Bioresource Technology 100 (2009) 5002–5009

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Bioresource Technology

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Systematic analysis of biochemical performance and the microbial communityof an activated sludge process using ozone-treated sludge for sludge reduction

Sang-Tian Yan a, Hao Zheng a, An Li a, Xue Zhang a, Xin-Hui Xing a,*, Li-Bing Chu a, Guoji Ding b,Xu-Lin Sun c, Benjamin Jurcik c

a Department of Chemical Engineering, Tsinghua University, Beijing 100084, Chinab School of Environmental and Chemical Engineering, Shanghai University, Shanghai 200072, Chinac Air Liquide Laboratories, Tsukuba-shi, Ibaraki-Pref. 300-4247, Japan

a r t i c l e i n f o a b s t r a c t

Article history:Received 3 March 2009Received in revised form 15 May 2009Accepted 17 May 2009Available online 9 June 2009

Keywords:Sludge reductionOzonationLysis-cryptic growthMicrobial activityMicrobial population

0960-8524/$ - see front matter � 2009 Elsevier Ltd. Adoi:10.1016/j.biortech.2009.05.029

* Corresponding author. Tel./fax: +86 10 62794771E-mail address: xhxing@tsinghua.edu.cn (X.-H. Xin

Two lab-scale bioreactors (reactors 1 and 2) were employed to examine the changes in biological perfor-mance and the microbial community of an activated sludge process fed with ozonated sludge for sludgereduction. During the 122 d operation, the microbial activities and community in the two reactors wereevaluated. The results indicated that, when compared with the conventional reactor (reactor 1), the reac-tor that was fed with the ozonated sludge (reactor 2) showed good removal of COD, TN and cell debris,without formation of any excess sludge. In addition, the protease activity and intracellular ATP concen-tration of reactor 2 were increased when compared to reactor 1, indicating that reactor 2 had a betterability to digest proteins and cell debris. DGGE analysis revealed that the bacterial communities in thetwo reactors were different, and that the dissimilarity of the bacterial population was nearly 40%. Reactor2 also contained more protozoa and metazoa, which could graze on the ozone-treated sludge debrisdirectly.

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1. Introduction

The use of activated sludge as a biological wastewater treat-ment process has been employed to treat a wide variety of waste-water. However, its primary by-product, excess sludge, has becomea large problem. Indeed, treatment and disposal of excess sludgecan account for up to 60% of the total operational costs of a waste-water treatment plant (Spellman, 1997). For this reason, it isimportant to develop methods of reducing excess sludge producedduring wastewater treatment in an economic, environmentallysafe and practical manner. One promising technique is the sludgeozonation process (Yasui et al., 1996; Wei et al., 2003; Yan et al.,2009). Ozone is a strong chemical oxidant capable of destroyingthe cell walls of microorganisms and solubilizing or oxidizing themto organic substances.

The sludge ozonation process has been defined as a sequence ofdecomposition processes that includes disintegration of suspendedsolids, solubilization of the solids (cells) and mineralization of thesoluble organic matter released from the microbial cells (Ahn et al.,2002). Recently, we comprehensively studied the mechanism ofsludge ozonation at different ozone dosages using a combinationof biological and chemical approaches (Yan et al., 2009), including

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.g).

analysis of the changes in biological response based on changes inthe CFU and PCR–DGGE profiles, bio-macromolecular activity andradical-scavenging activity. The results revealed that, when theozone dosage was as high as 0.14 to 0.27 g O3/g TSS, ozone failedto oxidize the sludge matrix efficiently due to the release of radicalscavengers such as lactic acid and SO2�

4 from the microbial cells inthe sludge. These findings indicate that prolongation of the sludgeozonation process causes the ozone to gradually lose its ability tooxidize sludge solids and soluble organic molecules. In addition,these findings indicated that the efficiency of sludge decomposi-tion by ozonation has its limits, even at high ozone doses. Thesludge ozonation process has been employed to reduce excesssludge by feeding ozonated sludge into the activated sludge (Yasuiand Shibata, 1994; Ahn et al., 2002; Bohler and Siegrist, 2004; Cuiand Jahng, 2004). During biological treatment, a portion of therecirculated ozone-treated sludge is oxidized to CO2 in the aerationtank, while another portion is turned into new microbial cells. Thisprocess is known as the lysis-cryptic growth process, which is con-sidered to be an important mechanism for the reduction of sludgeby sludge ozonation (Wei et al., 2003).

Many studies have evaluated the effects of introducing ozonat-ed excess sludge into a variety of activated sludge reactors, includ-ing traditional activated sludge reactors (Yasui and Shibata, 1994;Yasui et al., 1996), AO reactors (Cui and Jahng, 2004; Saktaywinet al., 2005), sequenced batch reactors (SBR) (Huysmans et al.,2001) and membrane bioreactors (MBRs) (Oh et al., 2007). In these

Table 1Experimental condition for reactors 1 and 2.

Items Reactor 1 Reactor 2

Volume (L)Bioreactor 4.6 4.6Settling tank 5.9 5.9HRT (h) 10 10Influent (average)COD (mg/L) 750 ± 65 790 ± 105TN (mg/L) 39.8 ± 2.2 40.0 ± 1.9TP (mg/L) 4.8 ± 0.5 5.0 ± 0.4Effluent (average)COD (mg/L) 67 ± 50 61 ± 42TN (mg/L) 20.2 ± 5.8 19.8 ± 4.9TP (mg/L) 2.4 ± 1.1 3.1 ± 1.1

MLSS concentration (mg/L) 2000 ± 400 1850 ± 450MLVSS concentration (mg/L) 1850 ± 400 1600 ± 450

S.-T. Yan et al. / Bioresource Technology 100 (2009) 5002–5009 5003

studies, a fraction of the recycled sludge passing through thereactor is always treated by ozonation, and the ozonated sludgeis then fed back to the aeration tank for biological treatment to-gether with the wastewater. Although ozonation-assisted sludgereduction processes have been successfully developed in practice,there are still several problems that should be considered. Forexample, the sludge system and wastewater conditions often differamong different experiments reported in previous studies; there-fore, comparison of activated sludge employing sludge ozonationis difficult. Furthermore, several studies have reported that in-creases in the amount of inorganic solid may occur in aerationtanks and effluent water quality may be dependent on the operat-ing conditions when the ozone-treated sludge was fed into the aer-ation tank (Yasui and Shibata, 1994; Yasui et al., 1996; Bohler andSiegrist, 2004). However, most previously conducted studies havefocused on the operating conditions and the performance of theactivated sludge process at the time of introduction of the ozonat-ed sludge (Yasui and Shibata, 1994; Yasui et al., 1996; Sakai et al.,1997; Ahn et al., 2002; Cui and Jahng, 2004; Lee et al., 2005; Sak-taywin et al., 2005). After the introduction of ozonation to the acti-vated sludge processes, nutrients released by cell lysis and celldebris may alter the influent characteristics of the entire system;however, there have been few in depth analyses of the biologicalperformance and microbial communities associated with the acti-vated sludge process when ozonated sludge was introduced to re-duce the sludge.

The sludge matrix includes bacteria, protozoa and metazoans,which all contribute to the function of the activated sludge in theaeration tank. Furthermore, each of the microorganisms has differ-ent growth patterns, growth rates and dependence on the environ-mental conditions. After the ozone-treated sludge consisting of thecell debris and soluble organics released from the disrupted cells(Yan et al., 2009) is returned to the bioreactor for degradation,the influent of the bioreactor will be altered. Indeed, the presenceof a large amount of cell debris and soluble organics in the influentleads to cryptic growth (Wei et al., 2003) when the ozonatedsludge solution is returned to the bioreactor, which may influenceall of the microbes in the bioreactor. Lysis-cryptic growth may beinduced in one of two ways. The first possibility is that the bacteriain the sludge may secrete hydrolysis enzymes that help hydrolyzethe debris. This causes changes in the bacterial hydrolysis activityand succession in the bacterial population (Mason et al., 1986). Thesecond way that lysis-cryptic growth may be induced is throughthe direct consumption of cell debris by protozoa and metazoansin the activated sludge, which leads to quick propagation of themicrofauna.

In this study, two lab-scale bioreactors (reactor 1 and reactor 2)were employed to systematically examine changes in the biologi-cal performance and microbial community structure of an acti-vated sludge process that was fed with ozonated sludge torealize sludge reduction. Reactor 1, which was used as a control,was a conventional activated sludge reactor from which the excesssludge was withdrawn periodically. Reactor 2 contained the sameactivated sludge as reactor 1, but this sludge was combined with abatch sludge ozonation process in which the excess sludge thatwas withdrawn was subjected to ozonation. The excess sludgewithdrawn from reactor 2 was treated by ozonation, after whichit was continuously fed to reactor 2. The sludge activities such asthe protease activity, intracellular ATP concentration and levelsof anti-oxidant enzymes such as superoxide dismutase (SOD) andcatalase (CA) were measured during the continuous operation.The bacterial population was analyzed by PCR–DGGE, and themicrofauna structure in the two reactors was evaluated by micro-scopic observation. Taken together, the results of this study canprovide a comprehensive basis for practical application of anozone-based sludge reduction system.

2. Methods

2.1. Lab-scale activated sludge reactors 1 and 2

Two lab-scale simulated activated sludge reactors (1 and 2)were established (Table 1). Each reactor contained a 4.6 L bioreac-tor and a 5.9 L settling tank. For start-up of the two reactors, thereturned sludge of a municipal wastewater treatment plant (Qin-ghe, Beijing, China) was inoculated and cultivated with syntheticwastewater (Feng et al., 2008) to domesticate the two reactorsfor about 30 d. The primary components of the synthetic wastewa-ter were glucose (chemical oxygen demand (COD) concentration ofabout 800 mg/L), (NH4)2SO4 (total nitrogen (TN) concentration ofabout 40 mg/L), KH2PO4 (total phosphorus (TP) concentration ofabout 5 mg/L), MgSO4 (9.5 mg/L), CaCl2 (0.91 mg/L) and Fe2(SO4)3

(0.14 mg/L). After 30 d of domestication, reactor 1 was operatedas a traditional activated sludge reactor, with 300 mL of excesssludge being withdrawn from the bioreactor once a day.

Reactor 2 contained a batch ozonation unit that was used toozonate 300 mL of the excess sludge that was discharged fromthe reactor at a dose of 0.15 g O3/g MLSS using a ozone generator(ED-OG-R4, Eco Design, Japan) (Yan et al., 2009). The ozonatedsludge was continuously returned to the bioreactor together withthe synthetic wastewater at a ratio of 0.3:8 (v/v). During the oper-ation in reactor 2, no sludge was discharged from the reactor. Fol-lowing treatment with 0.15 g O3/g MLSS ozone, approximately 70%of the mixed liquor suspended solids (MLSS) was reduced andapproximately 1000 mg/L soluble chemical oxygen demand(SCOD), 100 mg/L TN and 18 mg/L TP were released from the ozo-nated sludge. Table S1 (Supplementary data) shows the MLSS,SCOD, TN and TP of the sludge before and after ozone treatment.The results revealed that MLSS reduction rate of as high as 73.5%was obtained, indicating that the sludge in the lab was more easilyoxidized when compared with the practical sludge taken from thewastewater treatment plant (Yan et al., 2009). These differencesmight be caused by the sludge in the lab being cultivated usingsynthetic wastewater.

The average influent COD, TN and TP are shown in Table 1. TheHPLC method was used to determine the level of organic acid in theinfluent. The results of HPLC showed that the influent of reactor 2contained more organic acid than that of reactor 1, and that theseorganic acids were released by the ozonation of the sludge (datanot shown). The solid residence time (SRT) of the activated sludgewas approximately 15 d, and the hydraulic retention time (HRT) ofthe artificial sewage was 10 h. The dissolved oxygen (DO) was con-trolled at 3 mg O2/L in the two bioreactors and below 0.1 mg O2/Lin the settling tanks. The internal recycling rate (R) of the sludgefrom the settling tank to the bioreactor was controlled at about

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0.75 (v/v). To reach a stable state for an extended period, the totalrunning time of the two different reactors was greater than 120 dwithout temperature control (from April to August in 2008). Thetemperature of the reactors changed from 21 to 30 �C, with an in-crease from 22 to 27 �C occurring from the first to the 50th day(from May 26) and the temperature being maintained at approxi-mately 29 �C after the 60th day (Fig. S1).

For evaluation of the reactors, three identical samples in everymeasurement were taken at the same time to obtain the reliabledata. All the bars shown in the figures represented the standarddeviation (n = 3).

2.2. Chemical analysis of the reactors

During the 122 days of operation of the two reactors, 50 mLsludge samples were withdrawn three times a week and centri-fuged (12,000g, 10 min). The pellets were firstly heated at 105 �Cto examine the MLSS concentration, and then heated at 550 �C tomeasure the inorganic suspended solids concentrations. The con-centration of the mixed liquor volatile suspended solids (MLVSS)concentration was thus deduced from the two measurements.The effluent COD, TN and TP were measured three times a week.The effluent suspended solid (SS) concentrations were measuredby sampling 400 mL of effluent from the outlet of the settlingtanks. The COD, TN and TP concentrations were determinedaccording to the Chinese SEPA Standard Methods (House, 2002).The relative analytical errors for these parameters were 1–5%.

The organic acid content was determined by high-pressure li-quid chromatography (HPLC-10A, Shimadzu) using an SCR-102Gorganic acid analysis column (Shimadzu) and an RID-10A detector(Shimadzu) at a temperature of 40 �C. The mobile phase was 0.1%perchloric acid aqueous solution.

2.3. Biochemical and microbial community analysis of the reactors

The biological activity of the activated sludge in the two reac-tors was analyzed twice a week. Specifically, the protease activity,intracellular ATP concentration, total catalase (CA) activity andsuperoxide dismutase (SOD) activity were evaluated. The proteaseactivity was determined using a method that has been reportedelsewhere (Kim et al., 2002). One unit of the enzyme activity wasdefined as the amount of the enzyme that degraded 1 mg of azoc-asein in 60 min at 37 �C. The CA activity of sludge was determinedusing the procedure described by (Dallmier and Martin, 1988),with one unit of CA activity being defined as the amount requiredto decompose 1 lmol of H2O2 per min at 25 �C and pH 7. The SODactivity was measured using a SOD enzyme activity kit (JianchengTechnology, China). One unit of SOD activity was defined as theamount required to inhibit the rate of cytochrome c reduction by50% (Fisher et al., 2000). The intracellular ATP concentration wasmeasured with an ATP detection kit (Biyuantian Technology,China).

To evaluate the bacterial population, samples were analyzed byPCR–DGGE once a week. Briefly, 2 mL of fresh sludge were col-lected from each bioreactor and the total DNA was then extractedusing the phenol–alcohol extraction method (Ausubel, 1995). Thesamples were then subjected to PCR–DGGE under conditions thathave been described elsewhere (Yan et al., 2008) using the primersGC-341f and 907r (Muyzer et al., 1997). The Dice index (Cs), whichwas calculated from the DNA fingerprints, was used to evaluate thesimilarity of the bacterial community among and within the biore-actors after different operating times (LaPara et al., 2002). TheDGGE fingerprints were manually scored based on the presenceor absence bands without consideration of the band intensity.The pair-wise community similarity was determined using theDice index of similarity, Cs = 2j/(a + b), where j is the number of

common bands between samples A and B, and a and b are the totalnumbers of DGGE bands for samples A and B. The index rangedfrom 0 (no common band) to 1.0 (identical band patterns). Thesimilarity analysis was performed every week throughout the122 d operation period.

The DNA bands were excised from the DGGE gel using a previ-ously described method (Chen et al., 2007). The band’s PCR ampli-cons were then purified and ligated into the simple pMD18 cloningvector (TaKaRa Bio Inc., Otsu, Japan). Nearly 20 transformants wereused as template. Primers pMD18+(GAGCGGATAACAATTTCACA-CAGG) and pMD18-(CGCCAGGGTTTTCCCAGTCACGAC) were usedto amplify the DNA, and the amplicons were then subjected torestriction fragment length polymorphism (RFLP) analysis. Thenucleotide sequences were then determined by Shanghai Invitro-gen Biotech Co., Ltd. and compared with sequences in the GenBankdatabase (Benson et al., 1999). In addition, the sequences obtainedin this study were deposited in GenBank under the accession num-bers FJ463855 to FJ463861 and FJ750463 to FJ750468.

Protozoa and metazoans were counted weekly by microscopicobservation (Nikkon E-600, Japan). Three identical samples of thefresh sludge were used for the microscopic observation and everysample was observed three times. Ten fields of vision were ob-served for each observation and the average values were thenobtained.

3. Results and discussion

3.1. Performance of the two reactors

During the 122 d of operation, the average MLSS concentrationsof reactors 1 and 2 were approximately 2000 mg/L and 1850 mg/L,respectively (Table 1). In addition, the average MLVSS/MLSS was0.90 for reactor 1 and 0.88 for reactor 2 (Supplementary dataFig. S1). The slightly higher value for MLVSS/MLSS in reactor 1could be attributed to the slight accumulation of inorganic sus-pended solids in reactor 2 when compared with reactor 1 duringthe continuous introduction of the ozonated sludge. These findingsare similar to results of previously conducted studies (Yasui et al.,1996; Sakai et al., 1997). The average sludge volume index (SVI)values of the two reactors during the 122 d operation were 440and 400 mL/g, respectively, which indicates that the settling capa-bility of the activated sludge in reactor 2 was slightly better thanthat of reactor 1 (data not shown). These findings are similar tothose of the studies reported previously. The effluent SS from thesettling tanks of the two reactors was also examined during theoperation and found to be less than 5 mg/L for both reactors (datanot shown). No obvious increase in effluent SS from the settlingtank of reactor 2 was observed within the 122 d operation,although it has been reported that the effluent SS deterioratesslightly over time (Sakai et al., 1997). Based on the above results,there was no formation of excess sludge in reactor 2.

The water treatment performance of these two reactors isshown in Fig. 1a (COD), Fig. 1b (TN) and Fig. 1c (TP), and the aver-age change in the values of the effluent for the reactors over the122 d of operation are summarized in Table 1. Because previouslyconducted studies used different activated sludge reactors and dif-ferent types of wastewater, comparison of COD and TN treatmentabilities after feeding with the ozonated sludge is difficult. Forexample, Yasui and colleagues (Yasui and Shibata, 1994; Yasuiet al., 1996) reported that there was a slight increase in TOC con-centration when the reactor was fed ozonated sludge. In addition,Bohler and Siegrist (2004) found that the SCOD of the effluent in-creased in response to feeding of ozonated sludge. Conversely,Lee et al. (2005) and Sakai et al. (1997) observed no obvious in-crease in the effluent COD and BOD, respectively.

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Fig. 1. (a) Changes in the effluent COD of reactors 1 and 2 during the 122 doperation. (b) Changes in the effluent TN of reactors 1 and 2 during the 122 doperation. (c) Changes in the effluent TP of reactors 1 and 2 during the 122 doperation.

S.-T. Yan et al. / Bioresource Technology 100 (2009) 5002–5009 5005

The results of the present study indicated that the COD treat-ment capability of reactor 2 was almost completely unaffectedby feeding with ozone-treated sludge. Moreover, as shown inFig. 1b and Table 1, the average effluent TN in the effluent of the

two reactors was nearly the same, 20.2 and 19.8 mg/L (Table 1).When compared with the influent, approximately 50% of the nitro-gen was removed from the two reactors. The lack of obvious differ-ences in the TN removal of the two reactors evaluated in this studymay have occurred because the settling time in the settling tankwas about 12 h and the DO in the settling tank was nearly 0 mg/L, which may have induced denitrification in the tanks. Ahn et al.(2002) found that the TN removal efficiency of activated sludgefed with ozonated sludge was approximately 10% greater than thatof the control.

The average TP in the effluent of reactor 2 was nearly 30% great-er than that of the effluent of reactor 1 (Fig. 1c). The same phenom-enon has been reported by other researchers (Saktaywin et al.,2005). Since phosphorus could not be withdrawn with the excesssludge in reactor 2, other methods such as chemical precipitationand crystallization should be combined for phosphorus removal(Zhou et al., 2008).

The ozone-treated sludge contained decomposed proteins, DNA,polysaccharides and cell debris (Scheminski et al., 2000; Yan et al.,2009). After the ozonated sludge was returned to reactor 2, theinfluent of reactor 2 contained more debris, soluble COD, TN andTP than that of reactor 1; however, the effluent quality of reactor2 (except TP) did not deteriorate. The reason for this will be dis-cussed later.

Sakai et al. (1997) and Bohler and Siegrist (2004) found thatsome inert materials accumulated in the effluent of the activatedsludge process when it was conducted in combination with theozonation process. In this study, gel permeation chromatograph(GPC) analysis indicated that there were no larger molecularweight substances present in the effluent of reactor 2 (data notshown). Because synthetic sewage was used during the 122 d ofoperation, it may be difficult to form inert materials when com-pared with practical wastewater, which would explain why noobvious accumulation of inert materials was observed.

3.2. The microbial activity of the two reactors

The effluent water quality and other performance factors men-tioned above demonstrated that reactor 2 effectively removed thesoluble COD and debris introduced by feeding the reactor the ozo-nated sludge. Because the soluble components and cell debris inthe ozone-treated sludge had a high protein content, the biologicalactivity pertaining to the degradation of these components wasexamined (Yan et al., 2009). In this study, the protease and intra-cellular ATP concentrations were chosen as indicators of the sludgemicrobial activity after the ozone-treated sludge was fed back intothe bioreactor (Fig. 2). During the 122 d operation period, the pro-tease activity of reactor 2 was higher than that of reactor 1(Fig. 2a). Within the first 30 d, the protease activity in reactor 2had increased dramatically. However, the average protease activi-ties of reactors 1 and 2 for the entire experimental period were0.32 and 0.35 U/mg MLSS, respectively. The increase in proteaseactivity during the early stage suggests that the microbes in reactor2 increased their ability to decompose the proteins contained inthe soluble cell lysates, as well as the cell debris. Moreover, theaverage intracellular ATP concentration in reactor 2 throughoutthe 122 d operation period was 2.48 nmol/mg MLSS, which washigher than that of reactor 1 (1.75 nmol/mg MLSS) (Fig. 2b). A high-er ATP concentration may imply that reactor 2 showed better met-abolic reactivity.

Because ozone is a strong oxidizing agent, the radicals are alsopresent in the soluble part of the ozone-treated sludge (Yanet al., 2009). However, after recirculation of the ozone-treatedsludge, the respective CA and SOD activities of the activated sludgein the two reactors during the 122 d of operation were almost thesame level (Supplementary Data, Fig. S2a and b). These findings

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Fig. 3. Changes in the Dice index between the two reactors and each reactor’s Diceindex every two weeks during the 122 d operation.

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imply that the introduction of the ozonated sludge did not influ-ence the anti-oxidation activity of the sludge. This was becausethat because the half-life of ozone in water is only 15–30 min(Xu and Zhao, 2003), ozone and other radicals in the treated sludgesolution might have quickly deteriorated following feeding to thebioreactor in the present study due to the long retention time inthe storing tank.

A higher level of organic acid was released from the sludge andpresent in the influent of reactor 2. Some studies have shown thatozonated sludge can function as an effective carbon source fordenitrification (Ahn et al., 2002; Bohler and Siegrist, 2004; Cuiand Jahng, 2004). Therefore, the ozone-treated sludge that wasfed to reactor 2 may have functioned as a nutrient that led to anincrease in the microbial protease activity and ATP level in thesludge of reactor 2. For lysis-cryptic growth to occur (Masonet al., 1986), bacteria need to digest the sludge debris and the or-ganic materials using a hydrolysis enzyme; therefore, proteaseactivity was chosen as an indicator of hydrolysis in this study.Compared with the control reactor, the increase of the protease le-vel in reactor 2 was about 9.4%, demonstrating that the bacteriahad a better ability to digest proteins, so that the bacteria in thereactor could use the organics more easily. The influent of reactor2 contained a greater amount of cell debris and soluble organics;however, the reactor also contained higher levels of ATP whichwas increased by 41.7% compared with reactor 1, indicating that

the microbes were able to use organic materials and might explainthe more active microbial growth that was observed in reactor 2.Overall, these results indicated that feeding the soluble (COD)and insoluble (debris) organics contained in the ozonated sludgemay enhance the microbial activity in activated sludge, whichmaintained the performance of the wastewater treatment.

3.3. The microbial community diversity in the two reactors

Based on the results presented above, the protease activitiesand intracellular ATP content of the sludge were enhanced in reac-tor 2. Because the components in the influent of these two reactorsdiffered, the microbial communities of these two reactors mayhave differed during operation. The PCR–DGGE method is a usefultool to for analysis of the diversity of a microbial community (Muy-zer et al., 1997). The weekly DGGE banding patterns of reactors 1and 2 were used to calculate the microbial similarity, Cs, for thetwo reactors at the same operation time (Fig. 3). During the 122d of operation, the changes in the microbial community of eachreactor were also calculated based on the Cs values of the individ-ual reactors that were calculated using data obtained every twoweeks (Fig. 3). As shown in Fig. 3, the Cs of each reactor was around0.7 during the initial stage of operation, but this value increased togreater than 0.9 in the first 45 d, which suggests that the bacterialcommunities in each reactor were changing rapidly during the first45 d of operation. This was probably due to the increase in temper-ature from 22 to 27 �C that occurred during the first 45 d, from 22to 27 �C, which led to the increase of the bacterial populationdiversity in both reactors. After the temperature reached a con-stant level (around 29 �C), the Cs for each reactor also stabilized(at around 0.9), which implies that the bacterial populations ineach reactor became stable.

The differences in the microbial population of each reactor, thebacterial populations of reactors 1 and 2 were identical (the Cs was1) prior to introduction of the ozonated sludge since the sameseedling sludge was used for both reactors. After the first weekof operation, the bacterial communities in the two reactors didnot differ. However, after 10 d of operation, the Cs between thetwo reactors decreased to 0.92. This value then decreased rapidlyto 0.64 at around 45 d. After 45 d operation, the Cs between thetwo reactors remained constant (around 0.60). After reactor 2was fed with ozonated sludge, the bacterial population neededapproximately 10 d to adapt to the altered environment, after

Fig. 4. DGGE fingerprints of the activated sludge in reactors 1 and 2 during the 122d operation.

S.-T. Yan et al. / Bioresource Technology 100 (2009) 5002–5009 5007

which the difference in the communities increased dramatically.Indeed, in the first 45 days, the change of the Cs between the tworeactors was significant, especially during the first 30 days. Thesefindings indicated that the difference in the bacterial populationbetween the two reactors increased rapidly after reactor 2 wasfed with the ozonated sludge. Accordingly, when the recirculationof the ozone-treated sludge was started, the intracellular proteaseactivity in the activated sludge was triggered and increased dra-matically in the first 30 d. This increased activity helped the bacte-ria to utilize the proteins contained in the ozonated sludge, whichresulted in the alteration of the bacterial community structure.After approximately 45 d operation, Cs between the two reactorsbecame nearly stable, indicating that the microbes had adaptedto the ozone-treated sludge in the influent and the bacterial popu-lation had been stable.

Fig. 4 shows a comparison of the typical DGGE fingerprints ofthe two reactors during the 122 d operation, during which timethe bacterial population in each reactor was analyzed approxi-mately every 20 d. DGGE analysis indicated that the bacterial com-munities in the two reactors were the same prior the introductionof ozonated sludge. However, the DGGE fingerprints illustratedthat the communities in the two reactors were becoming muchmore revealed clear differences in the community structures ofreactors 1 and 2, which likely occurred due to the adaptation ofthe microbes to the changing conditions in the reactors. Because

Table 2Sequence length and closest phylogenetic affiliation of the bacteria in the reactors 1 and 2

No. Reactor Sequence length (bases) Phylogenetic relationship

Closet relatives

1 1 and 2 580 Sphaerotilus sp.2 1 and 2 567 Unclassified Nitrosomonas3 1 and 2 583 Uncultured Hydrogenophaga4 1 576 Uncultured Haliscomenobacter5 2 576 Uncultured Bacteroides6 1 and 2 583 Unclassified Methylibium7 1 and 2 494 Nitrospira sp.8 2 587 Methyloversatilis universalis9 2 587 Uncultured Betaproteobacteria10 1 and 2 577 Uncultured Bacteroidetes11 2 560 Uncultured Alphaproteobacterium12 1 and 2 562 Thiothrix eikelboomii13 1 and 2 574 Unclassified Proteobacteria

the sequences of each band of DGGE can provide detailed informa-tion regarding the bacterial structure of the microbial community,typical bands obtained from Fig. 4 were sequenced (Table 2). Bands1, 2 and 3 were the major bacteria at the beginning of the studyperiod. Band 1, corresponded to Sphaerotilus sp., which is a fila-mentous bacterium common in activated sludge. DGGE analysisrevealed that this organism was one of the major populationsthroughout the operation. Bands 2 and 3 corresponded to bacteriathat were dominant members of the community only during thebeginning stages of the experiment. Band 12, which representsThiothrix eikelboomii, was found in both the reactors. This organismbelongs to a filamentous group of bacteria that are common to acti-vated sludge. In this study, synthetic wastewater was used at a rel-atively high concentration; therefore, some filamentous bacteriawere found both reactors in the SEM observation after 90 d opera-tion (data not shown). However, the sludge floc structure did notdiffer greatly between the two reactors. DGGE analysis showedthat several bands (5, 8, 9 and 11) were found only in reactor 2,whereas 4, which corresponded to another common filamentousbacterium, Haliscomenobacter sp., was only found in reactor 1.Reactor 2 showed better settling than reactor 1 throughout theexperimental period; therefore, the difference in the filamentouspopulation may be related to this phenomenon. Although bands5, 9 and 11 corresponded to uncultured bacteria, band 8 found inreactor 2 corresponded to Methyloversatilis universalis, which canutilize methanol, methylated amines, formaldehyde and formate,as well as a variety of organic compounds (Kalyuzhnaya et al.,2006). The presence of this organism may account for the degrada-tion of cell-originated organics in the influent of reactor 2. Overall,the DGGE fingerprints of the samples from reactor 2 containedmore bands, which indicated that the bacterial population in reac-tor 2 was more abundant than that of reactor 1. Taken together,these results suggested that the bacterial population within reactor2 was altered in response to the introduction of the ozonatedsludge.

In addition to the bacterial population, the protozoa and meta-zoa also play important roles in the activated sludge. When celldebris is incorporated, the structure of the microfauna changes inresponse to the different operating conditions (Fried et al., 2000).During the 122 d of operation, the total number of protozoa andmetazoa in the two reactors increased rapidly during the first 40d, probably in response to changes in the bacterial population aswell as the increase in temperature. In addition, more protozoaand metazoans grew in reactor 2 than in reactor 1 (Fig. 5).

The difference in the structure of the microfauna in the tworeactors was also identified based on the morphology of the proto-zoa and metazoans. For protozoa, reactors 1 and 2 contained 15

.

Function in the sludge Accession no. % Similarity

Filamentous bacterium FJ463855 98Nitrification bacterium FJ463856 94– FJ463857 98Filamentous bacterium FJ463858 97– FJ463859 92– FJ463860 99Nitrification bacterium FJ463861 99Utilization of methanol, formate etc. FJ750463 100– FJ750464 99– FJ750465 99Filamentous bacterium FJ750466 99– FJ750467 100– FJ750468 99

0 10 20 30 40 50 60 70 80 90 100 110 120103

104

105

Prot

ozoa

and

met

azoa

(/m

g M

LSS)

Day (d)

Protozoa and metazoans Reactor 1 Reactor 2

Fig. 5. Changes in the total number of protozoa and metazoans in reactors 1 and 2(n = 3).

5008 S.-T. Yan et al. / Bioresource Technology 100 (2009) 5002–5009

and 17 species of protozoa, respectively. Among these, seven spe-cies of protozoa were commonly present in both of the reactors,including Centropyxis aculeate, Epistylis lacustris, Epistylis urceolata,Stentor polymorphru, Euplotes aediculatus, Paramecium aurelia andOpisthotricha similis. Centropyxis aculeate was the most commonprotozoa in both of the reactors, and reactor 1 contained more Epi-stylis lacustris than reactor 2. Conversely, seven species of protozoa,including Vannella platypodia, Discamoeba guttula, Epistylis plicatilis,Epistylis chrysemydis, Epistylis rotans, Vorticella putrina and Carche-sium batorligetiense were only found in reactor 1, while 10 speciesof protozoa, Mayorella hohuensis, Microchlamys patella, Gromia sp.,Paramecium trichium, Vorticella convallaria, Vorticella octava, Colepssp., Trachelophyllum pusillum, Chilodonella aplanata and Rhabdo-phrya sp. were present in reactor 2. The bacterial community Diveindex of the two reactors after 122 d (Fig. 3) was approximately0.6, indicating that the similarity of the bacteria communities inthe two reactors was approximately 60%; however, less than 50%of the protozoa present in the two reactors were the same species.

Four species of metazoans were present in reactor 1, whilethree were present in reactor 2. Two species of metazoa, Rotarianeptunia and Monostyla unguitata, were found in both reactors.Macrostomum tuba and Didymodactylos sp. were only found inreactor 1, while Nematoda sp. was only present in reactor 2. Thediversity of the metazoa in the two reactors was nearly the same;however, reactor 2 had more metazoa, especially Rotaria neptunia.The average number of protozoa and metazoans during the 122 dof operation in reactor 2 was increased by approximately 1.57times.

Protozoa grow either by grazing on suspended particulatematerial including bacteria cells or by predation upon other proto-zoa (Ginoris et al., 2007). Metazoans play an important role in theplant operation as predators via predation on bacterial cells andprotozoa. During the operation, the differences in the protozoaand metazoans of the two reactors indicated that feeding with celldebris contained in the ozonated sludge might trigger the growthof the predators in reactor 2. Moreover, the bacterial populationalso changed in response to the introduction of the ozonated,which might also have contributed to the alteration of the micro-fauna. However, these differences may also have been caused bythe differences in the influent of the two reactors. Therefore, therichness in microfauna combined with the lysis-cryptic growthprocess (Wei et al., 2003) contributed to the reduction in sludgein the bioreactor fed with the ozonated sludge.

4. Conclusions

The reactor fed with ozonated sludge (reactor 2) showed goodperformance for the removal of COD, TN and cell debris, whileforming no excess sludge. The mechanism analysis showed thatthe protease activity and intracellular ATP concentration in reactor2 were increased, and DGGE analysis revealed that the bacterialcommunities in the two reactors were different. Furthermore, reac-tor 2 contained more protozoa and metazoans than the controlreactor in both number and variety. These findings indicate thatthe positive impact of the feeding of ozonated sludge occurs viamaintenance of the diverse microbial community and metabolicactivity.

Acknowledgments

This study was supported by the Hi-Tech Research and Devel-opment Program of China (863 Plan) (Grant No. 2007AA06Z347),the National Science Fund of China (Grant No. 20777045) and theNational Key Basic Research Program (973 plan) (Grant No.2007CB109203).

Appendix A. Supplementary data

Supplementary data associated with this article can be found, inthe online version, at doi:10.1016/j.biortech.2009.05.029.

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