sludge system. The scope of

59

improvement in efficiency is still there when an enhanced P removal process will

applied. The anaerobic condition can be created by static fill or low mixing. The P is

released by the granules during the anaerobic fill period, and then rapidly taken up

during the aerobic react. The P content in aerobic granule is in the range of 1.9-9.3%,

depending on the ratio of P/COD of the influent (Mosquera-Corral et al., 2005; de

Kreuk et al., 2005).

3.6.4 Biosorption of Heavy Metals and Dyes

Many biomaterials have been tested as biosorbents for heavy metal removal.

These include marine algae, fungi, waste activated sludge, and digested sludge.

Because of its characteristics of aerobic granules are ideal biosorbent for heavy metals

removal. The granules have high settling velocity and have large surface area and

high porosity for adsorption. Moreover, because of its good settling ability the

granules can be easily separated from the effluent. Because of these properties

different dyes such as Rhodamine B and Malachite were found to adsorb by aerobic

granules effectively. Especially, the excellent settleability of aerobic granules could

ensure a rapid solid liquid separation of the treated water, which helps in a simple

process design (Liu et al., 2003c). The number of ways have been reported where

immobilized cell bioaccumulate metals (Merrounm 2003; Gorby 1992; renninger, N

2004; Nancharaiah et al., 2006) .

1. Biosorption to cell components or extracellular polymeric matrix

2. Bioaccumulation

3. Precipitation by reation with inorganic ligands

4. Microbial reduction of soluble metals to insoluble metals

The use of aerobic granules for Zn2+

and Cd2+

biosorption has been reported in

details (Liu et al., 2002, 2003c and d). it was shown that The biosorption of Zn2+

is

related to the initial Zn2+

concentration and granular bimass. In fact metal

concentration gradient is the main driving force for their Biosorption .Biosorption was

observed to be rapid in acidic range of 1-6 compared to a pH of 7.0 or above

(Venogopalan et al., 2006). They further observed that cation such as Na+, K

+, Mg

+

60

and Ca2+

were simultaneously released into the bulk solution during U(VI)

biosorbtion, showing the involvement of ion exchange mechanism in radionuclide

uptake. They have further established that in U-removal involve passive sorption as

live and dead biomass did not show any significant difference in removal.

3.7 Recent Trends in Aerobic Granulation in SBR

Studies Finding Reference

Effect of different

operating conditions on

aerobic granulation was

studied

Both a short HRT and a relative high shear were found

favorable for granulation. A substrate loading rate of

7.5 kg COD m-3

d-1

was applied. This led to formation

of granules with an average diameter of 3.3 mm and a

biomass density of 11.9 g-1

VSS L-1

granule.

Beun et

al., 1999

Aerobic granules were

cultivated using synthetic

urban WW containing

NaAc at low DO

Microscopic examination showed that the morphology

of the granules was nearly spherical (0.3-0.5 mm

diameter) with high COD removal efficiency and good

settleability.

Peng et

al., 1999

Effect of various

parameters (temperature,

shear force and DO) on

aerobic granulation

For better granulation, temperature should be in

between 29-38 °C, shear force should be greater than

2.0 cm sec-1

, DO should be greater than or equal to 2.5

mg L-1

.

Hailei et

al., 2006

Effect of Mg+2

on aerobic

sludge granulation in SBR

Mg+2

augmented SBR showed significant decrease in

the sludge granulation time 32 to 18 days. The results

demonstrated that Mg+2

enhanced the sludge

granulation process in SBR.

Li et al.,

2008b

Effect of long term

anaerobic and intermittent

anaerobic/

aerobic starvation on aero

bic granules

The loss of granular compactness was faster and more

pronounced under anaerobic/ aerobic starvation

conditions. These results suggested both anaerobic and

intermittent anaerobic/aerobic conditions are suitable

for maintaining granule structure and activity

during starvation.

Pijuan et

al., 2009

Cultivation of aerobic

granules by seeding

MMPs to bioreactor

Strain HSD can form MMPs under high

Mn2+

concentration. BAGs can be cultivated using

MMPs. MMPs result in the formation of aerobic

granules containing MMPs as nuclei and also induce

the formation of self-immobilized biogranules which

do not have the MMP at their core.

Hailei et

al., 2010

61

Effect of seed sludge Beer

waste-water treatment

plant (BWTP) and

Municipal wastewater

treatm-ent plant (MWTP)

on character-istics and

microbial community of

aerobic granular sludge

BWST was more suitable for cultivating aerobic

granules than that of sludge from MWTP. The results

suggested that dominant species in mature granules

cultivated by BWTP were Paracoccus sp., Devosia

hwasunensi, Pseudoxanthomonas sp., while the

dominant species were Lactococcus raffinolactis

and Pseudomonas sp. in granules developed from

MWTP.

Song et

al., 2010

Behavior of polymeric

substrates in an

aerobic granular sludge

system

Aerobic granules maintained on starch as sole influent

carbon source were filamentous and irregular. Low

starch hydrolysis rates, leading to available substrate

during the aeration period (extended feast period) and

resulting in increased substrate gradients over the

granules which induces a less uniform granule

development.

De Kreuk

et al.,

2010

Effect of filamentous

growth on aerobic

granular system

Filamentous granules exhibited low porosity and fast

settling velocity, and were more compact even than

bacteria granules. Filame-ntous microbes could form

compact granular structure, which may encourage the

utilization of filamentous microorganisms rather than

the inhibition of their growth, as the latter is frequently

used for sludge bulking control.

Li et al.,

2010a

Enhanced biological

phosphorus removal by

granular sludge

Results showed that positive charged particles were

formed with the release of phosphorus in the anaerobic

stage. These particles served as the cores of granules

and stimulate the granulation. Further, smaller granules

had a higher specific area, pore width and

phosphorus removal activity than bigger granules.

Wu et al.,

2010

The role of nitrate and

nitrite in a granular sludge

process

It was reported that aerobic granules allowed anoxic

phosphorus uptake using either nitrite or nitrate as an

electron acceptor. Nitrate presence was the best option

for phosphorus-rich effluents (3.04 mg P mg−1

N), while

nitrite was more useful with the simultaneous

nitrification-denitrification and phosphorus removal of

nitrogen-rich influents (1.68 mg P mg−1

N).

Coma et

al., 2010

Aerobic granules were

cultivated at a COD of 2.5

kg Acetate-COD m-3

in

SBAR and compared with

granules in BASR

The most importance difference was that the density of

the granules in the SBAR was much higher than the

density of the biofilms in the BASR

Beun et

al., 2002

62

Effect of H/D ratio on

aerobic granulation in a

SBR

These results recommended strongly that reactor H/D

ratio or reactor setting velocity do not affect granule

formation, and stable operation of granular sludge

reactor. Hence, reactor H/D ratio can thus be very

flexible in the practice, which is important for the

application of aerobic granular technology.

Pan et al.,

2004

Biodegradation of phenol

was studied in batch

reactor

The degradation rate was maximum 15.7 mg L-1

hrs-1

at

400 mg L-1

phenol and specific growth rate followed

Haldane and Han–Levenspiel models.

Saravanan

et al.,

2008

Granulation of activated

sludge in a pilot-scale

sequencing batch reactor

for the treatment of low-

strength MWW

The results shows that the volume exchange ratio and

settling time of an SBR were found to be two factors in

the granulation of activated sludge grown on the low-

strength municipal wastewater.

Ni et al.,

2009

Role of selective sludge

discharge as the

determining factor in

SBR aerobic granulation

The results suggested that aerobic granulation may not

require the dominance of any particular species. Small

and loose sludge flocs were found to have an advantage

over larger and dense granules in substrate uptake.

Li and Li,

2009

Effect of different settling

times on aerobic

granulation in SBR

The results demonstrated that short settling times at

initial stage principally determine the efficiency of

subsequent granulation processes. Early granulation of

aerobic granules is observed at a settling time of 5 min.

Adav et

al., 2009a

Role of denitrification on

aerobic granular sludge

formation in sequencing

batch reactor

The results proved that the presence of nitrate

(commonly provided by nitrification) in SBRs can

assist the densification of the biological aggregates in

aerobic condition and can be considered as a possible

factor which helps to maintain the granulation process.

Wan and

Sperandio,

2009

Enhanced storage stability

of aerobic granules seeded

with activated sludge flocs

and pellets

Compared with granule based on pellets, flocs based

granules had more decrease in biomass concentration,

settleability, hydrophobicity, and EPS concentration

after the storage. Aerobic granules seeded with pellets

were more resistant against storage, and thus would

have greater potential in practical applications.

Xu et al.,

2010

Aerobic granules

cultivated on low and high

ammonium and phosphate

salts

Aerobic granules cultivated with low ammonium and

phosphates lost structural stability within 3 days in

CFRs while stable aerobic granules were cultivated in

substrate with high levels of ammonium salts that could

stably exist for 216 days CFRs with or without

submerged membrane

Juang et

al., 2010

63

Effects of seed sludge

properties and selective

biomass discharge

on aerobic sludge

granulation

The results showed that aerobic granules could be

formed in the reactors from both seed sludge (small

loose flocs and larger denser flocs) of different

structural and settling properties. Selective sludge

discharge facilitated the growth and accumulation of

denser sludge in the reactor, leading to

complete granulation.

Sheng et

al., 2010

Cultivating autotrophic

nitrifying granules in SBR

Both ammonia and nitrite-oxidizing bacteria were

observed in the granular sludge. An NH4+-N

concentration range of 100-250 mg L-1

was found to be

favourable.

Shi et al.,

2010

Aerobic granules with

inhibitory strains and role

of extracellular polymeric

substances

The compact nature of aerobic granules was generally

assumed to provide spatial isolation, resulting in the

co-occurrence of diverse strains that have similar or

dissimilar functions.

Adav et

al., 2010

Comparing the effect of

Ca2+

and Mg2+

enhancing aerobic

granulation in SBR

Reactor fed with Ca2+

(R1) showed faster granulation

process and better physical characteristics compared

with reactor having Mg2+

(R2). However, the mature

granules in R2 had a higher production yield of

polysaccharides and proteins with a faster substrate

biodegradation rate.

Liu et al.,

2010

Evaluation of pre-anoxic

feast period on granular

sludge formation in a

SBAR

The presence of pre-anoxic phase clearly improved the

densification of aggregates and allowed granular sludge

formation at reduced air flow rate (0.63 cm sec-1

). A

low sludge volume index of 45 mL g-1

and a high

MLSS concentration (9–10 g L-1

) were obtained in the

anoxic/aerobic system compared to more conventional

results for the aerobic reactor.

Wan et al.,

2009

Effect of feeding time on

aerobic granular system

A short feeding time causes a rapid exposure of

microorganism to the high concentrations of toxic

substrates leading to their inhibition and it is

unfavorable for aerobic granulation.

Tomei et

al., 2005

Effect of the specific

surface area and operating

mode on biological phenol

removal using packed bed

reactors

The specific surface area strongly affects the observed

phenol biodegradation rate. Draw-fill operation with

recirculation of the pilot-scale packed bed reactor

proved to be a very effective operating mode, since

homogeneity is achieved in the reactor, resulting in a

better exploitation of the filter volume

Tziotzios

et al.,

2007