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ENVIRONMENTMonitoring the Bacteriological Quality of

Seawater by Flow Cytometry: Development of aRapid Method for the Control of Ballast Water

Volodymyr Ivanov (cvivanov@ntu.edu.sg)Tay Joo Hwa (cjhtay@ntu.edu.sg)

Eva Joachimsthal (ceva@ntu.edu.sg)Stephen Tay Tiong Lee (ctltay@ntu.edu.sg)

Introduction

Guidelines concerning the threat posed by invasive pathogensdue to uncontrolled discharge of ballast water and sedimentfrom ships are currently in place. Countries such as USA,Australia, Canada, New Zealand, have already taken measuresto prevent the introduction of invasive species from ballastwater. The International Maritime Organization (IMO) willsoon be adopting international regulations on ballast water(Alam, 2000). The conventional method of monitoring themicrobiological quality of ballast water is to determine thenumber of indicator microorganisms using cultivationtechniques or other conventional methods, which have to beperformed in a specialized laboratory by a skilled technicianand may take several days to complete.

For practical purpose, new methods for rapid and simple on-board monitoring of bacteriological quality of ballast watermust be developed. These ballast water assays must be ableto provide rapid results (in less than 8 hours), be easilyperformed by an operator with little or no microbiologicaltraining, and give an accurate enumeration of harmfulmicroorganisms. Molecular biological methods based on theidentification of sequences of specific nucleic acids or proteinsare faster and often simpler than conventional methods(MacGregor, 1999). Fluorescently labeled oligonucleotideprobes can be used to detect these specific microorganismswith a fluorescence microscope, a fluorescence spectrometer(Tay et al., 2001), or a flow cytometer (Vives-Rego et al.,2000).

Flow cytometry is a useful tool for assessing many parametersof microorganisms including enumeration, viability, andidentification. Practical enumeration of bacterial cells mustnot require the addition of mutagenic nucleic acid stains. Forthis reason, we studied a method for the enumeration ofbacterial cells without staining. The results were compared tospread plating and epifluorescence direct counting.

All our experiments with the monitoring of water quality wereperformed using a laboratory flow cytometer FACSCalibur(Becton Dickinson, USA), but according to commercialinformation, micro-cytometers with disposable cartridges,which are most suitable for bacteriological monitoring of waterquality, will be available soon and may be used for the on-sitemonitoring of ballast water.

The experiments were performed with samples of ballast waterand with pure cultures of Escherichia coli because enumerationof these indicator microorganisms is routine practice forbacteriological monitoring of water and the aquaticenvironment.

Materials and methods

The ballast water samples were taken from ships anchored tothe south and to the west of Singapore. Samples of seawaterwere also taken from beside these ships (Figure 1). Purecultures of Escherichia coli ATCC 53323 and Bacillus cereus

Figure 1. Sampling of ballast water

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ATCC 10876 were grown in nutrient broth (Difco Laboratories,USA). R2A agar (Difco Laboratories, USA) was used toenumerate heterotrophs. Oligotrophic counts were performedusing 1000 times diluted R2A agar with15 g/L Noble Agar (Difco Laboratories, USA).

A standard protocol for staining by DAPI (4’,6-diamidino-2-phenylindole, Merck, Germany) was used where 10 µL ofDAPI (1 mg/mL solution) was incubated in the dark with 1mL of sample for 20 minutes. The sample was filtered througha 0.2 µm Isopore membrane filter (Millipore, Ireland) andthe fluorescence of cells was detected using an Olympus BX60epifluorescence microscope.

The relation of microorganisms isolated on heterotrophic mediato oxygen was determined in semi-solid 0.2% nutrient agar(Difco Laboratories, USA) with 1 mg/L rezazurin (Sigma,USA) as an indicator. The cultures were incubated for 7 daysat 30°C.

Oligonucleotide probes specific for E. coli (Eco16S07C andEco440), enterobacteria (En1432), and eubacteria (Eub338)were used for the enumeration. Whole-cell fluorescent in situhybridizations (FISH) were carried out as per standardprotocols. Fluorescently labeled probe (final concentration 100pmol/µL) was added. Un-labeled probe (final concentration1000 pmol/µL) was added for assessment of non-specificbinding.

Pure cultures of E. coli and B. cereus, or size calibrationbeads (Molecular Probes, USA) were used in flow cytometryby FACSCalibur (Becton Dickinson, USA) to determine agate identifying bacterial cells in the samples. The nucleicacid stain SYTO 9 (Molecular Probes, USA) was used tostain the DNA and RNA of cells.

Results

A control sample involving pure cultures of E. coli and B.cereus showed an acquisition gate for bacterial cells in thesamples (Figure 2). It was found that the size range used forthe enumeration method by forward scatter contained 99.4%cells for each sample.

A comparison of enumeration methods performed on theenvironmental samples is shown in Table 1.

There was a significant decrease in the total number of cellscounted in the ballast sample relative to that in other seawatersamples.

As can be seen in all of the enumeration methods theconcentration of cells in the ballast water was significantlylower (an 85% decrease in prokaryotes, and a 76% decreasein eukaryotes) compared to the seawater samples.

The majority of microorganisms in the seawater samples wereeither obligate aerobes or microaerophilic. For the ballast water,the proportion of aerobic microorganisms was 11%. However,the proportion of facultative anaerobic microorganismsincreased, from 7% in the seawater samples, up to 50%.

Table 1. Concentration of microbial cellsin seawater and ballast water

Concentration of cells in1 mL of sample

Method and ParameterSeawater from Ballastnear the ship Water

DAPI StainingTotal cell number 1.4×106 3.9×105

Spread PlateNumber of heterotrophs 2.2×104 1.3×102

Number of oligotrophs 8.6×104 1.2×103

Total cell number 1.1×105 1.3×103

Flow CytometryNumber of prokaryoticcells 5.1×105 8.8×104

Number of eukaryoticcells 2.9×105 7.8×104

Total cell number 8.0×105 1.7×105

Figure 2. Forward scatter histograms of (A) a control involvingE. coli (peak 1) and B. cereus (peak 2), (B) clean seawater,

(C) seawater from near the ship, and (D) ballast water samples.Regions are marked on the histograms for: (R1) prokaryotes,

and (R2) eukaryotes. The references beads are indicated in (A).

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Pure cultures of E. coli were examined using theoligonucleotide probes. Each probe had approximately 30%of cells bound non-specifically after hybridization. It wasshown that the use of labeled and non-labeled Eco16S07Cprobe is a best choice for the purpose of E. coli detection andquantification in environmental samples by flow cytometry.

This probe was then used for environmental analysis of thecontent of E. coli. Figure 3 shows the fluorescence-forwardscatter plot for FISH of the Eco16S07C oligonucleotide probewith a sample of seawater. The concentration of allmicroorganisms in the sample can be enumerated within aspecific size range.

Table 2 lists the values determined from the FISH analysisperformed using the E. coli and enterobacterial oligonucleotideprobes with an environmental sample as well as theenumeration of total cells in the range 1–4 µm. This analysisshows that for this sample the concentration of E. coli accountsfor approximately 42% of the enteric bacteria present. E. colirepresents approximately 0.3% of the total bacterial population.In comparison to the total number of microorganisms (withinthe size range 1–4 µm) in the environmental sample theconcentration of enterobacteria was low.

Table 2. Enumeration of cells by molecular probes

OrganismsCell

Probe Name identified

numberper 1 mL

Eco16S07C E. coli 1.00×103

En1432 Enterobacteria 2.40×103

Total number All cells in theof cells (1–4 µm) range 1–4 µm 3.42×105

Conclusion

Enumeration by flow cytometry, based on light scatter withreference cell cultures as standards was safer and quicker, andfound to be an efficient alternative to conventional methodsof enumeration for samples of ballast water.

The concentration of cells in the ballast water sample was muchsmaller than that in either of the seawater samples. However,the ballast water sample contained a much higher proportion offacultative anaerobic microorganisms compared to the otherseawater samples. This is of importance as facultative anaerobesmight be more likely than other microbial groups to containpotential pathogens and therefore might be more likely to be anenvironmental and health risk when discharged.

The E. coli-specific probe Eco16S07C was the better choiceout of the probes tested for the identification and enumerationof E. coli cells in the samples of ballast water. Theconcentrations of E. coli in the sample of seawater analyzedwere found to contribute to approximately 0.3% of the totalbacterial population within the size range 1–4 µm. Theconcentration of enterobacteria found in the seawater sampletested was 0.7% compared to the total concentration ofmicroorganisms determined within the size range 1–4 µm.

Figure 3. Fluorescence-forward scatter dot-plots for seawater(A) un-stained, with (B) labeled and un-labeled Eco16S07C, and

(C) stained by SYTO 9. Legend: rejected region (Gate),total (R1) and weak non-specifically (R2) stained cells (1–4 µm),

non-specificallystained cells (R3) <1 µm, and beadsfor counting (B1 and B2).

Acknowledgments

This research was funded by Nanyang TechnologicalUniversity and the Maritime Port Authority of Singapore.

References

[1] Alam, Z. (2000). Recent developments in ballast waterissue at the International Maritime Organization (IMO),Singapore Maritime Journal, pp. 63–65.

[2] MacGregor, B.J. (1999). Molecular approaches to the studyof aquatic microbial communities. Current Opinion inBiotechnology, 10: 220–224.

[3] Tay, S.T.L., Ivanov, V., Kim, I.S., Feng, L., & Tay, J.H.2001 Quantification of ratios of Bacteria and Archaea inmethanogenic microbial community by fluorescence insitu hybridization and fluorescence spectrometry. WorldJournal of Microbiology and Biotechnology, 17, 583–589.

[4] Vives-Rego, J., Lebaron, P., Nebe-von-Caron, G. (2000).Current and future applications of flow cytometry inaquatic microbiology. FEMS Microbiology Reviews, 24:429–448.

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Microstructural Optimization of EngineeringProperties of Microbial Granules Used for

Wastewater TreatmentTay Joo Hwa (cjhtay@ntu.edu.sg)

Volodymyr Ivanov (cvivanov@ntu.edu.sg)Stephen Tay Tiong Lee (ctltay@ntu.edu.sg)

Introduction

The development of aerobically grown microbial granules forbiological wastewater treatment is a research area that is attractingincreasing interest. These microbial aggregates have good settlingability and are capable of treating high strength wastewater (Beunet al., 2002; Tay et al., 2001a, b; Etterer and Wilderer, 2001).However, dense microbial aggregates may encounter problemsassociated with diffusion limitation, such as slow diffusion of thenutrients into and metabolites out of the granules. When theenvironmental conditions within the granule become unfavourable,cell death may occur in zones that do not receive sufficient nutritionor that contain inhibitory metabolites. A little is known (Tay et al.,2002) about the existence of these zones and about how thesezones are affected as a result of changes in granule diameter. Theextent of the biomass and the porosity profiles within the granuleinterior were visualized and quantified using specific fluorochromestogether with confocal laser scanning microscopy (CLSM). Thegoal of this study was to determine the biomass and porosity profilesinside granules of different diameters and to develop a simpleapproach to determine a range of granule diameters appropriate forapplication in biological wastewater treatment.

Materials and methods

The microbial granules were cultivated in an aerobic granularsludge reactor using a mineral medium containing acetate, amixture of ethanol with acetate, or glucose only as the carbonsource (Tay et al., 2001a, b). To detect microbial biomass, granulesamples were incubated with 2 µM (final concentration) of thepermeable nucleic acid stain SYTO9 (Molecular Probes, USA)for 1 hr in the dark at room temperature. To determine the porosityprofiles, granule samples were added to a suspension of 0.1 µmmicrospheres from TetraSpec Fluorescent Microsphere Standards®

(Molecular Probes, OR, USA). The mixtures were placed in the0.2 ml Eppendorf tubes and incubated for 4 hr in an orbitalshaker at 100 rpm. The final concentration of the microsphereswas 2⋅108 particles⋅ml-1. Granule images and fluorescence profileswere acquired with a Fluoview300 confocal laser scanningmicroscope (CLSM) (Olympus, Japan) as described earlier (Tayet al., 2002). Green, red, and far red fluorescence was excited bya 10 mW argon laser at 488 nm and helium neon laser at 633 nm.

Results

The central core of the large granules likely contained lysed ordead microbial cells (Figure 1; Figure 2b, c).

Small granules with diameters of approximately 0.6 mm consistedentirely of biomass (Figure 2a). Larger granules possessed a sack-like structure with a thick wall of microbial biomass. This wallwas between 0.3 to 0.4 mm thick in the 1.3 mm-diameter granule

Figure 1. CLSM image of granule cross-section.(Green colour shows a region dominated by dead cells).

(Figure 2b) and 0.8 mm thick in the 3.0 mm-diameter granule(Figure 2c).

Cell lysis and cell death occur in granules that are large enoughthat problems associated with diffusion limitation within thegranule interior probably became insurmountable. Although theseproblems can be mitigated by the presence of pores and channelswhich can facilitate the transport of substrates into and wasteproducts out of deep biofilm regions, these pores and channelsmay eventually be plugged by the secondary colonization ofmicroorganisms or by the formation of extracellular polymers.

The microspheres did not adhere to the cell surface. Therefore,their distribution within the granule was not a measure of theadsorption of the beads onto the granule matrix but indicated thepenetration of the beads into the granule interior by passage throughpore and channel structures larger than 0.1 µm in size (Figure 3).

The thickness of the porous layer in the granule positivelycorrelated with the granule diameter. For example, a granule witha diameter of 550 µm had a porous layer with a thickness of 250µm and a granule with a diameter of 1,000 µm had a porouslayer with a thickness of 350 µm. The biomass content in thegranules and the porosity profiles were also observed to drop atthe same depth below the granule surface.

A useful engineering consequence of these measures of biomassand porosity in aerobically grown granules is the determinationof a cut-off or critical granule diameter above which diffusionlimitation would set in. The value of this critical diameter, Dc, isdetermined by assuming that diffusion limitation sets in when thegranule diameter is greater than two times the biomass layer ortwo times the porous layer. Figure 4 shows a plot of granulediameter versus two times the biomass layer or two times theporous layer (coefficient of correlation is 0.99). The equation ofthis correlation is:

2Hl = 0.4Dg + 0.3 mm (1),

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where Hl is the thickness of the porous layer or biomass layerand Dg is the diameter of the granule. The critical dameter, Dc,may be calculated from Equation 1 with the condition that 2Hl

= Dg, which means that whole granules with diameters less thanor equal to Dc consists entirely of a porous matrix with activecells. Dc calculated from Equation 1 for this condition is 0.5 mm.

Biomass- and porosity-based measures of critical diameter canbe used in the development of aerobic granulation systems forwastewater treatment provided the choice of critical diameter issupported by the physiological studies. Conditions in the aerobicgranular sludge reactor can be imposed to enhance wastewatertreatment efficiency by selecting for microbial granules with adiameter smaller than Dc.

Conclusion

1. Aerobically grown microbial granules with diameters rangingfrom 0.45 mm to 3.0 mm contained a porous layer ofmicrobial biomass. The amount of microbial biomass andthe extent of porous zones on a per unit volume basisdecreased with increasing granule diameter. Larger granulespossessed thicker biomass layers.

2. The critical diameter of 0.5 mm for aerobically grownmicrobial granules was calculated on the assumption thatwhole granules should consist entirely of a porous biomassmatrix. This value can be used in the design of aerobicgranulation systems for wastewater treatment.

Figure 2. Biomass profile in the granules of different size.(Arrow points the granule center; dotted line shows

right edge of the granule cross-section).

Figure 3. CLSM image of granule porosity.(Porosity is shown as green fluorescence of 0.1 µm fluorescent

microspheres that penetrated into the granule interior).

Acknowledgements

This research was supported by Nanyang Technological University,Singapore.

References

[1] Beun J.J., van Loosdrecht M.C.M., Heijnen J.J. (2002)Aerobic granulation in a sequencing batch airlift reactor.Water Research 36: 702-712.

[2] Etterer,T. and Wilderer, P.A. (2001) Generation and propertiesof aerobic granular sludge. Water Science and Technology43:19-26.

[3] Tay, J.H., Liu, Q.S., Liu, Y. (2001a) Microscopic observationof aerobic granulation in sequential aerobic sludge reactor.Journal of Applied Microbiology 91: 168-175.

[4] Tay, J.-H., Liu, Q.-S., Liu Y. (2001b) The role of cellularpolysaccharides in the formation and stability of aerobicgranules. Letters in Applied Microbiology 33: 222-227.

[5] Tay, J.-H., Ivanov V., Pan S., Tay S. T.-L. (2002) Specificlayers in aerobically grown microbial granules. Letters inApplied Microbiology, 34 (4): 254-258.

Figure 4. Plot of double the thickness of the biomass layer( � ) or porous layer (�) against granule diameter.

The dotted line depicts the condition in which the whole granuleconsists entirely of a porous biomass matrix (2Hl = Dg).

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Role of Filamentous Microorganisms in theFormation of Glucose-fed Aerobic Granules

A M Maszenan (cmaszenan@ntu.edu.sg)J H Tay (cjhtay@ntu.edu.sg)

B Y P Moy (cypmoy@ntu.edu.sg)S T L Tay (ctltay@ntu.edu.sg)

Introduction

Microbial granulation process is widely used in biologicalwastewater treatment whereby bacterial consortia capable ofremoving biodegradable organic, nitrogen and phosphorus areorganised into highly structured granules. This self-immobilisation of microbes has been extensively applied inthe upflow anaerobic sludge blanket (UASB) reactor, whereanaerobic granules are used to perform anaerobic wastewatertreatment (Lettinga et al. 1980). However, anaerobicgranulation suffers from several disadvantages, including along start-up period of several months due to the slow growthof anaerobic biomass.

On the other hand, aerobic granulation is a recently describedphenomenon in which granules can be fully formed insequential batch reactors within three weeks of start-up (Tayet al. 2001). These aerobic granules also have severaladvantages over conventional activated sludge systems,including a strong and compact microbial structure, goodsettling ability, high biomass retention and ability to accept ahigh organic loading rate (Moy et al. 2002).

The studies that have been documented thus far have focusedlargely on the influence of reactor conditions on the formationof aerobic granules. However, little is known about the typesof microorganisms that reside in these granules. Based onmicroscopy images, Beun et al. (1999) observed theproliferation of filamentous fungal pellets during the initialstage of operation of a sequential batch reactor. These mycelialpellets would eventually expand and lyse as a result of oxygenlimitation. The resulting fragments would then serve as animmobilisation matrix for bacteria to attach and formmicrocolonies that would in turn grow into compact granules.

In this study, the occurrence of filamentous bacteria wasinvestigated using microscopy image analysis. By studyingmicrobial population involved in aerobic granulation, this workcould help better understand the microbiology of aerobicgranules and provide useful information for optimisation ofthe granulation process.

Results

A sequencing batch reactor was used to cultivate the aerobicgranules using glucose as substrate. Granules first appearedin the reactor after 16 days of operation. Filamentous bacteriawere present both in the initial seed flocs as well as in theaerobic granules that were eventually formed. After reactorstart-up, the filament counts increased significantly from 329± 26 µl-1 on day 2 to 406 ± 33 µl-1 on day 6. The sludgevolume index (SVI) value also increased from 211 ml g-1 onday 2 to 331 ml g-1 on day 6. However, the formation of

aerobic granules was accompanied by dramatic decreases inboth filament counts and SVI. Filament counts and SVIdecreased significantly to 91 ± 10 µl-1 and 100 ml g-1,respectively, on day 16, and subsequently stabilized at theselevels.

Microscopy images revealed that the number of pin-point flocsincreased in the reactor from day 2 to day 13. The flocsinitially grew to a size large enough that these flocs werefragmented by the aeration-induced shear into smaller flocs(Figure 1). Filamentous bacteria formed an interconnectingbridge that facilitated the aggregation of flocs into morecompact biological particles. The pin point flocs wereeventually colonized by microorganisms and developed intogood settling granules that could be retained in the reactor.

The filamentous bacteria that were present in the sludgebiomass were identified based on cell morphology. Thesefilamentous bacteria shared the same descriptions as previouslydescribed morphotypes of Sphaerotilus natans, and Types 1701,1851 and 1863 (Jenkins et al. 1993). Sphaerotilus natans andType 1701 were more abundant, while Types 1851 and 1863were present to a lesser extent. Sphaerotilus natans-likebacteria stained gram-negative, contained an empty regionwithin the bacterial sheaths, and were characterized by typicalfalse branching (Figure 2A). Type 1701-like bacteria werecurved rods with clear indentation septa and intertwined withinthe floc structure (Figure 2B). Type 1851-like bacteriadisplayed a characteristic sparse attached growth that extendedout of surface flocs and appeared as bundles of intertwinedfilaments (Figure 2C). Type 1863-like bacteria consisted ofirregularly bent filaments of gram negative cells with noattached growth, occurred in chains and displayed indentationsat the septa (Figure 2D).

Figure 1. Filament bridging of aerobic flocs.Scale bar represents 10 µm.

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Discussion

In the current study, activated sludge from a municipalwastewater treatment plant was used to inoculate a sequencingbatch reactor to produce aerobic granules. After sludge seeding,both the filament count and SVI increased as a result of theoutgrowth of filamentous bacteria. This outgrowth hinderedsettling ability and is a common phenomenon known as sludgebulking in the activated sludge process. Prolonged periods ofanoxia and substrate shortages or overloads are known topromote the growth of filamentous bacteria such asHeliscomenobacter hydrossis and S. natans (Jenkins et al. 1993).These filamentous bacteria may upset sludge settling propertiesas a result of increased steric interactions or increased flocporosity.

Based on established identification schemes for filamentousbacteria, S. natans and Type 1701 were found to be dominantmembers in the glucose-fed flocs and granules. Types 1851and 1863 were less abundant. An excessive abundance ofType 1851 and S. natans is known to result in poor filamentbridging, while Type 1701 has been associated with an openfloc structure, which has a negative effect on sludgecompaction and settling in activated sludge systems (Jenkinset al., 1993). Nevertheless, a well-settled sludge developedand was retained in the reactor from day 14 of the currentstudy, while filamentous and suspended bacterial cells werewashed out. These observations coincided with dramaticdecreases in filament count and SVI to 98 filaments µl-1 and109 ml g-1, respectively, and signalled the onset of granuleformation. Paradoxically, the filamentous bacteria were

present in optimal numbers that they provided aninterconnecting bridge that appeared to promote flocattachment and enhanced the mechanical strength of the flocstructure. One of the early models of aerobic granulationinvolved the growth and subsequent fragmentation offilamentous fungal pellets under stressed conditions (Beunet al. 1999). These fungal fragments functioned as animmobilization matrix for colonization by floc-formingbacteria, and eventually grew out to become granules withgood settling properties that were easily retained in thereactor. The filamentous bacteria in the current study likelyplayed a similarly important role to the filamentous fungi inthe formation of aerobic granules, by facilitating flocaggregation and providing attachment sites for colonizationby other microorganisms.

Although fungi had previously been implicated in the formationof aerobic granules (Beun et al. 1999), the vast diversity inherentin the microbial world suggest that other microorganisms couldparticipate in different modes of granule formation. Selectionfor different microorganisms could be achieved by a judiciouschoice of substrate, seed sludge, and other operating conditions.For these reasons, it was not unexpected to discover in thiscurrent study a different mode of aerobic granule mediated byfilamentous bacteria.

References

[1] Beun J. J., Hendriks A., van Loosdrecht M. C. M.,Morgenroth E., Wilderer P. A. and Heijnen J. J. (1999).

Figure 2. Light microscopy images of aerobic flocs containing Sphaerotilus natans (A), and Types 1701 (B), 1851 (C)and 1863 (D). Scale bars for A, B and D represent 10 µm. Scale bar for C represents 100 µm.

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Aerobic granulation in a sequencing batch reactor. WaterResearch 33, 2283-2290.

[2] Jenkins D., Richard M. G. and Daigger G. T. (1993). Manualon the causes and control of activated sludge bulking andfoaming. Lewis Publishers, Chelsea, Michigan.

[3] Lettinga G., van Velsen A. F. M., Hobma S. M., de ZeeuwW and Klapwijk A. (1980). Use of the Upflow Sludge Blanket(USB) reactor concept for biological wastewater treatment.Biotechnology & Bioengineering 22, 699-734.

Achieving High Organic Loads withGlucose-fed Aerobic Granules

B Y P Moy (cypmoy@ntu.edu.sg)J H Tay (cjhtay@ntu.edu.sg)S K Toh (csktoh@ntu.edu.sg)

Y Liu (cyliu@ntu.edu.sg)S T L Tay (ctltay@ntu.edu.sg)

Introduction

Aerobic granules have several advantages over conventionalactivated sludge systems, including a strong, compact microbialstructure, good settling ability and high biomass retention.Because of their ability to retain biomass, granules can handlehigh organic loading rates (OLRs). While most wastewatertreatment systems have relatively low OLRs (0.5 to 2.0 kgchemical oxygen demand [COD] m-3 d-1 for activated sludgesystems), compact upflow anaerobic sludge blanket reactors havebeen developed to handle high OLRs of up to 40 kg COD m-

3 d-1 under anaerobic conditions. However, little is known aboutthe application of high OLRs to aerobic granules grown insequential batch reactors, and earlier studies employed OLRsbelow 7.5 kg COD m-3 d-1 (Beun et al. 1999; Tay et al. 2001a).The main objective of this study was therefore to investigatethe effect of high OLRs on the physical characteristics of aerobicgranules. This work could contribute to a better understandingof the ability of aerobic granules to handle high-strengthwastewaters.

Results

Aerobic granules appeared in the glucose-fed reactor during thethird week of reactor start up. The OLR was maintained at 6kg COD m-3 d-1 for an additional three weeks. The fluffy outerappearance of the granules at this initial OLR (Fig. 1a) wasreplaced by a smooth irregular morphology when the OLRincreased to 9 kg COD m-3 d-1. The granules maintained thesame irregular morphology but had larger mean diameters whenthe OLR increased to 12 and 15 kg COD m-3 d-1 (Fig. 1b).

Scanning electron microscopy (SEM) images revealed thatglucose-fed granules cultivated at an OLR of 6 kg COD m-3 d-

1 had a hairy appearance with a loose microbial structuredominated by filamentous bacteria. On the other hand, glucose-fed granules cultivated at higher OLRs had a compact andirregular structure characterized by folds, crevices and

Fig. 1 Morphology of glucose-fed granules at OLRs of(a) 6 kg COD m-3 d-1; and (b) 15 kg COD m-3 d-1.

depressions. Settling ability and strength are generally improvedwith increased OLR in aerobic glucose-fed granules. The SVIvalues improved from 106 ml g-1 at 6 kg COD m-3 d-1 to 31 mlg-1 at 15 kg COD m-3 d-1. A higher OLR contributed to a lowerSVI, with denser and more compact granules and better settling.Settling velocities measurements increased from 53 m h-1 at 6kg COD m-3 d-1 to 84 m h-1 at 15 kg COD m-3 d-1, and thestrength of the aerobic granules also showed an improving trendwith higher OLR; the integrity coefficient increased from 93.0% at 6 kg COD m-3 d-1 to 99.0 % at 15 kg COD m-3 d-1.

Discussion

In the current study, glucose-fed granules were able to sustain themaximum OLR of 15 kg COD m-3 d-1 without compromisinggranule integrity. The type of granule microstructure developedinfluenced the ability of the granules to handle high OLRs.Glucose-fed granules showed improving settling and strengthcharacteristics with increasing OLR, and had better SVI andintegrity coefficient values at the highest OLR of 15 kg COD m-

3 d-1 compared to other substrate-fed granules. At higher OLRs,the glucose-fed granules developed an irregular appearancecharacterised by folds, crevices and depressions that allowed forshorter diffusion distances and better penetration of nutrients intothe granule interior. Diffusion was also enhanced by the highersubstrate concentration that existed in the bulk solution at higherOLRs. These factors probably enabled the glucose-fed granules

[4] Moy B.Y.-P., Tay J.-H., Toh S.-K., Liu Y. and Tay S.T.-L.(2002). High organic loading influences the physicalcharacteristics of aerobic sludge granules. Letters in AppliedMicrobiology 34, 407-412.

[5] Tay J.-H., Liu Q.-S. and Liu Y. (2001). Microscopic observationof aerobic granulation in sequential aerobic sludge blanketreactor. Journal of Applied Microbiology 91, 168-175.

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to sustain the high OLR of 15 kg COD m-3 d-1. Diffusion limitationsymptoms were not detected in the glucose-fed granules duringthis investigation. Paradoxically, the initial loose filamentousmatrix resulted in an irregular morphology that enabled glucose-fed granules to sustain significantly higher OLRs before masstransfer became restrictive.

Although the maximum OLR investigated here was 15 kg CODm-3 d-1, it is worth noting that aerobic granules are likely able tohandle higher OLRs. This could be accomplished by adjustingparameters such as superficial gas velocity. A higher superficialgas velocity favors the formation of smaller and more compactgranules where diffusion limitation is less likely to adverselyaffect reactor operation (Tay et al. 2001b). This studydemonstrates that high organic loading rates were possible withaerobic granules and this ability would be helpful in thedevelopment of aerobic granule-based systems for high-strengthwastewaters.

References

[1] Beun, J.J., Hendriks, A., Van Loosdrecht, M.C.M.,Morgenroth, E., Wilderer, P.A. and Heijnen, J.J. (1999)Aerobic granulation in a sequencing batch reactor. WaterResearch 33,2283-2290

[2] Tay, J.H., Liu, Q.S. and Liu, Y. (2001a) The role of cellularpolysaccharides in the formation and stability of aerobicgranules. Letters in Applied Microbiology 33,222-226.

[3] Tay, J.H., Liu, Q.S. and Liu, Y. (2001b) The effects ofshear force on the formation, structure and metabolism ofaerobic granules. Applied Microbiology and Biotechnology57,227-233.

Treatment of Food Processing Wastewater byPolymer-Conditioned UASB Bioreactors

Tay J H (cjhtay@ntu.edu.sg)Show K Y (ckyshow@ntu.edu.sg)

Foong S F (PS7835647B@ntu.edu.sg)Sim K H (PS7609778Z@ntu.eu.sg)

Lam W Y (cwylam@ntu.edu.sg)Lua C H (cchlua@ntu.edu.sg)

Introduction

The Upflow Anaerobic Sludge Blanket (UASB) system is abioreactor cum clarifier provided by a three-phase separator ina single unit. The most important feature of this process is itsself-immobilization of sludge that shows good settling propertiesafter a start-up period of aggregation and granulation. The mostsignificant drawback lies in the long start-up time often lastingseveral months due to the slow growth rate of the anaerobicbacteria and also the loss in activity caused by the initial washout.

Materials and methods

The industrial food-processing wastewater was collected fromthe JTC Industrial Estate in Woodlands that houses more thana hundred food processing companies. Samples were collectedweekly and analysed for Chemical Oxygen Demand (COD)and Suspended Solids (SS) based on the Standard Methods1998.

Four plexiglass UASB reactors (R1, R2, R3 and R4) with aninternal diameter of 100 mm, a height of 680 mm and effectiveworking volume of 4.4 L were used for the study. The experimentalset-up was placed in a temperature control room at 35 ± 1°C andthe feeding tank was kept at 4°C after primary treatment by asedimentation tank (see Figure 1). For seeding, the reactors wereinoculated with 70% polymer-conditioned anaerobic digested

sludge by volume obtained from the Jurong Water ReclamationPlant. The synthetic polymer used in this study was supplied bycourtesy of Ondeo-Nalco Pacific Pte Ltd.

Serving as a control reactor, R1 was not polymer-conditioned.R2, R3 and R4 were inoculated with 1, 2 and 4 mg polymer/g SS respectively. The UASB reactors were loaded in 2 phaseswith hydraulic retention time (HRT) of 30 and 15 h, which isequivalent to organic loading rate (OLR) of 1.5 and 3 gCOD / L .d, respectively.

Results and discussion

The COD and SS fluctuations in the period from 25 March2002 to 21 Sep 2002 are presented in Figure 2. CODconcentrations ranged from 1583 to 3533 mg/L and SS from404 to 1357 mg/L with an average of 877 and 2197 mg/L,respectively.

Table 1. COD removal and time to reach steady-statein R1 to R4 at 1.5 g COD/L .d

Reactor R1 R2 R3 R4

COD removal (%) 77 70 72 75

Time to reach steady-state (days) 57 48 48 53

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40Figure 3 shows SS concentrations of the wastewater after 30minutes settling time in the gravity settler. Results showed ahigh removal efficiency of more than 80%, as well as consistentresidue SS below 400 mg/L, which is the discharge limit asspecified by Ministry of the Environment (ENV).

R2 and R3 reached steady-state after 48 days of operation attaininga removal efficiency of 70 and 72% respectively. R4 took 5 dayslonger in comparison, with 75% COD removal. All of the reactorswere able to reach steady-state in a shorter duration compared tothe control, at a reduced COD removal rate, however, as shownin Table 1. The greatest reduction in time taken was 9 days in R2and R3 dosed with 1 and 2 mg polymer per g SS, which meansthat the start up was favourably accelerated by 16%.

Figure 4 presents the variations in effluent COD concentrationsand the COD removal efficiencies of all four reactors. An organicshock was experienced in the period of day 79 to 81 caused bya sudden 40% increase in wastewater COD concentration. Thisshock was well received by R2 with a further improvement of14% in COD removal efficiency and a simultaneous decreaseof 200mg/L in the effluent COD concentration. On the contrary,R1 and R4 experienced a sharp increase in effluent CODconcentration of 200 to 300 mg/ L. Onset of granulation asdefined by the formation of biological solids having diametersmore than 0.5 mm was observed on day 44 in the control reactorR1 compared to day 37 of the three polymer-conditionedreactors.

Conclusion

The results obtained indicated the positive effects of polymer-conditioned sludge both in start-up and granulation acceleration.It suggested that the optimum polymer dosage could lay between1 to 2 mg polymer/g SS as R2 and R3 was able to shorten bothstart-up time and accelerated granulation by 16%. Furtherobservations of the subsequent load stages are still beinginvestigated.

Figure 1. Schematic experimental set-up of 1 unitof UASB reactor

Figure 2. COD and SS variations of food processing wastewater

Figure 3. Settled SS concentration and removal efficiency

Figure 4. COD removal efficiencies of R1 to R4

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Introduction

The removal of nitrogen from wastewater has greatersignificance in recent years as a means of protecting andpreserving the environment. Many modified biologicalprocesses were developed and implemented in order to meetthe increasingly stringent environmental regulations. Due tothe sensitivity of nitrifying bacteria to environmental factorsas well as their lower growth rates, it is difficult to obtainand maintain sufficient nitrifying bacteria in conventionalwastewater treatment plants.

Previous research showed that hydraulic selection pressurein terms of upflow flow velocity and hydraulic retentiontime (HRT) in an UASB reactor, and cycle time or HRT insequencing batch reactor (SBR) would play a crucial role inanaerobic granulation process. It should be pointed out thatthere is no information regarding the role of hydraulicselection pressure in the development of nitrifying granules.Therefore, the main objectives of this work were to showthe feasibility of nitrifying granulation in SBR and further toinvestigate the effects of hydraulic selection pressure in termsof the SBR cycle time on the formation, structure andmetabolic behaviours of nitrifying granules.

Materials and methods

Four columns (120 cm height and 5 cm in diameter) with aworking volume of 2.4 litres were used as sequencing batchreactors, each with the same geometrical configuration.Reactors one to four (R1 to R4) were supplied with an airflow of 3.5 l/min, which resulted in a superficial upflow airvelocity of 3.0 cm/s. R1 to R4 were operated at a cycle timeof 24 hours, 12 hours, 6 hours and 3 hours, respectively,while other operation conditions were kept as the same forall four reactors (4 minutes of influent filling, 30 minutes ofsettling and 4 minutes of effluent withdrawal). Effluent wasdischarged at the middle port of the column. The influentammonium concentration was fixed at 80 mg N/l for all fourreactors. The experiments were conducted in a temperaturecontrol room of 25°C. Ammonium, nitrite and nitrateconcentrations were measured by a flow injection analyser(QuikChem Method 10-107-06-1-I, Lachat Instruments, Inc.).The size of granular sludge was obtained by using laserparticle size analysis system (Malvern Mastersizer Series2600) or image analyser (Quantimnet 500 Image Analyzer,Lecia Cambridge Instruments).

Nitrifying Granulation: An InnovativeBiotechnology for Wastewater Nitrification

Tay Joo-Hwa (cjhtay@ntu.edu.sg)Yang Shu-Fang (P149512399@ntu.edu.sg)

Liu Yu (cyliu@ntu.edu.sg)

Results

A unique feature of the SBR system is the cycle operation,and SBR cycle time is interrelated to the HRT. In this study,sludge wasting was conducted through the cycling effluentwithdrawal from the system, thus the SBR cycle time wasinversely related to the daily frequency of sludge discharge,i.e., represented the hydraulic selection strength on microbialcommunity (Shizas and Bagley 2001). In the experiments,R1 to R4 were operated at different cycle times, rangingfrom 24 hours to 3 hours. A complete washout of sludgeblanket occurred and led to a failure of nitrifying granulationin R4 run at the shortest cycle time of 3 hours, while onlytypical bioflocs were cultivated in R1 operated at the longestcycle time of 24 hours. However, after two weeks’ operation,tiny nitrifying granules with a respective mean diameter of0.22 and 0.24 mm appeared in R2 and R3, which wereoperated at a cycle time of 12 and 6 hours. In contrast withthe seed sludge that had a very loose and irregular structurewith a mean size of 0.08 mm and SVI value of 265 ml/g, thenitrifying granules developed in R2 and R3 showed a compactand round-shaped structure with a clear outer shape. Theevolution of sludge size in R1, R2 and R3 was further shownin Fig. 1, indicating that the SBR cycle time had a profoundinfluence on the formation of nitrifying granules. After theformation of nitrifying granules in R2 and R3, the SVIdropped to as low as 30 to 50 ml/g.

Figs. 2a to 2c show nitrification profiles in R1 to R3. Thesalient points of the data are: (i) a rapid fall in the ammonium-N concentration; (ii) almost all input ammonium-N convertedto nitrite and nitrate; (iii) no a lag nitrate production withrespect to nitrite formation. Nitrification is mainly completedby two kinds of bacteria, Nitrosomonas responsible for nitriteformation, while Nitrobacter for converting nitrite to nitrate.

Figure 1. Sludge size versus operation time in R1 to R3

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In fact, under the conditions without inhibition, at leasttwo factors can influence the nitrification profiles, theywere the relative specific growth rates of Nitrosomonas toNitrobacter in microbial community, and the relative massratio between Nitrosomonas and Nitrobacter in the system.It appears from Figs. 2a to 2c that the nitrifying granulesdeveloped in R2 and R3 had a higher activity in terms ofammonium oxidation than the bioflocs in R1.

Conclusion

The nature of SBR is cycle operation, thus SBR cycle timecan serve as a main hydraulic selection pressure imposedon microbial community in the system. No nitrifyinggranulation was observed in the SBR operated at the longestcycle time of 24 hours due to a very weak hydraulicselection pressure, while the washout of nitrifying sludgewas found in the SBR run at the shortest cycle time of 3hours, and led to a failure of nitrifying granulation. Excellentnitrifying granules with a mean diameter of 0.25 mm andspecific gravity of 1.014 were developed in the SBRoperated at a cycle time of 6 hours and 12 hours,respectively. This work shows that nitrifying granules canbe developed at a proper hydraulic selection pressure interm of SBR cycle time. Nitrifying granulation indeed wasa novel biotechnology that has great potentials forwastewater nitrification.Figure 2. Nitrification profiles. a: R1; b: R2; c: R3

Coagulation of Humic Acidsby Ferric Chloride in Seawater

Niu Aixiang (p148909824@ntu.edu.sg)Francis Wilson (cfwilson@ntu.edu.sg)

Introduction

Development in membrane technologies in the last fewdecades has made it economic to employ reverse osmosis(RO) to produce freshwater from seawater. However,membrane fouling has been an important limiting problemin the process. Extensive pretreatment is usually required toremove the potential membrane foulants from feed seawater.

Traditional coagulation has been used in seawaterpretreatment in the last few decades (Duan et al., 2002). Itwould be a new trend to combine a membrane module, suchas nanofiltration (NF) and ultrafiltration (NF), withcoagulation as the pretreatment (Kurihara et al., 2001). Adetailed study of seawater coagulation coupled withmembrane filtration is therefore important for the applicationof coagulation in seawater pretreatment. Because of theirimportance in contributing to membrane fouling, humic acids(HA) were chosen as the target of coagulation in this research.The goal of this study was to investigate coagulation behavior

in seawater and establish a coagulation diagram to predictgeneral pH and iron dosage domains for HA removal.

Materials and methods

Materials and reagents

Raw seawater was collected from Raffles Marina in WesternSingapore and filtered through a 0.2-µm filter. The averageHA concentration of filtered seawater was 1.3 mg/L. Thetest sample was filtered seawater spiked with HA.

Ferric chloride (FeCl3·6H

2O) was used as the coagulant. HCl

and NaOH were used for pH adjustment, and NaHCO3 was

used for alkalinity addition.

Experiments

The experiments were carried out using the jar test technique.The procedure involved rapid mixing for 2 minutes at 250

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rpm, followed by slow stirring for 20 minutes at 50 rpm andquiescent settling for 2 hours.

At the end of the flash mixing, a 50-mL sample was collectedfrom the supernatant and analyzed for zeta potential. Afterthe test sample had settled for 30 minutes, a 50-mL samplewas collected from the supernatant and analyzed for pH. Atthe end of the settling period, a 250-mL sample was collectedfrom the supernatant, filtered through a 0.45-µm filter andanalyzed for UV absorbance. Absorbance at 254 nm wasinterpreted as HA concentration using a calibration curve.Usage of the 0.45-µm filtration was to simulate microfiltration(MF).

Equipment

The equipment used for the jar test was a scientific flocculator(Stuart, SWI). It consists of six stainless steel paddles, eachhaving a height of 2.5 cm and a width of 6.4 cm. Theapparatus for zeta potential measurement was a zetasizer(Malvern 3000). The apparatus for absorbance measurementwas a UV/Vis spectrometer (Perkin Elmer, Lambda Bio 20).

Results and discussion

Jar tests were performed over a wide range of iron dosage(16 - 2500 µmol/L) and pH (4 - 10). The pH of theexperiments was maintained within ± 0.1 of the target pH.Duplicate experiments were performed and the resultspresented in this article are the mean values. The standarddeviations for the duplicate experiments are less than 5 % ofthe HA removal efficiency and 2 mV of the zeta potential.

Figure 1 shows the variation of HA removal efficiency atiron dosage 0 - 500 µmol/L over the pH range of 4 - 9. Forthe same pH conditions, the HA removal increased withincreasing iron dosage and no restabilization was observed.For all the pH values examined, HA removal of around 40% was achieved at zero iron dosage. Compression of theelectrical double layer, contributed by the cationic ions, suchas Ca2+, Mg2+ and Na+, present in seawater at very highconcentrations, might be the dominant coagulationmechanism under this condition. For each pH, an iron dosageof around 125 µmol/L resulted in a HA removal equivalentto 90 % of the respective highest HA removal achieved, andHA removal increased only slightly with further increase iniron dosage. So, the iron dosage of 125 µmol/L might beconsidered a critical value. At iron dosages lower than 125µmol/L, the HA removal was sensitive to changes in irondosage; at iron dosages higher than 125 µmol/L, the HAremoval was almost stable. The dominant mechanism ispostulated to be sweep coagulation at iron dosages higherthan 100 µmol/L. The iron coagulant performed best at pH5 in terms of both the highest HA removal achieved and theHA removal achieved for the same iron dosage. HA removalat pH 4 was also satisfactory, the result being close to thatat pH 5. Both the highest HA removal achieved and the HAremoval achieved for the same iron dosage decreased withincreasing pH from 5 to 9. Overall, the HA removal variedconsiderably with the pH at iron dosages below 150 µmol/L while less significant influence of the pH on HA removalwas observed in the iron dosage range of 150 - 500 µmol/L.

Figure 1. HA removal versus iron dosage

Figure 2. Zeta potential versus iron dosage

Figure 3. Coagulation diagram for HA removal

Figure 2 shows the variation of zeta potential over the irondosage range of 0 - 500 µmol/L. For the same pH conditions,the net charge showed a decreasing trend with increasingiron dosage. No direct relationship between the variation ofzeta potential and that of HA removal with pH was found.Stronger charge neutralization always occurred at pH 10than at other pH values. At pH 4 - 9, the zeta potentialvaried markedly with the pH in the iron dosage range of 150- 500 µmol/L while much less significant influence of thepH on charge neutralization was observed at iron dosageslower than 150 µmol/L. As the iron dosage range where thepH influences HA removal significantly was not the same asthat in which the pH influences the zeta potential significantly,the zeta potential could not serve as an accurate indicationof particle destabilization and aggregation.

A coagulation diagram for HA removal was defined basedon the experimental results (Figure 3). An iron dosage of

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120 µmol/L at pH 5 resulted in HA removal of 95 %. Aniron dosage of 500 and 300 µmol/L at pH 4 and 6,respectively, also resulted in HA removal of 95 %. HAremoval of 95 % was not achieved at pH 7 - 9 over thewhole iron dosage range examined (0 - 2500 µmol/L).

Conclusion

1. A coagulation diagram for HA removal was defined.The diagram may be used to predict the domain wheresatisfactory HA removal will be achieved for ferricchloride coagulation in seawater.

2. For the same pH conditions, the degree of chargeneutralization increased with increasing iron dosage andno restabilization was observed.

3. HA removal by coagulation was best at pH 5.

4. The zeta potential was not an accurate indication ofparticle destabilization and coagulation.

References

[1] Duan J., Wang J., Graham N and Wilson F., (in press,2002). Coagulation of humic acids by aluminiumsulphate in saline water conditions. Desalination.

[2] Kurihara M., Fusaoka Y. and Yamamura H., (2001).Recent progress on membrane technology and challengesfor 21st century. The IDA International Conference onDesalination, Singapore.

Intensive Aerobic Bioconversion ofSewage Sludge and Food Waste into Biofertilizer

Olena Stabnikova (cstab@ntu.edu.sg)Chen Xiaoge (CXGChen@ntu.edu.sg)

Tay Joo Hwa (cjhtay@ntu.edu.sg)Tay Stephen Tiong Lee (ctltay@ntu.edu.sg)

Wang Jing-Yuan (jywang@ntu.edu.sg)

Introduction

Sewage sludge can be disposed by landfilling, incineration, incement production, in agriculture, and by composting. Theadvantage of agricultural use of sewage sludge is the recyclingof nitrogen, phosphorus, potassium, calcium, microelements,and organic matter for plant growth by the enhancement ofsoil. However, the obstacles for the agricultural use of sewagesludge are as follows: high content of heavy metals, lowconcentration of potassium, high moisture content, and thepossible presence of chemical pollutants. There is a certaintrend towards increasing the use of sludge in agriculture.Utilization of sewage sludge in agriculture in the USA was55% in 1997. Utilization of sewage sludge in agriculture inEuropean countries in 1990 varied from 10% in Greece to80% in Portugal and Luxembourg (Vesilind and Spinosa,2001). In Japan, 2.3 million cubic meters of sewage sludgeare produced annually and 24% have been reused onagricultural lands as a fertilizer after composting (Ito et al.,1998). However, sewage sludge in other Asian countries isusually disposed, not utilized (Vesilind and Spinosa, 2001).For example, 50,100 tonnes of sewage sludge and 1,081,700tonnes of food waste are disposed of in Singapore annually.

One of the ways of sludge utilization is composting.Composting lasts about one month and then long stages ofcuring and maturation follow. Starter cultures are rarely usedin the composting. The aeration usually is provided only byperiodical turning of the composting material. Composting of

sewage sludge occupied large space of land and is not suitablefor application in large scale in a country with a shortage ofland such as in Singapore.

The objective of the research was to develop a biotechnologyto convert dewatered sewage sludge, with addition of foodwaste, into biofertilizer.

Materials and methods

The fresh vegetable food waste was added to sewage sludgeto improve the chemical composition and the texture of theinitial material. The mixture of sewage sludge and food wastewas treated in a closed reactor under controlled aeration,stirring, pH, and temperature at 60oC, after addition of starterbacterial culture. Starter culture was selected by the growthrate on the sewage sludge as a source of the nutrients. It wasidentified by 16S rDNA gene sequencing as Bacillusthermoamylovorans. The biodegradation of sewage sludge,which is mainly anaerobic microbial biomass, was studied bydecrease of volatile solids (VS), content of organic carbon,autofluorescence of coenzyme F420, and the quantification ofspecific binding of Archaea-specific oligonucleotide probeArc915 (Tay et al., 2001). Stability of the biofertilizer wasmeasured by the carbon dioxide evolution rate. Thephytotoxicity was determined by the modified germinationand root Elongation Test. The quality of the fertilizer wasexamined by the plant test.

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Results and discussion

The removal of pathogens and parasites is an important taskin the thermophilic treatment of sewage sludge. The contentof viable and dead cells detected by Ent1432 oligonucleotideprobe, specific for enterobacteria, was measured by flowcytometry. It was shown that the treatment under 60oC didnot completely remove enterobacteria from the treated material.The content of enterobacteria1 cells increased from 78×106

cells/g of dry matter to 740×106 cells/g of dry matter after 4days of the biodegradation and then dropped to 36×106

cells/g of dry matter at the end of the process. The share ofviable enterobacterial cells (EBC) was changed from 3% to12% of viable bacterial cells (Table 1). The thermal pre-treatment of sludge at 100oC for 15 minutes increased theprocess rate, provided sludge disinfection and ensured thepresence of only starter bacterial culture.

The images of the particles of sewage sludge and food wasteduring bioconversion were made by scanning electronmicroscope (SEM) LEICA (Cambridge Ltd., UK) atmagnification of 5000 (Figure 2).

The colonization of the particles of sewage sludge and foodwaste by bacterial cells was determined as a ratio of the areaof the bacterial cells projection on the image and an area ofthe SEM image. Six images were analysed for onedetermination. The colonization of the particles was changedfrom 2.2±0.5% on the first day to 4.4±0.2% and 3.8±0.7%after six and eight days of bioconversion, respectively (mean± standard deviation are shown).

The degradation of anaerobic biomass was faster thanbiodegradation of total organic matter (VS) because thedecrease of VS, specific binding of Arc915 probe, andfluorescence of coenzyme F420 was 25%, 54%, and 87% ofinitial value, respectively. Absolute reduction of volatile solidsduring bioconversion was 74% from initial quantity. The bestfertilizer was obtained when sewage sludge was thermallypre-treated, mixed with food waste (in a ratio 1:1 by TS),chalk and artificial bulking agent. VS decreased on 24.8%,total carbon content on 13.5%, total and nitrogen content on27.7% of initial values (Table 2). For the comparison, 25-30%of degradation of volatile suspended solids is typical level ofaerobic stabilization of sewage sludge.

Table 1. Number of enterobacterial cells duringthe biodegradation process

Duration, daysTotal number Share of viableof EBC, 106/g EBC cells, %

0 78 12

2 300 3

4 740 5

13 36 7

The addition of food vegetable waste to sewage sludge, withthe ratio of 1:1 by total solids, was made to improve C/N ratioin initial material for bioconversion, to increase the potassiumcontent, to dilute the content of heavy metals and to reducethe viscous texture of sludge.

The study of the content of the main heavy metals in sewagesludge from Seletar municipal water reclamation plant showedthat their content was lower than the US and EU limits for thebiosolids. Food waste addition to sewage sludge provideshigher quality of final product (Figure 1).

Initial content of thermophilic bacterial cells was 4.2×107

c.f.u.g-1 107 to 6.0×108 c.f.u.g-1 of dry matter because thebiomass of thermophilic strain Bacillus thermoamylovoransSW25 was added at the beginning of bioconversion. It is thesame as the highest content of bacterial cells in hot syntheticcompost (Dees, Ghiorse, 2001).

Figure 1. The content of heavy metals in sewage sludgeand biofertilizer, % of EU limits

Figure 2. SEM image of the mixture of sewage sludgeand food waste after six days of bioconversion carried out

with addition of Bacillus thermoamylovorans SW25.

Table 2. Characteristics of the material andfinal products of bioconversion

Initial FinalParameters matter product

pH 7.2 6.9

TS, % 22.5 95.5

VS, % of TS 82.8 62.3

Ctotal, % of TS 37.7 32.5

Ntotal, % of TS 4.7 3.4

C/N 8.0 9.7

P, % of TS 0.4 0.4

K, % of TS 2.8 2.9

Bacterial content, c.f.u. g-1 4.2⋅107 2.2⋅108

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Biofertilizer had stability index of 1.3 mg CO2 g-1 OM d-1

measured by carbon dioxide evolution rate. It is consideredvery stable. The toxicity of the fertilizer from sewage sludgeand food waste determined by the seed germination tests waslower than that of commercial organic fertilizer.

The plant growth experiments were performed by the commonscheme of the research on agricultural use of sewage sludgeand compost (Wong and Su, 1997). Three plant species,cucumber, tomato and kangkong, were used to study theinfluence of the fertilizer on plant growth in the pots. Twokilograms of fresh subsoil with moisture of 14% were placedin every pot. No fertilizer was added to the control. Thefertilizers were added in the experimental pots at an applicationrate of 0.2, 0.5, 1.0, 1.5, and 1.75% (weight of dry matter/weight of soil). Each mixture was incubated one week priorto plant growth experiments for the release of nutrients. Thegrowth of the plants was observed and measured over a periodof 5 weeks. The optimal addition of fertilizer to the subsoilwas 1.0% – 1.5% (Figure 3). Addition of 1.0 – 1.5% of thisfertilizer to the subsoil increased the growth of different plantsby 113 – 164 % depending on the types of plants tested.Higher content of biofertilizer decreased the length and weightof the roots and the stems of the plants.

Figure 3. The growth of cucumber in the soil withdifferent content of the biofertilizer.

The biofertilizer produced by the bioconversion of sewagesludge and food waste was compared with that of raw sewagesludge, dried sewage sludge, and commercial organic fertilizerin the germination and root elongation test and in the plantexperiments. The biofertilizer, raw sludge and dried sludgewere added to the subsoil in dosage of 1% (weight of drymatter/weight of soil). Commercial organic fertilizer wasadded to the subsoil in quantity 0.7% (weight of dry matter/weight of soil) to equalize the content of nitrogen in thebiofertilizer and in commercial organic fertilizer. Best growthof the plants was obtained when biofertilizer was applied.The parameters of plant growth with biofertilizer were 40 –100% higher than the parameters of the plant growth on thesubsoil with the addition of raw sewage sludge or driedsewage sludge (Figure 4).

The results demonstrated that the biotechnology of theconversion of sewage sludge and food waste into biofertilizercan be used in larger scale application in Singapore.

Conclusion

1. A new biotechnology for intensive aerobic bioconversionof the mixture of sewage sludge and vegetable foodwaste into biofertilizer was developed and demonstratedpromising results.

2. The biofertilizer obtained from the mixture of pre-treatedsewage sludge and vegetable food waste in the ratio 1:1by TS was non-toxic, stable, and its application to thesoil resulted in faster growth and development ofagricultural plants.

References

[1] Dees, P.M., Ghiorse,W.C. 2001. Microbial diversity inhot synthetic compost as revealed by PCR-amplifiedrRNA sequences from cultivated isolates and extractedDNA. FEMS Microbial Ecology, 35, 207-216.

[2] Ito, A., Umita, T., Aizawa J., and Kitada K. 1998. Effectof inoculation of iron oxidizing bacteria on elution ofcopper from anaerobically digested sewage sludge. WaterScience and Technology, 38 (2), 63-70 (1998).

[3] Tay, S. T.-L., Ivanov, V., Kim, I.-S., Feng, L., Tay, J.-H.2001. Quantification of ratios of Bacteria and Archaeain methanogenic microbial community by fluorescencein situ hybridization and fluorescence spectrometry.World Journal of Microbiology and Biotechnology 17,583-589.

[4] Vesilind P.A., and Spinosa L. 2001. Sludge productionand characterization. Production and regulations. In:Sludge into Biosolids: Processing, Disposal, Utilization,Spinosa L. and Vesilind A. (eds.), IWA Publishing,London, pp. 3 – 8.

[5] Wong, J.W.C., Su D. C. 1997. Growth of Agropyronelongatum in an artificial soil mix from coal fly ashand sewage sludge. Bioresource Technology, 59 (1),57-62.

Figure 4. An effect of raw sewage sludge (RSS), dry sewagesludge (DSS), commercial organic fertilizer (COF) and

biofertilizer produced by the bioconversion of sewage sludge andfood waste (BF), on the growth of tomato.

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RSS DSS COF BF