state of the art of biogranulation technology for wastewater treatment

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Research review paper State of the art of biogranulation technology for wastewater treatment Yu Liu * , Joo-Hwa Tay Division of Environmental and Water Resources Engineering, School of Civil and Environmental Engineering, Nanyang Technological University, 50 Nanyang Avenue, Singapore 639798, Singapore Accepted 6 May 2004 Available online Abstract Biogranulation technology developed for wastewater treatment includes anaerobic and aerobic granulation processes. Anaerobic granulation is relatively well known, but research on aerobic granulation commenced only recently. Many full-scale anaerobic granular sludge units have been operated worldwide, but no report exists of similar units for aerobic granulation. This paper reviews the fundamentals and applications of biogranulation technology in wastewater treatment. Aspects discussed include the models of biogranulation, major factors influencing biogranulation, characteristics of biogranules, and their industrial applications. This review hopes to provide a platform for developing novel granules-based bioreactors and devising a unified interpretation of the formation of anaerobic and aerobic granules under various operation conditions. D 2004 Elsevier Inc. All rights reserved. Keywords: Biogranulation; Aerobic granules; Anaerobic granules; Operating parameters; Microbial structure; Diversity; Mechanism of granulation 1. Introduction Biogranulation involves cell-to-cell interactions that include biological, physical and chemical phenomena. Biogranulation can be classified as aerobic and anaerobic granulation. Biogranules form through self-immobilization of microorganisms. 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 0734-9750/$ - see front matter D 2004 Elsevier Inc. All rights reserved. doi:10.1016/j.biotechadv.2004.05.001 * Corresponding author. E-mail address: [email protected] (Y. Liu). www.elsevier.com/locate/biotechadv Biotechnology Advances 22 (2004) 533 – 563

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Page 1: State of the art of biogranulation technology for wastewater treatment

www.elsevier.com/locate/biotechadv

Biotechnology Advances 22 (2004) 533–563

Research review paper

State of the art of biogranulation technology for

wastewater treatment

Yu Liu *, Joo-Hwa Tay

Division of Environmental and Water Resources Engineering, School of Civil and Environmental Engineering,

Nanyang Technological University, 50 Nanyang Avenue, Singapore 639798, Singapore

Accepted 6 May 2004

Available online

Abstract

Biogranulation technology developed for wastewater treatment includes anaerobic and aerobic

granulation processes. Anaerobic granulation is relatively well known, but research on aerobic

granulation commenced only recently. Many full-scale anaerobic granular sludge units have been

operated worldwide, but no report exists of similar units for aerobic granulation. This paper reviews

the fundamentals and applications of biogranulation technology in wastewater treatment. Aspects

discussed include the models of biogranulation, major factors influencing biogranulation,

characteristics of biogranules, and their industrial applications. This review hopes to provide a

platform for developing novel granules-based bioreactors and devising a unified interpretation of the

formation of anaerobic and aerobic granules under various operation conditions.

D 2004 Elsevier Inc. All rights reserved.

Keywords: Biogranulation; Aerobic granules; Anaerobic granules; Operating parameters; Microbial structure;

Diversity; Mechanism of granulation

1. Introduction

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

chemical phenomena. Biogranulation can be classified as aerobic and anaerobic granulation.

Biogranules form through self-immobilization of microorganisms. 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

0734-9750/$ - see front matter D 2004 Elsevier Inc. All rights reserved.

doi:10.1016/j.biotechadv.2004.05.001

* Corresponding author.

E-mail address: [email protected] (Y. Liu).

Page 2: State of the art of biogranulation technology for wastewater treatment

Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563534

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

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

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

Formation of anaerobic granules has been extensively studied and is probably best

recognized in the upflow anaerobic sludge blanket (UASB) reactor.Manywastewater treatment

plants already apply anaerobic granulation technology (Alves et al., 2000). 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 successfully demonstrated (Lettinga et al., 1980; Fang and Chui, 1993; Schmidt and

Ahring, 1996). Anaerobic granular sludge is a dense microbial community that typically

includes millions of organisms per gram of biomass. None of the individual species in these

microecosystems is capable of completely degrading the influent 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 bemaintained 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. Anaerobic granular sludge has proved capable of treating

high-strength wastewater contaminated with soluble organic pollutants.

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. The

development of aerobic granules was first reported by Mishima and Nakamura (1991) in a

continuous aerobic upflow sludge blanket reactor. Aerobic granules with diameters of 2 to

8 mm were developed, with good settling properties. Aerobic granulation has since been

reported in sequencing batch reactors (SBRs) by many researchers (Morgenroth et al., 1997;

Beun et al., 1999; Peng et al., 1999; Etterer and Wilderer, 2001; Tay et al., 2001a; Liu and

Tay, 2002) and has been used in treating high-strength wastewaters containing organics,

nitrogen and phosphorus, and toxic substances (Jiang et al., 2002; Moy et al., 2002; Tay et

al., 2002e; Lin et al., 2003; Yang et al., in press). Development of biogranules requires

aggregation of microorganisms. For bacteria in a culture to aggregate, a number of

conditions have to be met. The formation of anaerobic granules is a multiple-step process

that involves physicochemical and biological forces. This review is focused on progresses in

biogranulation technology developed for wastewater treatment.

2. Aerobic granulation technology

2.1. The formation of aerobic granules

Sludge is the microbial biomass that utilizes nutrient substrates present in wastewater.

Microbial granules can be regarded as compact and dense microbial aggregates with a

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 535

spherical outer shape. The growth of aerobic granules is sometimes regarded as a special

case of biofilm development (Liu 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. Aerobic granulation may be initiated by

microbial self-adhesion. Bacteria are not likely to aggregate naturally because of the

repulsive electrostatic forces and hydration interactions among them.

Tay et al. (2001a) used different microscopic techniques to investigate how an aerobic

granule formed from seed sludge. For comparison, granules were cultivated in two reactors

fed with glucose in one case and acetate in the other case, as sole carbon sources. The

results showed that the seed sludge had a very loose and irregular structure, dominated by

filamentous bacteria. After operation in SBR for 1 week, compact aggregates appeared.

The filamentous bacteria gradually disappeared in the acetate-fed reactor; however, in the

glucose-fed reactor, filamentous bacteria still prevailed. Two weeks after the start-up, the

granular sludge with clear round outer shape was formed in both reactors. Although the

filamentous bacteria disappeared completely in acetate-fed reactor, they were still

predominant in glucose-fed reactor. This may imply that a high-carbohydrate feed

composed of glucose supports the growth of filamentous bacteria as reported in activated

sludge process previously (Chudoba, 1985). After operation for 3 weeks, aerobic granules

matured in both reactors. At this stage, both glucose- and acetate-fed granules had a very

regular round-shaped outer surface. The average aspect ratio of glucose-fed granules was

0.79 and 0.73 for acetate-fed granules. (Aspect ratio of a particle is the ratio of the lengths

of minor axis and major axis of an ellipse that is equivalent to the particle.) Compared to

acetate-fed granules, glucose-fed granules had a fluffy outer surface because of the

predominance of filamentous bacteria (Fig. 1). Scanning electron microscope (SEM)

observations further revealed that the glucose-fed mature aerobic granules indeed had a

filamentous dominant outer surface, while the acetate-fed aerobic granules had a very

Fig. 1. Macrostructures of glucose-fed (a) and acetate-fed (b) aerobic granules (Tay et al., 2001a).

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compact microstructure in which cells were tightly linked together and rodlike bacteria

were predominant (Fig. 2). It seems certain that aerobic granulation is a gradual process

involving the progression from seed sludge to compact aggregates, further to granular

sludge and finally to mature granules.

2.2. Factors affecting aerobic granulation

For cells in a culture to aggregate, a number of conditions have to be satisfied. This

section attempts to discuss the factors that may influence aerobic granulation.

2.2.1. Substrate composition

Aerobic granules have been successfully cultivated with a wide variety of substrates

including glucose, acetate, ethanol, phenol, and synthetic wastewater (Beun et al., 1999;

Peng et al., 1999; Tay et al., 2001a, 2003b; Moy et al., 2002; Jiang et al., 2002; Yang et al.,

in press; Schwarzenbeck et al., 2003). However, granule microstructure and species

diversity appear to be related to the type of carbon source. The glucose-fed aerobic

granules have exhibited a filamentous structure, while acetate-fed aerobic granules have

had a nonfilamentous and very compact bacterial structure in which a rodlike species

predominated. Aerobic granules have been also cultivated with nitrifying bacteria and an

inorganic carbon source (Tay et al., 2002b; Tsuneda et al., 2003). These nitrifying granules

showed an excellent nitrification ability. More recently, aerobic granules were also

successfully developed in laboratory-scale SBR for treating particulate organic matter-

rich wastewater (Schwarzenbeck et al., 2003).

2.2.2. Organic loading rate

The essential role of organic loading rate (OLR) in the formation of anaerobic granules

has been recognized. Relatively high organic loading rates facilitate the formation of

anaerobic granules in UASB systems. In contrast to anaerobic granulation, the accumu-

lated evidence suggests that aerobic granules can form across a very wide range of organic

loading rates from 2.5 to 15 kg chemical oxygen demand (COD)/m3 day (Moy et al., 2002;

Liu et al., 2003a). It seems that aerobic granulation is not sensitive to the organic loading

rate. Although the effect of organic loading rate on the formation of aerobic granules is

Fig. 2. Microstructures of glucose-fed (a) and acetate-fed (b) aerobic granules (Tay et al., 2001a).

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 537

insignificant, the physical characteristics of aerobic granules depend on the organic

loading rate. The mean size of aerobic granules increased from 1.6 to 1.9 mm with the

increase of the organic loading from 3 to 9 kg COD/m3 day (Liu et al., 2003a). A similar

trend was also observed in 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., in press). Similarly, in anaerobic granulation, a

high organic loading rate has been found to reduce strength of anaerobic granules; that is,

partial loss of structural integrity and disintegration can occur at high organic loading rates

(Morvai et al., 1992; Quarmby and Forster, 1995). It should be stressed that an increased

organic loading rate can raise the biomass growth rate and this in turn reduces the strength

of the three-dimensional structure of the microbial community (Liu et al., 2003c).

2.2.3. Hydrodynamic shear force

Evidence shows that a high shear force favors the formation of aerobic granules and

granule stability (Shin et al., 1992; Tay et al., 2001a). It was found that aerobic granules

could be formed only above a threshold shear force value in terms of superficial upflow air

velocity above 1.2 cm/s in a column SBR, and more regular, rounder, and compact aerobic

granules were developed at high hydrodynamic shear force (Tay et al., 2001a). The

granule density and strength were also proportionally related to the shear force applied

(Tay et al., 2003c). These observations may imply that the structure of aerobic granules is

mainly determined by the hydrodynamic shear force present in a bioreactor. However, it is

well known that extracellular polysaccharides can mediate both cohesion and adhesion of

cells and play a crucial role in maintaining the structural integrity in a community of

immobilized cells. Tay et al. (2001a) reported that the production of extracellular

polysaccharides was closely associated with the shear force and the stability of aerobic

granules was found to be related to the production of extracellular polysaccharides (Tay et

al., 2001c). The extracellular polysaccharides content normalized to protein content,

increased with the shear force estimated in terms of superficial upflow air velocity. Thus,

a high shear force stimulated bacteria to secrete more extracellular polysaccharides. In fact,

shear force-induced production of extracellular polysaccharides has been observed in

biofilms (Ohashi and Harada, 1994). Consequently, the enhanced production of extracel-

lular polysaccharides at high shear can contribute to the compact and stronger structure of

aerobic granules. Effects of shear on microorganisms and aggregates have been discussed

further elsewhere (Chisti, 1999a).

2.2.4. Settling time

In a SBR, wastewater is treated in successive cycles each lasting a few hours. At

the end of every cycle, the biomass is settled before the effluent is withdrawn. The

settling time acts as a major hydraulic selection pressure on microbial community. A

short settling time preferentially selects for the growth of fast settling bacteria and the

sludge with a poor settleability is washed out. Qin et al. (2004) reported that aerobic

granules were successfully cultivated and became dominant only in the SBR operated

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563538

at a settling time of 5 min. Mixtures of aerobic granules and suspended sludge were

observed in the SBRs run at settling times of 20, 15, and 10 min. The production of

extracellular polysaccharides was stimulated and the cell surface hydrophobicity

improved significantly at short settling times. These findings illustrate the fact that

aerobic granulation is driven by selection pressure and the formation and characteristics

of the granules may be controlled by manipulating the selection pressure. Therefore,

choice of an optimal settling time is very important in aerobic granulation. Generally,

the mature aerobic granules tend to settle within 1 min, leaving a clear supernatant in

the reactor (Tay et al., 2001a). The easily retainable biomass in the reactor ensures a

faster and more efficient removal of organic pollutants in wastewater. Granules with

excellent settling properties are essential for the effective functioning of biological

systems treating wastewater.

2.2.5. Hydraulic retention time

In aerobic granulation, the light and dispersed sludge is washed out and the relatively

heavy granules are retained in the reactor. The SBR cycle time represents the frequency of

solids discharge through effluent withdrawal, or the so-called washout frequency, and it is

related to the hydraulic retention time (HRT) at a given exchange ratio. The latter is

defined as the volume of effluent discharged divided by the working volume of the SBR.

A short cycle time would suppress the growth of suspended solids because of frequent

washout of the suspended material. However, if the SBRs are run at an extremely short

cycle time, sludge loss has been observed through hydraulic washout because bacterial

growth has been unable to compensate. As a result, a complete washout of sludge blanket

occurs and leads to a failure of microbial granulation. Thus, the HRT should be short

enough to suppress the suspended growth, but long enough for microbial growth and

accumulation.

By its nature, a SBR is cyclic in operation. The SBR cycle time can serve as a main

hydraulic selection pressure on the microbial community in the system. Tay et al.

(2002b) investigated the effect of hydraulic selection pressure on the development of

nitrifying granules in column-type sequencing batch reactors. No nitrifying granulation

was observed in the SBR operated at the longest cycle time of 24 h because of a weak

hydraulic selection pressure while the nitrifying sludge was washed out in the SBR run

at the shortest cycle time of 3 h, which also prevented the development of nitrifying

granules. Excellent nitrifying granules were successfully developed in the SBR operated

at cycle times of 6 and 12 h. A short cycle time stimulates microbial activity and

production of cell polysaccharides and also improves the cell hydrophobicity. These

hydraulic selection pressure-induced microbial changes favor the formation of nitrifying

granules.

2.2.6. Aerobic starvation

The SBR operation is a sequencing cycle of feeding, aeration, settling, and

discharging of supernatant fluid. As a result, microorganisms growing in the SBR are

subject to periodic fluctuations in the environmental conditions. During operation cycles,

an important period of aerobic substrate starvation has been identified (Tay et al.,

2001a). The waste degradation time required tends to reduce with the increase in the

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 539

number of operation cycles. The aeration period of the operation actually consists of two

phases: a degradation phase in which the substrate is depleted to a minimum, followed

by an aerobic starvation phase in which the external substrate is no longer available.

Under starvation conditions, bacteria became more hydrophobic which facilitates

microbial adhesion (Tay et al., 2001a). It is likely that aggregation is a strategy of

cells against starvation. It appears that the microorganisms are able to change their

surface characteristics when they face starvation (Tay et al., 2001a). Bossier and

Verstraete (1996) reported that under starvation conditions, bacteria become more

hydrophobic which likely facilitates adhesion or aggregation. Such changes contribute

to microbial ability to aggregate. Thus, starvation plays a role in the microbial

aggregation process and leads to stronger and denser granules. Although the periodical

starvation in SBR is important for microbial aggregation, the contribution of other

operation conditions should not be neglected.

2.2.7. Presence of calcium ion in feed

Jiang et al. (2003) reported that addition of Ca2 + accelerated the aerobic granulation

process. With addition of 100 mg Ca2 +/l, 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 poly-

saccharides contents. 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 nondeformable polymeric gel-like

matrix.

2.2.8. Intermittent feeding strategy

A periodic starvation occurs during the course of SBR operation (Tay et al., 2001a).

This periodic starvation has been shown to have a profound effect on cell hydrophobicity,

which is a key factor that affects aerobic granulation (Liu et al., 2003b). Tay et al. (2001a)

found that cell surface hydrophobicity was proportionally related to the starvation time in

SBR. McSwain et al. (2003) recently developed an operation strategy to enhance aerobic

granulation by intermittent feeding; that is, different filling times were applied to SBR

reactors to vary the feast-fast cycle. A feast-fast cycle or pulse feeding of the SBR favored

the formation of compact and dense aerobic granules. Under starvation conditions, bacteria

became more hydrophobic and this in turn facilitated microbial adhesion and aggregation

(Bossier and Verstraete, 1996).

2.2.9. Dissolved oxygen, pH and temperature

Dissolved oxygen (DO) concentration is an important variable that influences the

operation of aerobic wastewater treatment systems. Aerobic granules have formed at DO

concentration as low as 0.7 to 1.0 mg/l in a SBR (Peng et al., 1999). In addition, aerobic

granules have been successfully developed at DO concentrations of >2 mg/l (Tay et al.,

2002c; Yang et al., in press). It appears, therefore, that DO concentration is not a decisive

variable in the formation of aerobic granules. Concerning the roles of the reactor pH and

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563540

temperature on aerobic granulation, detailed studies are lacking. Our unpublished work

suggests that these effects are not as important in aerobic granulation as they are in

anaerobic granulation.

2.2.10. Seed sludge

Aerobic granular sludge SBRs have been seeded with conventional activated sludge.

In anaerobic granulation, there is evidence that the characteristics of the seed sludge

profoundly influence the formation and properties of anaerobic granules. The important

factors that determine the quality of seed sludge for aerobic granulation appear to

include the macroscopic characteristics, settleability, surface properties (a high surface

hydrophobicity and low surface charge density are preferred), and microbial activity.

Little information is available on the role of seed sludge in aerobic granulation.

2.2.11. Reactor configuration

In almost all cases reported, aerobic granules were produced in column-type upflow

reactors. 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). Column-type upflow

reactor and completely mixed tank reactor (CMTR) have very different hydrodynamic

behaviors in terms of the interaction between flow and microbial aggregates. The air or

liquid upflow in column reactors can create a relatively homogenous circular flow and

localized vortexing 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. In a column-type

upflow reactor 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. However, in CMTRs microbial aggregates stochastically move with dispersed

flow in all directions. Thus, microbial aggregates are subject to varying localized

hydrodynamic shear force, upflow trajectories and random collisions. Under such circum-

stances, only flocs with irregular shape and size instead of regular granules occasionally

form (Liu and Tay, 2002). 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.

2.2.12. Inhibition to aerobic granulation

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

granulation in a SBR fed with acetate as the sole carbon source. Aerobic granules formed

only when the free ammonia concentration was less than 23.5 mg/l and nitrification was

completely inhibited at a free ammonia concentration of >10 mg/l. The specific oxygen

utilization rates (SOURs) of heterotrophic and nitrifying bacteria were reduced by a factor

2.5 and 5.0 as the free ammonia concentration increased from 2.5 to 39.6 mg N/l. The high

free ammonia concentration resulted in a significant decrease of cell hydrophobicity and

also repressed the production of extracellular polysaccharides. Changes in hydrophobicity

and polysaccharide production were likely responsible for the failure of aerobic granula-

tion at high free ammonia concentrations. Yang et al. (2004b) demonstrated that free

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ammonia could hinder the formation of aerobic granules by inhibiting the energy

metabolism of microorganisms.

2.3. Characteristics of aerobic granules

Compared to the loose, fluffy, and irregular conventional activated sludge flocs, the

aerobic granular sludge is known to: (i) have denser and stronger microbial structure; (ii)

have regular, smooth round shape, and a clear outer surface; (iii) be visible as separate

entities in the mixed liquor during both the mixing and the settling phases; (iv) have a high

biomass retention and excellent settleability; (v) be capable to withstanding high flow

rates; (vi) be able to withstand high organic loading rates; (vii) be less vulnerable than the

suspended sludge to the toxicity of organic chemicals and heavy metals in wastewater. The

excellent settleability of aerobic granules simplifies the separation of treated effluent from

the granular sludge.

2.3.1. Morphology

Microscopic examination shows that the morphology of the aerobic granular sludge is

completely different from the floclike sludge. The shape of the granules is nearly

spherical with a very clear outline (Peng et al., 1999; Tay et al., 2001a,c; Zhu and

Wilderer, 2003). The granule size is an important parameter in the characterization of

aerobic granulation. The average diameter of aerobic granules varies in the range of 0.2

to 5 mm. This is mainly due to a balance between growth and abrasive detachment due

to the relatively strong hydrodynamic shear force in aerobic reactors (Liu and Tay, 2002;

Liu et al., 2003g). Hydrodynamic shear forces are known to control the prevailing size

of the suspended biosolids in many situations (Chisti, 1999a). Methods of estimating the

magnitudes of these forces under various conditions of operation have been discussed by

Chisti (1999a).

2.3.2. Settleability

The settling properties of aerobic granules determine the efficiency of solid–liquid

separation that is essential for the proper functioning of wastewater treatment systems.

The sludge volume index (SVI) of aerobic granules can be lower than 50 ml/g, which is

much lower than that of conventional bioflocs (Liu et al., 2003f; Qin et al., 2004). This

implies that from an engineering perspective, the settleability of sludge can be improved

significantly through the formation of aerobic granules so that it can be settled in a

more compact clarifier. The settling velocity of aerobic granules is associated with

granule size and structure and is as high as 30 to 70 m/h. This is comparable with that

of the UASB granules, but is at least three times higher than that of activated sludge

flocs (typical settling velocity of around 8 to 10 m/h). The high settling velocities of

aerobic granules allow the use of relatively high hydraulic loads to the reactors without

having to worry about washout of biomass (Beun et al., 2000; Tay et al., 2001b). Thus,

aerobic granulation can lead to more biomass retention in the reactor and this can

enhance the performance and stability of the reactor. A high concentration of the

retained biomass ensures a faster degradation of pollutants and relatively compact

reactors.

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2.3.3. Density and strength

The specific gravity of aerobic granules typically ranges from 1.004 to 1.065 (Etterer

and Wilderer, 2001; Tay et al., 2001a; Yang et al., in press). The granules with a high

physical strength withstand high abrasion and shear. The physical strength, expressed as

integrity coefficient (i.e., the ratio of residual granules to the total weight of the granular

sludge after 5 min of shaking at 200 rpm on a platform shaker, expressed as percent), is

higher than 95% for the aerobic granules grown on glucose and acetate (Tay et al., 2002c).

The physical strengths of aerobic and anaerobic granules are comparable. Aerobic

granules with smaller sizes tend to be more compact compared to larger aerobic granules

(Toh et al., 2003; Yang et al., 2004a).

2.3.4. Cell surface hydrophobicity

Cell surface hydrophobicity is an important affinity force in cell self-immobilization

and attachment processes (Pringle and Fletcher, 1983; Kos et al., 2003; Liu et al., 2003b).

The role of cell surface hydrophobicity in the formation of aerobic granules has not been

clear. Liu et al. (2003b) have linked the formation of heterotrophic and nitrifying granules

to the cell surface hydrophobicity. The hydrophobicity of granular sludge was nearly

twofold higher than that of conventional bioflocs. A high shear force or hydraulic selection

pressure imposed on microorganisms resulted in a significant increase in the cell surface

hydrophobicity, while the cell surface hydrophobicity seemed not to be sensitive to the

changes in the organic concentrations or loading rates in the range of 500 to 3000 mg

COD/l.

2.3.5. Specific oxygen utilization rate

In the environmental engineering field, microbial activity of microorganisms is

characterized by the specific oxygen utilization rate (SOUR). A very wide range of SOUR

values for aerobic granules have been reported (Morgenroth et al., 1997; Tay et al., 2001b;

Yang et al., 2003b; Zhu and Wilderer, 2003). The SOUR has been found to increase with

the increase in shear force specified in terms of superficial air velocity. It is obvious that

the increased shear force can stimulate the respiration activities of microorganisms in a

very significant manner (Tay et al., 2001b). This may have to do with the effect of

hydrodynamic shear on increasing the rate of oxygen transfer at the granule–liquid

interface (Chisti, 1999b). The biochemical reactions associated with bacterial metabolism

show in an approximately linear relationship between oxygen utilization and carbon

dioxide production: that is, relatively less cell mass is produced at high oxygen utilization

as the metabolism is faster and more of the substrate is converted to carbon dioxide. The

microbial activity represented by SOUR is inversely related to the hydraulic selection

pressure in terms of the settling time (Qin et al., 2004). The shorter settling time tends to

significantly stimulate the respiratory activity of these bacteria. This implies that the

microorganisms attempt to regulate their energy metabolism in response to changes in

hydraulic selection pressure.

During long storage of aerobic granules under anaerobic conditions, SOUR is used as

an indicator for evaluating the metabolic activity of the granular sludge. A decreased

SOUR indicates a reduced metabolic activity of granules after storage (Tay et al., 2002c;

Zhu and Wilderer, 2003). Once a reactor restarts, the activity of the granules returns to

Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563542

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 543

normal within a few days. When treating the wastewater containing toxic chemicals such

as phenol, SVI and SOUR values are necessary for monitoring the physiology changes of

granule. Biomass growth and substrate removal are linked to batch measurement of

SOUR. From the SOUR values, we can estimate the highest permissible loading rate.

Therefore, SOUR is an important characteristic for assessing the ability of aerobic granules

to handle high-strength industrial wastewaters.

2.3.6. Storage stability

The loss of granule stability and activity during an extended idling period is related to

the storage temperature. A high storage temperature accompanied with the absence of

external substrate can lead to endogenous respiration and a rapid disintegration of the

granules. Tay et al. (2002c) examined the effect of storage period of glucose- and acetate-

fed aerobic granules on granule activity and structure and found that the aerobic granules

grown on glucose only lost about 60% of initial metabolic activity in terms of specific

oxygen utilization rate (SOUR). In comparison, the activity of the acetate-fed granules was

reduced by about 90%. Broth granules were stored for 4 months in tap water at 4 jC. Zhuand Wilderer (2003) found that after 7 weeks of storage of aerobic granules in ambient

environment, the granules could regain microbial activity within a week.

Ng (2002) studied the effect of different storage solutions, e.g., tap water, physiological

and nutrient solutions, on the stability and activity of aerobic granules stored for a period

of 8 weeks at 4 jC. The granules stored in all storage media became more irregular and

smaller at the end of week 8 as compared to fresh granules at week 0 and released soluble

organic material with an accompanying drop in the solution pH due to cell hydrolysis. A

partial desegregation of granules was observed. The decay rate of aerobic granules during

the anaerobic storage appears to correlate with the content of volatile solids (VS) in the

granules. In summary, the loss of granule activity and structural integrity during storage

depend on the storage temperature, duration, the storage medium, and the characteristics of

the granules.

Compared to fresh granules, the strength of the stored granules has been observed to

decrease by 7–8% for glucose- and acetate-fed aerobic granules after 4 months storage

at 4 jC (Tay et al., 2002c). However, the size of aerobic granules stored in tap water and

physiological solution decreased by 34%, and 22%, respectively, at the end of 8-week

storage (Ng, 2002). In addition, the apparent color of aerobic granules stored in tap

water and physiological solution turned from brownish-yellowish (fresh granules) to

gray and dark black at the end of storage. The granules that were stored in the phosphate

buffered saline (PBS) experienced the least change in color. It is reasonable to expect

that the change of apparent color during storage results from anaerobic metabolism of

the granules.

Zhu and Wilderer (2003) reported that aerobic granular sludge could be stored for up to

7 weeks without loss of integrity and metabolic potential of the granules; that is, even

without the substrate and oxygen supply, the granular sludge remained intact. The

metabolic activity of the stored aerobic granules recovered progressively as soon as the

reactor was resupplied with substrate and oxygen. This implies that aerobic granular

sludge can be kept on standby in case the wastewater flow and concentration fall below the

normal levels or when toxic effects occur occasionally. Clearly, the use of aerobic

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granulation is applicable to situations where substrate availability varies greatly with time.

This is a very common scenario.

2.4. Microbial structure and diversity

In wastewater treatment system, studying the microbial structure and diversity is

essential for a comprehensive understanding of the microbial community. However, such

a study is quite complicated. This is especially so in the microbial ecology of aerobic

granules.

2.4.1. Microbial structure

Confocal laser-scanning microscopy (CLSM) has been used with different oligonucle-

otide probes, specific fluorochromes, and fluorescent microspheres for studying the

microstructure of aerobic granules (Tay et al., 2002d, 2003a; Toh et al., 2003; Jang et

al., 2003; Meyer et al., 2003). The obligate aerobic ammonium-oxidizing bacterium

Nitrosomonas spp. was found mainly at a depth of 70 to 100 Am from the granule surface,

and aerobic granules contained channels and pores that penetrated to a depth of 900 Ambelow the granule surface. The porosity peaked at depths of 300 to 500 Am from the

granule surface (Tay et al., 2002d, 2003a). These channels and pores would facilitate the

transport of oxygen and nutrients into and metabolites out of the granules. Polysaccharide

formation peaked at a depth of 400 Am below the granule surface. The anaerobic

bacterium Bacteroides spp. also detected at a depth of 800 to 900 Am from the granule

surface (Tay et al., 2002e), while a layer of dead microbial cells was located at a depth of

800 to 1000 Am (Toh et al., 2003). In order to fully utilize the aerobic microorganisms in

the granules, the optimal diameter for aerobic granules should be less than 1600 Am, which

is twice the distance from the granule surface to the anaerobic layer (Tay et al., 2002d).

Consequently, smaller granules will be more effective for aerobic wastewater treatment as

these granules have more live cells within a given volume of granules.

More recently, Liu et al. (2004) observed the mushroomlike structure of aerobic

granules developed at high substrate N/COD ratios. The nitrifying population was mainly

located at a depth of 70 to 100 Am from the surface of the granule (Tay et al., 2002d).

Research has shown that biofilms of mixed bacterial communities can form thick layers

consisting of differentiated mushroomlike structures (Costerton et al., 1981) that are

similar to structures observed in aerobic granules. There is strong evidence that bacteria

sense and move towards nutrients (Prescott et al., 1999). It has been demonstrated that

biofilms can form the mushroomlike structures by simply changing the diffusion rate;

that is, the biofilm structure is largely determined by nutrient concentration (Wimpenny

and Colasanti, 1997). Since nitrifying bacteria are slow-growing, the mushroomlike

structures seem to result from the demand of nitrifying population for nutrients. These

structures improve the access of the nitrifying population to nutrients. As Watnick and

Kolter (2000) noted, in mixed biofilms, bacteria distribute themselves according to who

can survive best in the particular microenvironment and the high complexity of the

resulting microbial community appears to be beneficial to its stability. Consequently, the

distribution of different microbial populations in a granule may have an effect on its

stability.

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2.4.2. Microbial diversity

Microbial diversity of aerobic granules has been studied by molecular biotechnology

techniques (Yi et al., 2003; Tay et al., 2002d; Jang et al., 2003; Meyer et al., 2003; Tsuneda

et al., 2003). Heterotrophic, nitrifying, denitrifying, P-accumulating bacteria, and glyco-

gen-accumulating bacteria have been identified in aerobic granules developed under

different conditions (Jang et al., 2003; Meyer et al., 2003; Tsuneda et al., 2003; Lin et

al., 2003; Liu et al., 2003f; Yang et al., 2003b). The microbial diversity of aerobic granules

is closely related to the composition of culture media, in which they are developed and

structure of aerobic granules. Anaerobiosis and dead cells have been documented at the

centers of aerobic granules (Tay et al., 2002e). The presence of anaerobic bacteria in

aerobic granules is likely to result in the production of organic acids and gases within the

granules. These end products of anaerobic metabolism can destroy the granules or at least

diminish their long-term stability.

2.5. Mechanisms of aerobic granulation

For bacteria to form aerobic granules a number of conditions need to be met and the

physical, chemical, and biological forces contributing to granulation need to be viewed in

combination. Liu and Tay (2002) proposed a model for the aerobic granulation as

consisting of the following steps. Step 1. Physical movement to initiate bacterium-to-

bacterium contact. The factors involved in this step are hydrodynamics, diffusion mass

transfer, gravity, thermodynamic effects, and cell mobility. Step 2. Stabilization of the

multicell contacts resulting from the initial attractive forces. These attractive forces are

physical forces (e.g., Van der Waals forces, opposite charge attraction, thermodynamically

driven reduction of the surface free energy, surface tension, hydrophobicity, filamentous

bacteria that can bridge individual cells), chemical forces, and biochemical forces

including cell surface dehydration, cell membrane fusion, signaling, and collective action

in bacterial community. Step 3. Maturation of cell aggregation 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. Step 4. Shaping of the steady state three-dimensional structure of

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

Cell surface hydrophobicity may play a crucial role in the initiation of aerobic

granulation (Liu et al., 2003b). According to the thermodynamic theory, increasing cell

surface hydrophobicity would cause a corresponding decrease in the excess Gibbs energy

of the surface and promote cell–cell interaction to further drive the self-aggregation of

bacteria out of suspending liquid (hydrophilic phase). Hydrophobic binding is considered

of prime importance for cell–cell attachment (Pringle and Fletcher, 1983; Bos et al., 1999;

Liu et al., 2003h). A high cell surface hydrophobicity would result in a stronger cell–cell

interaction and the formation of a denser structure. Extracellular polysaccharides can

mediate both cohesion and adhesion of cells and play a crucial role in maintaining the

structural integrity in a community of immobilized cells. The polysaccharide contents of

aerobic granules tend to be much higher than that of sludge flocs (Tay et al., 2001c). Cell

polysaccharides also contribute greatly to aerobic granulation. Qin et al. (2004) observed

that aerobic granules were successfully cultivated in the SBRs operated at a settling time

Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 545

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563546

of < 15 min, while only bioflocs appeared in the reactor run at the a 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 (Wimpenny and Colasanti, 1997). A cell can determine its position

in a concentration gradient of an extracellular signal factor and uses this to modify its

development (Gurdon and Bourillot, 2001). Research on cell–cell communication (Davies

et al., 1998; Pratt and Kolter, 1999; Ben-Jacob et al., 2000) 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.

2.6. Applications of aerobic granulation technology

The performance of a biological system for wastewater treatment depends significantly

on the active biomass concentration, the overall biodegradation rates, the reactor configu-

ration, and the feeding rates of the pollutants and oxygen. 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. Treatment

capacities can be easily varied to accommodate varying loading rates, wastewater compo-

sition, and treatment goals by bioaugmentation with specifically developed granules.

2.6.1. High-strength organic wastewater treatment

Granulation of the sludge can lead to high biomass retention in the reactor because of

the compact and dense structure of the granules. Biomass concentrations as high as 6.0 to

12.0 g/l have been obtained in SBRs operated with a volumetric exchange ratio of 50%

(Tay et al., 2002a,c). The feasibility of applying aerobic granulation technology for the

treatment of high-strength organic wastewaters was demonstrated by Moy et al. (2002),

who examined the ability of aerobic granules to sustain high organic loading rates by

introducing step increases in organic loading only after the COD removal efficiencies had

stabilized at values of >89% for at least 2 weeks. Aerobic granules cultivated this way on

glucose were exposed to organic loading rates that were gradually raised from 6.0 to 9.0,

12.0, and 15.0 kg COD/m3 day. Aerobic granules were able to sustain the maximum

organic loading rate of 15.0 kg COD/m3 day while removing more than 92% of the COD.

The granules initially exhibited a fluffy loose morphology dominated by filamentous

bacteria at low loadings and evolved into smooth irregular shapes characterized by folds,

crevices, and depressions at higher loadings. These irregularities were thought to allow for

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better diffusion and penetration of nutrients into the interior of the granule. Mass transfer

of nutrients was also enhanced by the higher substrate concentration that existed in the

bulk wastewater at higher loadings. These factors enabled the aerobic granules to sustain

high organic loading rates without compromising the granule integrity.

2.6.2. Simultaneous organic and nitrogen removal

Complete nitrogen removal involves nitrification and denitrification. Nitrite and nitrate

produced from nitrification are reduced to gaseous nitrogen by denitrifiers. Yang et al. (in

press, 2004a) investigated the simultaneous removal of organics and nitrogen by aerobic

granules. Heterotrophic, nitrifying, and denitrifying populations were shown to success-

fully coexist in microbial granules. Increased substrate N/COD ratio led to significant shifts

among the three populations within the granules. Coexistence of heterotrophic and

nitrifying populations in aerobic granules was also observed by Jang et al. (2003).

Enhanced activities of nitrifying and denitrifying populations were achieved in granules

developed at high substrate N/COD ratios; however, the heterotrophic populations in

granules tended to decrease with the increase of substrate N/COD ratio. 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 (Yang et al., 2003b).

Similar phenomena were reported by Beun et al. (2001). It appears that complete organics

and nitrogen removal can be efficiently and stably achieved in a single granules-based SBR.

2.6.3. Phosphorous removal

Environmental regulations in many jurisdictions require a reduction of phosphorus

concentration in wastewater to levels of 0.5 to 2.0 mg/l before discharge. The well-known

enhanced biological phosphorus removal (EBPR) process removes P without the use of

chemical precipitation and is an economical and reliable option for P removal from

wastewater. The EBPR process operates on the basis of alternating anaerobic and aerobic

conditions with substrates feeding limited to the anaerobic stage. Most EBPR processes

are based on suspended biomass cultures and require large reactor volumes. Although full-

scale experience shows a strong potential of the EBPR, difficulties in assuring stable and

reliable operation have also been recognized. Often, the reasons for failure of biological

phosphorus removal are not clear (Barnard et al., 1985; Bitton, 1999).

In attempts to overcome the problems associated with the conventional bioremoval of P,

Lin et al. (2003) successfully developed phosphorus-accumulating microbial granules in

SBRs operated at substrate P/COD ratios ranging from 1/100 to 10/100 by weight. The

soluble COD and PO4–P profiles showed that the granules had typical P-accumulating

characteristics, with concomitant uptake of soluble organic carbon and the release of

phosphate in the anaerobic stage, followed by rapid phosphate uptake in the aerobic stage.

The size of P-accumulating granules exhibited a decreasing trend with the increase of

substrate P/COD ratio. The structure of the granules became more compact and dense as

the substrate P/COD ratio increased. The P uptake by granules was in the range of 1.9% to

9.3% by weight, or comparable to that of the conventional enhanced biological phospho-

rus removal (EBPR) processes. These results will certainly spur the further development of

novel granule-based EBPR technologies.

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2.6.4. Phenolic wastewater treatment

Phenol is a toxic and inhibitory substrate, but also a carbon source for the bacteria. The

consequence of the presence of phenol in biological wastewater treatment is process

instability, which can lead to the washout of the microorganisms (Allsop et al., 1993). In

low concentrations, phenol is biodegradable, but high concentrations can kill phenol-

degrading bacteria. Industrial wastewaters from fossil fuel refining, pharmaceutical and

pesticide processing are the major sources of phenolic pollution. Jiang et al. (2002, 2004)

investigated the feasibility of treating phenol-containing wastewater with aerobic granular

sludge. Granular sludge is less susceptible to toxicity of phenol because much of the

biomass in the granules is not exposed to the same high concentration as present in the

wastewater. The phenol-degrading aerobic granules displayed an excellent ability to

degrade phenol (Jiang et al., 2002, 2004). For an influent phenol concentration of 500

mg/l, a stable effluent phenol concentration of less than 0.2 mg/l was achieved in the

aerobic granular sludge reactor (Jiang et al., 2002, 2004). The high tolerance of aerobic

granules to phenol can be exploited in developing compact high-rate treatment systems for

wastewaters loaded with a high concentration of phenol. Aerobic granules may prove

powerful bioagents for removing other inhibitory and toxic organic compounds from high-

strength industrial wastewaters. Aerobic granules appear to be highly tolerant of toxic

heavy metals (Xie, 2003).

2.6.5. Biosorption of heavy metals by aerobic granules

Heavy metals are often found in a wide variety of industrial wastewaters. More

stringent metal concentration limits are being established in view of their relatively high

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

These include marine algae, fungi, waste activated sludge, and digested sludge (Lodi et al.,

1998; Taniguchi et al., 2000; Valdman and Leite, 2000). In view of the physical

characteristics of aerobic granules as discussed earlier, these granules are ideal biosorbent

for heavy metals. The granules are physically strong and have large surface area and high

porosity for adsorption. In addition, the granules can be easily separated from the liquid

phase after biosorption capacity is exhausted. The biosorption of Zn2 + and Cd2 + by

aerobic granules has been reported (Liu et al., 2002, 2003c,d). The biosorption of Zn2 +

was shown to relate to both the initial Zn2 + and granule concentrations (Liu et al., 2002).

The concentration gradient of Zn2 + was the main driving force for Zn2 + biosorption by the

granules. The maximum biosorption capacity for Zn2 + was 270 mg/g of granules. For

Cd2 +, this capacity was 566 mg/g (Liu et al., 2003d).

Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563548

3. Anaerobic granulation technology

3.1. Mechanisms and model for anaerobic granulation

A number of models for anaerobic granulation have been developed over the past 20

years to enhance the understanding of the mechanisms of anaerobic granulation. These

models mainly include the inert nuclei model, divalent cation-bridge model, proton

translocation–dehydration model, extracellular polymer model, ‘‘Spaghetti’’ model,

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syntrophic microcolony model, thermodynamic models, quorum sensing model, etc. These

models are discussed by Liu et al. (2003i). The factors influencing anaerobic granulation

are discussed next.

3.2. Factors influencing anaerobic granulation and its performance

A major problem associated with the upflow anaerobic sludge blanket (UASB) reactors

is the long start-up period (2–4 months or longer) required for the development of

anaerobic granules. Enhanced granulation processes are desirable to reduce the space time

requirements of bioreactors. This section discusses the main factors that influence

anaerobic granulation and its performance.

3.2.1. Upflow liquid velocity and hydraulic retention time

It has been observed that anaerobic granulation can proceed well at relatively high

liquid upflow velocity, but does not occur under conditions of low hydrodynamic shear

(Alphenaar et al., 1993; Arcand et al., 1994; O’Flaherty et al., 1997; Alves et al., 2000).

According to Alphenaar et al. (1993), granulation in UASB reactors is favored by a

combination of high upflow liquid velocity and short hydraulic retention time (HRT).

Usually, the effects of upflow liquid velocity on anaerobic granulation are explained by the

selection pressure theory (Hulshoff Pol et al., 1988). A long HRT accompanied with a low

upflow liquid velocity may allow dispersed bacterial growth and be less favorable for

microbe granulation. In contrast, a short HRT combined with a high upflow liquid velocity

can lead to washout of nongranulation competent bacteria and thus promote sludge

granulation. These findings are consistent with those reported for aerobic granulation.

Evidence shows that flocculent anaerobic sludge can be converted to a relatively active

anaerobic granular sludge by manipulating hydraulic stress and the settleability of the

anaerobic granules. Granules with improved SVI and sludge settling velocity develop

when the upflow liquid velocity is increased (Noyola and Mereno, 1994). While the

upflow liquid velocity has a significant positive effect on the mean granule size, its effect

on the specific washout rate of the smaller particles is insignificant (Arcand et al., 1994).

3.2.2. Organic loading rate

The organic substrate loading rate (OLR) describes the degree of starvation of the

microorganisms in a biological system. At a low OLR, microorganisms are subject to

nutrient starvation, while a high OLR sustains fast microbial growth (Bitton, 1999).

Evidence shows that anaerobic granulation can be accomplished by gradually increasing

the OLR during the start-up (Hulshoff Pol, 1989; Kosaric et al., 1990; Campos and

Anderson, 1992; Tay and Yan, 1996). It is critical to select a reasonably high OLR during

start-up, to ensure rapid granulation and a stable treatment process. A simple and practical

strategy for rapid start-up of anaerobic granular sludge reactors is to increase the OLR to

attain only 80% reduction of biodegradable chemical oxygen demand (COD) with

supplementary monitoring of effluent for washout of suspended solids (de Zeeuw, 1988;

Fang and Chui, 1993). However, if the applied OLR is too high during start-up of UASB

reactors, the biogas production rate increases and this causes hydrodynamic turbulence and

can lead to washout of the seed sludge from the reactor. This often leads to unsuccessful

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start-up of a UASB reactor. As discussed earlier, compared to aerobic granulation, the

anaerobic granulation is highly sensitive to the organic loading applied. This is probably

because of different growth characteristics of aerobic and anaerobic bacteria.

3.2.3. Characteristics of feed

Characteristics of the feed are considered a key factor influencing the formation,

composition, and structure of anaerobic granules. As discussed earlier, the effect of feed on

aerobic granulation seems to be insignificant because of fast growth of aerobic bacteria.

Based on their free energy of oxidation, organic substrates can be roughly classified into

high-energy and low-energy feeds. During the UASB start-up period, high-energy

carbohydrate feeding can sustain the acidogens and facilitate the formation of extracellular

polymers. The more readily the acidogens take up and metabolize the substrate, the more

rapidly the proton pumps will be activated, and sooner the methanogens will obtain the

substrate (Tay et al., 2000). Thus, the rapid growth of acidogens due to the presence of

high-energy substrate in the influent would facilitate the overall process of sludge

granulation in the UASB reactors.

The granules grown on volatile fatty acid mixture (acetate, propionate, and butyrate)

under mesophilic conditions can be classified into three distinct types according to the

predominant acetate utilizing methanogens present: (1) rod-type granules, which are

mainly composed of rod-shaped bacteria in fragments of about four to five cells

resembling Methanothrix; (2) filament-type granules, which consist predominantly of

long multicellular rod-shaped bacteria; and (3) sarcina-type granules, which develop when

a high concentration of acetic acid is maintained in the reactor (Hulshoff Pol et al., 1983;

de Zeeuw, 1984). A trend has been observed towards increasing diversity of methanogenic

subpopulations with an increasing complexity of the waste composition. At least four

distinct microcolonies have been observed in granules treating brewery wastewater (Wu,

1991). One of these microcolonies was composed of Methanothrix-like rods only, while

the other microcolonies consisted of hydrogen–carbon dioxide utilizing Methanobacte-

rium-like rods juxtapositioned with three different rods-shaped synthrophs (Hickey, 1991).

Full-scale UASB experience confirms that anaerobic sludge granulation occurs in many

different types of wastewaters. Because of the extremely low growth rate of anaerobic

bacteria, the energy content of the substrate are important for anaerobic granulation;

however, the complexity of substrate also exerts a selection pressure on the microbial

diversity in anaerobic granules. This selection pressure may in turn influence the formation

and microstructure of granules through its effect on the food chain and community

signaling communications.

3.2.4. Seed sludge

Theoretically any medium containing the proper bacterial flora can be used as seed

sludge for a UASB reactor. Potential seed sludges include manure, fresh water sediments,

septic tank sludge, digested sewage sludge, and surplus sludge from anaerobic treatment

plants. The start-up of UASB reactors using digested sewage sludge as seed may take

several months (de Zeeuw, 1984). A long start-up period is the major deterrent to the use

of UASB systems. With respect to the use of digested sewage sludge as seed, de Zeeuw

(1984) found that heavy, relatively inactive sludge was preferred over lighter, more active

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sludge because of differences in washout. Two types of sludge washouts were distin-

guished. These were erosion washout and sludge bed washout. Sludge bed erosion

washout is a selective washout based on differences in settleability. Although digested

sewage sludge is usually used for the start-up of UASB reactors, other types of seed

sludges have been used successfully when granular sludge for seeding is unavailable. Wu

et al. (1987) used aerobic activated sludge from a sewage treatment plant and primary

sludge from an aerobic plant treating textile dyeing wastewater for start-up.

Addition of a small amount of granules to nongranular inoculum is proven to stimulate

the granulation process (Hulshoff Pol et al., 1983; Xu and Tay, 2002). This is probably a

consequence of supplying a specific inoculum that is responsible for granulation. UASB

systems can also be started up using existing granules. This lends an advantage to the

UASB start-up process, although a successful start-up is not dependent on the use of

precultivated granules. When feasible, inoculation with a large amount of seed granular

sludge from a healthy UASB reactor is desirable. However, the availability of granular

seed sludge is limited and the purchase and transport of the inoculum can be expensive.

Hulshoff Pol et al. (1983) reported that the addition of crushed granular methanogenic

sludge to digested sewage in a UASB reactor fed with acetate and propionate can give rise

to the development of methanogenic sludge granules with a diameter of 1 to 2 mm.

Two different types of sludges may develop on the same medium depending on the

source of the inoculum. Xu and Tay (2002) used methanol-precultured anaerobic sludge to

inoculate a UASB reactor. This approach accelerated the formation of embryonic granules

in a laboratory-scale UASB reactor. The granulation process reached its postmaturation

stage about 15 to 20 days ahead of the control reactor. Use of methanogen-enriched seed

sludge for UASB inoculation can reduce the time required for start-up. It seems certain

that anaerobic granulation can be expedited simply by manipulating the composition of

seed sludge. This approach can greatly facilitate the start-up of UASB reactors. It is still

not entirely clear as to which species in seed sludge contribute the most to anaerobic

granulation.

3.2.5. Addition of polymer and cations

Synthetic and natural polymers have been widely used in coagulation/flocculation

processes. These polymers are known to promote particle agglomeration and have been

used to enhance the formation of anaerobic granules. El-Mamouni et al. (1998) found that

the supplementation with the polymer chitosan (a polymer that is similar in structure to

polysaccharides) significantly enhanced the formation of anaerobic granules in the UASB-

like reactors. For example, the granulation rate in the chitosan-containing reactor was 2.5-

fold higher than that in the control reactor without the polymer supplementation. The

polymer enhanced granules had about the same specific activity of methane production as

the granules formed without the polymer. Polymeric chains enhance flocculation by

bridging microbial cells. Such initial microbial nuclei are the first step in microbial

granulation. In cationic polymer-assisted anaerobic granulation processes, it has been

observed that the start-up period required for the development of granular sludge blanket

can be shortened significantly compared to when no polymers are used (Uyanik et al.,

2002). Two mechanisms appear to be involved in polymer enhancement of anaerobic

granulation. The addition of polymers to anaerobic systems likely changes the surface

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properties of bacteria to promote association of individual cells. Polymer may also form a

relatively solid and stable three-dimensional matrix within which bacteria multiply and

daughter cells are then confined. The polymer additives appear to play a similar role as do

the naturally secreted extracellular polymeric substances (EPS) in aggregating anaerobic

sludge.

Evidence shows that the presence of divalent and trivalent cations ions, such as Ca2 +,

Mg2 +, Fe2 +, and Fe3 +, helps bind negatively charged cells together to form microbial

nuclei that promote further granulation (Mahoney et al., 1987; Schmidt and Ahring, 1993;

Teo et al., 2000; Yu et al., 2001). De Zeeuw (1984) reported that the rate of sludge

granulation was significantly enhanced at a calcium concentration of 100 mg/l in the

wastewater. Similarly, Mahoney et al. (1987) observed that granule formation was

stimulated by the presence of calcium in a concentration range of 100 to 200 mg/l.

Research by Teo et al. (2000) showed that an increase in Ca2 + concentration from 0 to 80

mg/l substantially improved the strength of anaerobic granules, as indicated by a 60%

decrease in turbidity. The role of Ca2 + in anaerobic granulation processes is still uncertain.

Calcium concentrations exceeding about 500 mg/l (Guiot et al., 1988; Thiele et al., 1990;

Yu et al., 2001) are detrimental to granulation. At high calcium concentrations, problems

such as the precipitation of calcium on the surface of granules and accumulation of

calcium inside anaerobic granules with consequent reduced microbial activity have been

reported (Yu et al., 2001).

3.2.6. Temperature

As a core microbial component of anaerobic granules, methanogenic bacteria grow

slowly in wastewater and their generation times range from 3 days at 35 jC to as high as

50 days at 10 jC (Bitton, 1999). When the reactor temperature is below 30 jC, the activityof methanogens is seriously reduced. This is the main reason why mesophilic UASB

reactors must be operated at a temperature of 30 to 35 jC for successful functioning. In

addition, sludge washout and deterioration of COD removal efficiency have been reported

in UASB reactors when temperature is increased in steps from 37 to 55 jC (Fang and Lau,

1996). Lepisto and Rintala (1999) further reported that effluent quality from a UASB

reactor operated at 70 jC was lower than that from reactors operated at 35 jC and 55 jC.High temperatures are known to encourage the growth of suspended biosolids; however,

extremely high temperatures inhibit bacterial growth. Extreme thermophilic UASB

reactors (i.e., temperature above 55 jC) seem not to be practicable because of the

additional energy that is required to maintain the high temperature and the relatively poor

effluent quality. A high-temperature operation is also difficult to control.

Recently, attention has been given to the impact of low temperature on the performance

of anaerobic granular sludge reactors (Angenent et al., 2001; Lettinga et al., 2001; Lew et

al., 2003; Singh and Viraraghavan, 2003). Singh and Viraraghavan (2003) showed that

COD removal efficiency can be as high as 70 to 90% in a UASB reactor operated at 11 jCwith a hydraulic retention time of 6 h. Similarly, the expanded granular sludge bed (EGSB)

reactors have been shown to be practicable systems for anaerobic treatment of mainly

soluble and preacidified wastewaters at temperatures of 5 to 10 jC (Lettinga et al., 2001).

In addition, anaerobic migrating blanket reactors (AMBRs) have also been successfully

applied to treat low-strength wastewaters at low temperatures (Angenent et al., 2001).

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 553

Clearly, therefore, anaerobic granular sludge systems are suitable for the treatment of

municipal wastewater at low and moderate temperatures.

3.2.7. pH

Based on the sequence of anaerobic reaction, microbial species involved can be roughly

divided into the following three categories: (a) bacteria responsible for hydrolysis; (b)

acid-producing bacteria; and (c) methane-producing bacteria. In general, the acid-produc-

ing bacteria tolerate a low pH and have an optimal pH of 5.0 to 6.0; however, most

methane-producing bacteria can function optimally in a very narrow pH range of 6.7 to 7.4

(Bitton, 1999). This explains why pH is more inhibitory to methane-producing bacteria

than to acidogenic bacteria in UASB reactors. Once the reactor pH falls outside 6.0 to 8.0,

the activity of methane-producing bacteria is reduced to a low level and this decline in

activity in turn poses serious operational problem that can lead to the failure of the reactor.

Under normal operation conditions, the pH reduction caused by acid-producing bacteria

can be buffered by bicarbonate produced by the methane-producing bacteria.

3.3. Characteristics of anaerobic granules

3.3.1. Microstructure

Based on the microscopic observations, a multilayer model for anaerobic granulation

was initially proposed by MacLeod et al. (1990) and Guiot et al. (1992). According to this

model, the microbiological composition of granules is different in each layer. The inner

layer mainly consists of methanogens that may act as nucleation centers that are necessary

for the initiation of granule development. H2-producing and H2-utilizing bacteria are

dominant species in the middle layer and a mix of species, including rods, cocci, and

filamentous bacteria, predominates in the outermost layer. The conversion of a target

organic compound to methane depends on the spatial organization of the methanogens and

other species in UASB granules. The layered structure of UASB granules is further

evidenced by immunological and histologic methods (Ahring et al., 1993; Lens et al.,

1995); dynamic models (Arcand et al., 1994); studies with microelectrodes (Santegoeds et

al., 1999); and fluorescence in situ hybridization using 16S rRNA-targeted oligonucleo-

tides (Sekiguchi et al., 1999; Tagawa et al., 2000).

A distinct layered structure has been found also in the methanogenic–sulfidogenic

aggregates, with sulfate-reducing bacteria in the outer 50 to 100 Am and methanogens in the

inner part (Santegoeds et al., 1999). Unlike the initial multilayer model proposed by

MacLeod et al. (1990), recent research shows that UASB granules have large dark

nonstaining centers in which neither archaeal nor bacterial signals are observed (Rocheleau

et al., 1999). These nonstaining centers may be formed as a result of the accumulation of

metabolically inactive, decaying biomass and inorganic material (Sekiguchi et al., 1999).

Granules with a homogeneous nonlayered structure have also been reported (Groten-

huis et al., 1991; Fang et al., 1995). Filamentous microorganisms were predominant

throughout such granules. Some researchers have argued that layered and nonlayered

microstructures can be developed with carbohydrates and proteins as substrates, respec-

tively (Fang et al., 1995). This phenomena is said to originate in the different initial steps

involved in the degradation of carbohydrates and proteins. The carbohydrate degradation

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563554

to small molecules is faster than the subsequent degradation of the intermediates. In

contrast, the initial step in the protein degradation is a rate-limiting step. Furthermore,

different types of granules can also form on a given substrate (Daffonchio et al., 1995). It

has been suggested that the granular microstructure is dependent on the degradation

kinetics of the substrate. In other words, different dominating catabolic pathways may give

rise to different granules (Daffonchio et al., 1995; Schmidt and Ahring, 1996).

3.3.2. Methanogenic activity

Methane-producing bacteria are one of the major species in anaerobic granules. In most

research on anaerobic granules, the activity of methane-producing bacteria has been used

to quantify the metabolic activity of the granules. In general, the methanogenic activity can

be defined as the methane production per unit biomass per unit time or the methane

production per unit reactor volume per unit time. The methanogenic activity can be used to

evaluate the performance of a reactor and as an indicator of toxic or inhibitory effects on

anaerobic granules.

3.3.3. Physicochemical properties

Research suggests that cell surface hydrophobicity plays a crucial role in both aerobic

and anaerobic granulation (Tay et al., 2000; Teo et al., 2000; Liu et al., 2003b). Increasing

cell surface hydrophobicity causes a decrease in the excess Gibbs energy of the surface

and further facilitates cell-to-cell interactions leading to a stable structure. Some environ-

mental factors are known to influence cell surface hydrophobicity. These factors include

starvation, oxygen level, selection pressure, and the ionic strength of the medium (Rouxhet

and Mozes, 1990; Castellanos et al., 2000; Liu et al., 2003b; Qin et al., 2004).

Biosolids washed out from the UASB reactors have been found to be more hydrophilic

than the sludge retained in the reactor (Mahoney et al., 1987). This seems to indicate that

in the presence of a selection pressure such as a high liquid upflow velocity, micro-

organisms having a high surface hydrophobicity can be self-immobilized to form denser

aggregates that remain in the reactor. Anaerobic granules tend to become weaker as the

surface negative charge increases (Quarmby and Forster, 1995). In general, the surface of

microorganisms is negatively charged at physiological pH values. The maximum cell

immobilization strength was observed at the isoelectric point of the cells (Rouxhet and

Mozes, 1990; Liu, 1995). Therefore, it can be concluded that the anaerobic granulation is

closely correlated with cell surface properties.

The size of anaerobic granules has a dual effect on the performance of a UASB system.

If the granule is too small, it is likely to be easily washed out of the reactor and this leads to

operational instability. On other hand, a large size of granules reduces mass transfer within

it. The large anaerobic granules in UASB reactors are usually associated with fluffy

bacteria. These large granules can be easily lost with the effluent because of their low

density. The size and density of anaerobic granules depend on many factors including

hydrodynamic conditions, COD loading rate, and the microbial species involved. In

industrial practice, a narrow size distribution of granules is preferred and medium-sized

granules with a diameter of 1.0 to 2.0 mm seem the most attractive. The density of

anaerobic granules indicates their compactness. In UASB reactors, anaerobic granules can

grow up to 2 to 5 mm or larger and can range in specific gravity from 1.033 to 1.065

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 555

(Pereboom and Vereijken, 1994; Tay and Yan, 1996). The relatively high specific gravity

of individual anaerobic granules allows them to settle rapidly. This permits good

separation of solids and liquid in a compact separator.

The strength of anaerobic granules strongly influences the stability of granules. A high

strength means a more compact and stable structure of the granules. Quarmby and Forster

(1995) found that high COD loading rates result in low-strength anaerobic granules. This

is expected. The strength of anaerobic granules depends on many factors including the

microbial diversity, organics loading rate, the feed, exopolysaccharide production, and

hydrodynamic shear forces. High-strength anaerobic granules are desirable in industrial

applications.

3.4. Anaerobic granulation in other types of reactors

Almost all research on anaerobic granulation has been carried out in UASB reactors.

The feasibility of other types of bioreactors in the development of anaerobic granular

sludge has not been sufficiently demonstrated. The reasons for this are not clear. This

section discusses some of the work done on anaerobic granules-based bioreactors other

than UASBs.

3.4.1. Anaerobic continuous stirred tank reactor

Anaerobic granules have been successfully produced in continuously stirred tank

reactors (CSTR), but they disappeared within 3 weeks when the reactor was incubated

statically instead of being agitated (Vanderhaegen et al., 1992). This seems to indicate that

hydrodynamic shear forces play a crucial role in maintaining the integrity of granular

sludge, as has been confirmed for aerobic granulation. Thus, anaerobic granulation may

not depend on the type of reactor, but on how it is operated. Of course, the hydrodynamic

flow patterns in UASB and CSTR are different and the question of how the nature of flow

affects the development of anaerobic granules remains to be answered definitively.

3.4.2. Internal circulation reactor

A major problem of UASB reactors is the washout of seed sludge during the start-up

phase. A new generation of more advanced anaerobic reactor systems has been developed

recently in attempts to overcome the washout-related operational problems of the

conventional UASB reactors. One advanced design is the ‘‘internal circulation’’ (IC)

reactor. This is characterized by biogas separation in two stages within a reactor that has a

high height-to-diameter ratio and a gas-driven internal circulation of the effluent

(Pereboom and Vereijken, 1994). The IC reactor consists of two interconnected UASB

compartments on top of each other. Most of the biogas is produced in the first stage that is

located in the bottom part of the IC reactor. The gas is trapped in gas hoods and then rises

through the riser zone to a gas–liquid separator placed on top of the reactor. The biogas

production thus drives an internal circulation flow through the airlift action (Chisti, 1998).

This results in excellent mixing in the bottom zone of the reactor. In the second stage on

top of the reactor, the biomass retention and effluent polishing take place.

Its design makes the IC reactor more capable of handling high upflow liquid and gas

velocities to allow the treatment of low-strength effluents at short hydraulic retention

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563556

times. In addition, the treatment of high-strength effluents at very high volumetric loading

rates is feasible. Furthermore, the high turbulence and sufficient mixing characteristics of

the IC reactors reduce susceptibility to clogging. In general, the performance of IC reactors

is comparable with or better than that of UASBs for high-strength industrial wastewater

treatment. The IC reactor has been successfully used to commercially treat a wide variety

of wastewaters (Habets et al., 1997; Driessen and Yspeert, 1999).

3.4.3. Expanded granular sludge bed (EGSB) reactor

Anaerobic expanded and fluidized bed reactors have been developed for treating

wastewaters. These reactors rely on dense carrier particles (e.g., sand, activated carbon)

that support a microbial biofilm for degrading the waste. Expanded and fluidized bed

reactors are anaerobic biofilm systems and not granular sludge systems. Zoutberg and de

Been (1997) devised a new type of anaerobic reactor, the expanded granular sludge bed

(EGSB) reactor. The most important feature of the EGSB reactor is the sludge granulation

without carrier materials. Granular sludge could grow and be maintained under high liquid

and gas velocities in the EGSB reactor. The EGSB is suitable for treating both industrial

and municipal wastewaters (Zoutberg and de Been, 1997; Lettinga et al., 2001). However,

the maintenance of expanded sludge bed in EGSB is relatively difficult because of an

absence of solid carriers.

3.4.4. Anaerobic sequencing batch reactor (ASBR)

The major differences between ASBR and UASB are the following: (1) a feed

distribution system is not required in ASBR; (2) there is no three-phase separator in the

ASBR; (3) an upflow hydraulic pattern is absent in ASBR; and (4) the ASBR is operated

in discontinuous mode. Wirtz and Dague (1996) successfully cultivated a granular sludge

blanket in an ASBR. Angenent and Sung (2001) found that the mixed liquor volatile

suspended solids retention in the ASBR was 2.5-fold higher than that in the UASB.

Furthermore, the performance in terms of COD removal in the ASBR was comparable to

that of the UASB. Evidence suggests that at low organic loading rates, the performances of

continuous UASB and anaerobic SBR are quite similar; however, continuous UASB

reactors perform better than the anaerobic SBRs at high organic loading rates (Kennedy

and Lentz, 2000).

3.4.5. Anaerobic migrating blanket reactor

Angenent and Sung (2001) reported a novel anaerobic wastewater treatment system, the

anaerobic migrating blanket reactor (AMBR). This is a continuously fed, compartmen-

talized reactor that does not require an elaborate gas–solid separator and systems for feed

distribution. Anaerobic granules developed in the AMBR tended to be darker in color,

smaller, and denser than granules formed in a UASB reactor operated under similar

conditions. The AMBR has some advantages over the UASB reactor. For example, the

AMBR has a low biomass migration rate, less chance of short-circuiting, efficient removal

of poorly biodegradable compounds, and can be operated in step feed mode for high-

strength wastewaters during shock loads. However, the internal structure of AMBR is

more complex than for the UASB. For example, the AMBR requires multipoint

mechanical mixing to improve feed distribution and prevent clogging by sludge.

Page 25: State of the art of biogranulation technology for wastewater treatment

3.5. Industrial application of anaerobic granulation technology

Kassam et al. (2003) analyzed global trends in the industrial use of anaerobic

wastewater treatment systems. Their data showed exponential growth in the use of

industrial anaerobic wastewater systems worldwide up to the mid-1990s. After 1994,

the number of annual installations declined, but it has remained relatively constant over the

last 3 years. As compared to conventional biological processes, the anaerobic granules-

based biosystems have the benefits of: (1) being simple in construction and operation; (2)

requiring no power from external grid; (3) being compact; (4) generating a low amount of

biological sludge; (5) having a high treatment efficiency; (6) being low in capital and

operating costs; (7) requiring no oxygen; and (8) generating methane fuel. The UASB

technology has been successfully applied to treat industrial wastewater from pulp/paper

industry, the food industry, breweries, distilleries, and the chemical industry. However, an

analysis of the market trends shows that the traditional UASB systems are being phased

out in favor of the high-capacity and high-rate systems such as the EGSB and IC (Lettinga

et al., 2001; Kassam et al., 2003).

Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563 557

4. Future work

This work reviewed the state of the art of biogranulation technology developed for

wastewater treatment. Although extensive work has been done in this area, future research

needs to look into some of the following aspects:

(1) In view of extremely long start-up period of anaerobic granular sludge reactors,

strategies need to be examined for accelerating anaerobic granulation.

(2) Apparently, the selection pressure is a main driving force of aerobic granulation (Qin et

al., 2004), but this needs to be elucidated conclusively.

(3) The question of whether the formation of anaerobic, aerobic granules and biofilms are

subject to the similar mechanisms needs to be addressed.

(4) Aerobic granulation has been observed only the SBRs. The feasibility of attaining

aerobic granulation in continuous culture systems needs to be investigated.

(5) Compared to anaerobic granules, aerobic granules have relatively low stability because

of their fast growth rate. Liu et al. (2004) showed that selecting slow-growing bacteria

in aerobic granules improved the stability of the granules. It is desirable to develop a

practical strategy for improving the stability of aerobic granules by manipulating

operational conditions or through selecting for slow-growth bacteria.

(6) Biogranule-associated bacteria live in a confined space. One advantage of the struc-

tured microbial granules is the ability to acquire transmissible, genetic elements at

accelerated rates. Rapid evolution by horizontal transfer of genetic material has been

observed in biofilms (Watnick and Kolter, 2000). Such transfer likely occurs also in

granules, but this needs to be examined.

(7) It would be interesting to look at the feasibility of transplanting engineered species into

microbial granules to tailor microbial granules for treating specific types of

wastewaters.

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Y. Liu, J.-H. Tay / Biotechnology Advances 22 (2004) 533–563558

Acknowledgements

M.A. Nay, S.F. Yang, and Q.S. Liu are thanked for assisting with the collection of

materials for this work.

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