Improvement of nitrification efficiency by bioaugmentation in ...

Front. Environ. Sci. Eng. 2014, 8(6): 937–944 DOI 10.1007/s11783-014-0668-7

RESEARCH ARTICLE

Improvement of nitrification efficiency by bioaugmentation in sequencing batch reactors at low temperature Di CUI1, Ang LI (✉)1, Tian QIU1, Rui CAI1, Changlong PANG1, Jihua WANG2, Jixian YANG1, Fang MA (✉)1, Nanqi REN1 1 State Key Laboratory of Urban Water Resource and Environment, School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China 2 School of Life Science and Technology, Harbin Normal University, Harbin 150025, China

© Higher Education Press and Springer-Verlag Berlin Heidelberg 2014

Abstract Bioaugmentation is an effective method of treating municipal wastewater with high ammonia concentration in sequencing batch reactors (SBRs) at low temperature (10°C). The cold-adapted ammonia- and nitrite- oxidizing bacteria were enriched and inoculated, respectively, in the bioaugmentation systems. In synthetic wastewater treatment systems, the average NHþ 4 -N removal efficiency in the bioaugmented system (85%) was much higher than that in the unbioaugmented system. The effluent NHþ 4 -N concentration of the bioaugmented system was stably below 8 mg$L–1 after 20 d operation. In municipal wastewater systems with bioaugmentation, the –1 effluent NHþ 4 -N concentration was below 8 mg$L after þ 15 d operation. The average NH4 -N removal efficiency in unbioaugmentation system (about 82%) was lower compared with that in the bioaugmentation system. By inoculating the cold-adapted nitrite-oxidizing bacteria (NOB) into the SBRs after 10 d operation, the nitrite concentration decreased rapidly, reducing the NO2– -N accumulation effectively at low temperature. The functional microorganisms were identified by PCR-DGGE, including uncultured Dechloromonas sp., uncultured Nitrospira sp., Clostridium sp. and uncultured Thauera sp. The results suggested that the cold-adapted microbial agent of ammonia-oxidizing bacteria (AOB) and NOB could accelerate the start-up and promote achieving the stable operation of the low-temperature SBRs for nitrification. Keywords nitrification, sequencing batch reactors (SBRs), bioaugmentation, low temperature

Received March 8, 2012; accepted July 15, 2013 E-mail: [email protected], [email protected]

1

Introduction

Since the beginning of the 20th century, numerous studies have investigated nitrogen removal from municipal wastewater. Accordingly, various nitrogen removal processes have been developed to decrease the emission of nitrogen compounds that pollute aquatic environments [1,2]. Still, the demand for more efficient nitrogen removal processes is globally increasing [3]. SBR process is one of the traditional biologic nitrogen removal processes, and has been widely used to treat the municipal wastewater in China [4–7]. The advantages of SBRs are its simplicity, high stability, low operating cost, desirable ecological effects, and efficient nitrogen and phosphorus removal [7,8]. Nitrogen removal by SBR process includes two stages as follow: the first step is nitrification by autotrophic nitrifiers during aerobic reaction period, oxidizing the – influent NHþ 4 to NO3 ; the second step is denitrification through endogenous respiration by heterotrophic nitrosobacteria during an anoxic period [9]. Therefore, SBR is an effective biologic process to remove the total nitrogen in the wastewater under common condition [10]. In winter, the temperature of the municipal wastewater can range from 10°C to 25°C in northern China at locations with latitudes higher than 40° N [11]. The implementation of ammonia removal technology in wastewater treatment is difficult due to low efficiency of nitrification in winter months. The rate of nitrification by ammonia-oxidizing bacteria (AOB) and nitrite-oxidizing bacteria (NOB) is severely affected by temperature [10,12,13], since low temperatures ( < 15°C) not only decrease the specific growth rate of nitrifying bacteria, but also negatively affect their activity, substrate utilization rate, adsorption, as well as the settling ability of activated sludge [11,14–16]. Based on previous theoretical and experimental studies, bioaugmentation can effectively improve nitrification rate

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in biologic municipal wastewater treatment systems [17– 20]. Bioaugmentation by adding the nitrifying bacteria is an attractive alternative in the short solids retention time (short-SRT) nitrification process at cold temperatures [21]. Compared with the common methods, all bioaugmentation approaches can decrease the minimum SRT for nitrification [20,21]. However, there are few reports on the improvement of nitrification by bioaugmentation using AOB and NOB at low temperatures. Therefore, the application of bioaugmentation by inoculating acclimatized cold-adapted AOB and NOB could be a highly promising method for municipal wastewater treatment at low temperatures. The aim of this study is to use the bioaugmentation technology to accelerate the start-up and achieve the stable operation of the laboratory-scale low-temperature SBRs for removing the nitrogen pollution in the wastewater. The community structure of the acclimated cold-adaptive microorganisms and the nitrogen conversion efficiency of the bioaugmentation and unbioaugmentation processes at a low temperature (10°C) were investigated in details.

2

Materials and methods

2.1

SBR operating conditions and characteristics

Two sets of experimental SBR apparatuses were constructed as the nitrifying and denitrifying reactors. Figure 1 shows the schematic of the SBR, which was divided into two sections. One SBR system comprised two types of reactors, one containing unbioaugmented synthetic wastewater (R1) and the other bioaugmented synthetic wastewater (R2). The other system comprised two types of

reactors with unbioaugmented municipal wastewater (R3) and bioaugmented municipal wastewater (R4). Each reactor comprised influent and effluent valves, aerators, and an agitator. A mechanical mixer was set at a suitable rotation rate (about 125 r$min–1) in the SBR tank to ensure complete mixing during the aerobic and anoxic period [5]. During the aerobic operation period, the dissolved oxygen (DO) concentration in the mixed-liquor should be above 3 mg$L–1 according to previous reports [13]. The average DO concentration during the aerobic phase was fixed at 5.5–6.0 mg$L–1. The 12 L SBR systems were operated in successive cycles of 12 h as follows: 30 min static anoxic fill (wastewater feed with no mixing or aeration), 8 h mixed aerated reaction (mixing with the mechanical mixer and aeration), 2.5 h mixed anoxic reaction (mixing with the mechanical mixer and no aeration), 40 min settling (sludge settling with no mixing or aeration) and 20 min effluent discharge. Wastewater feed (6.0 L) was fed to these reactors during each cycle (i.e. 50% volumetric exchange ratio). All these reactors were placed in the cool room, and the operating temperature was maintained at 10°C. Characteristics of the synthetic and municipal wastewater in this study are listed in Table 1. The synthetic wastewater used in this study mainly contained (per liter) glucose (400 mg – 650 mg as carbon resource), (NH4)2SO4 (100 mg NHþ 4 N as N), 0.25 g NaH2PO4, 0.75 g K2HPO4, 5.0 g CaCO3, 0.03 g MgSO4$7H2O, and 0.01 g MnSO4$4H2O. 2.2 Inoculated sludge and microbial agent used in the bioaugmentation

These reactors were used to treat synthetic and municipal wastewater separately with aerobic activated sludge

Fig. 1 Schematic of the SBR system in this study

Di CUI et al. Improvement of nitrification efficiency by bioaugmentation in SBR at low temperature

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Table 1 Characteristics of the synthetic and municipal wastewater in this study wastewater

parameter

range

mean

pH

7.1

–

synthetic wastewater

–1

SS / (mg$L )

0

0

COD / (mg$L–1)

382–632

479

–1 NHþ 4 -N / (mg$L )

86.6–113.7

98.3

municipal wastewater

pH

6.5–8.1

SS / (mg$L–1)

110–210

160

211–428

313

40.9–58

50.4

–1

COD / (mg$L ) NHþ 4

–1

-N / (mg$L )

obtained from the secondary settling tank of the Harbin Taiping municipal wastewater treatment plant at 10°C. The average concentrations of mixed liquor suspended solids (MLSS) and mixed liquor volatile suspended solids (MLVSS) in two bioaugmented reactors were about 3020 mg$L–1 and 2380 mg$L–1, respectively. And the average concentrations of MLSS and MLVSS in two unbioaugmented reactors were about 3300 mg$L–1 and 2392 mg$L–1, respectively. The AOB and NOB were enriched separately in this study. The medium used for the enrichment and acclimation of AOB contained (per liter) 2.0 g (NH4)2SO4, 0.25 g NaH 2 PO 4 , 0.75 g K 2 HPO 4 , 5.0 g CaCO 3 , 0.03 g MgSO4$7H2O, and 0.01 g MnSO4$4H2O. The medium used for the enrichment and acclimation of NOB contained (per liter) 1g NaNO2, 1g Na2CO3, 0.25 g NaH2PO4, 0.75 g K 2 HPO 4 , 0.01 g MnSO 4 , 1 g CaCO 3 and 0.03 g MgSO4$7H2O. Bioaugmentation bacteria (AOB and NOB) were collected from the activated sludge system in the Harbin municipal wastewater treatment plant at 10°C. The microbial agent (including AOB and NOB) were enriched by repeated batch cultivation at low temperature (10°C), thereby efficiently converting ammonia to nitrite or nitrite to nitrate. The suspended microbial agent of AOB was added to R2 and R4 with a total dry mass of 75 mg$L–1 at the beginning of operation. After 10d operation, the NOB agent was added into R2 and R4 with a total dry mass of 75 mg$L–1 in order to reduce the nitrite accumulation. The total inoculation biomass amounts including AOB and NOB were about 5% (the dry weight of the supplemented microbial agent to MLSS). 2.3

Analysis of microbial community in SBRs

Polymerase chain reaction-denaturing gradient gel electrophoresis (PCR-DGGE) can effectively monitor microbial communities in environmental samples without inherent cultivation biases [22,23]. As shown in Table 2, biosamples were collected from the four tanks (R1 to R4) of the SBRs at the initial operation stage (day 5) and stable operation stage (day 30). A 50 mL sample was collected from each biologic treatment unit in the SBRs using a

plastic dipper. The sample was decanted to remove as much foam as possible before the liquid was transferred to a sterile tube. The collected samples were centrifuged at 8000 r$min–1 for 20 min, and the pellets were stored at – 80°C until analysis. Table 2 Designations and locations of the collected samples SBRs

samples initial operation stage (day 5)

stable operation stage (day 30)

R1

lane 1

lane 2

R2

lane 3

lane 4

R3

lane 5

lane 6

R4

lane 7

lane 8

Genomic DNA was directly extracted using a Bacterial Genomic DNA Extraction Kit (TaKaRa, Dalian, China) for PCR-DGGE analysis. The universal primers for the V3 region of the 16S rDNA genes were R534-GC (5′CGCCCGCCGCGCGCGGCGGGCGGGGCGGGGGCACGGGGGGATTACCGCGGCTGCTGG-3′), which contained an X-base GC clamp, and F101 (5′TGGCGGACGGGTGAGTA-3′) [24]. The gene fragments were amplified through a PCR thermal cycler dice (BioRad Ltd., Tokyo, Japan) using rTaq polymerase (TaKaRa, Dalian, China). The PCR procedure was as follows: 8 min initial denaturation at 94°C, 30 cycles of 94°C for 40 s (denaturation), 55°C for 40 s (annealing) and 72°C for 30 s (extension), followed by an extension at 72°C for 10 min [11]. The PCR products were stored at 4°C and detected by electrophoresis on a 1% agarose gel. The GC-clamp PCR products were separated according to their sequences with a DCode Universal Mutation System (Bio-Rad Ltd., USA). The samples were applied onto 8% (w/v) polyacrylamide gels in a running Trisacetate-EDTA buffer. The denaturing gradient ranged from 35% to 60% (7 mmol$L–1 urea and 40% (v/v) formamide as 100% denaturants). Electrophoresis was performed at 60°C, initially at 20 V for 30 min and then at 150 V for 9 h. After electrophoresis, the gels were stained with 120 mL of

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0.5 Tris-HCl EDTA-2Na buffer containing 12 μL of GeneFinder (BIO-V Ltd., Xiamen, China), and then photographed using a gel-imaging instrument (Bio-Rad Ltd., Tokyo, Japan). The bacterial community structures were analyzed by visually identifying DNA bands that migrated at various distances in each lane on the denaturing gels [11]. 2.4

Analytical methods

Effluents and influents of the SBR process were collected for the off-line testing of ammonia nitrogen (NHþ 4 -N), NO2– -N, NO3– -N and chemical oxygen demand (COD) according to the existing standard methods [25]. The DO concentration and temperature of the wastewater were measured using a DO sensor (YSI 550A, USA).

3

Results and discussion

3.1

COD removal performance in SBRs

The COD removal performances in the SBRs are shown in Fig. 2. The performances of R1 to R4 were evaluated based on effluent quality and whether each SBR met Level B of –1 Criteria I (COD£60 mg$L–1, NHþ 4 -N£8 mg$L ) of the Chinese National Discharge Standard of Pollutants for Municipal Wastewater [26]. As shown in Fig. 2(a), the average influent COD concentrations of synthetic wastewater from R1 (unbioaugmented) and R2 (bioaugmented) were approximately 479.7 mg$L–1. The average effluent COD concentration was 88.8 mg$L–1 (average COD removal rate is 83.8%) for R1 and 67.9 mg$L–1 (average COD removal rate is 85.8%) for R2. As shown in Fig. 2(a), the effluent COD concentration of these two reactors was all beginning to decrease to 60 mg$L–1 after 8d operation. Figure 2(b) shows that the average influent COD concentrations of municipal wastewater from R3 and R4 were approximately 313.3 mg$L–1. For municipal wastewater treatment, the average effluent COD concentrations were 70.4 mg$L–1 (average COD removal rate is 77.5%) for R3 and 48.4 mg$L–1 (average COD removal rate is 84.6%) for R4. The effluent COD concentrations of R2 and R4 met effluent quality standards when the two SBR systems were assessed on the 5th day (initial operation stage) and 30th day (stable operation stage). Two previous reports indicated that low temperature is one of the biggest obstacles in the start-up of biologic system operations, particularly when wastewater temperatures vary between 13 and 22 and 12–23°C, respectively [4,27]. Bioaugmentation can lead to rapid COD removal (using heterotrophic microorganisms to degrade organic matter at low temperatures) and nitrification. No association is found between COD removal and nitrification. However, the improvement of COD removal in R2 and R4

Fig. 2 Variations in the COD and removal efficiencies using the SBRs for synthetic (a) and municipal (b) wastewater treatment

could be caused by the increase of MLVSS/MLSS value (from 0.72 to 0.80) by bioaugmentation. 3.2

Nitrification performance in SBRs

Figure 3 shows the daily variations in effluent NHþ 4 -N concentrations throughout the entire experimental period. Figure 3(a) shows that the effluent NHþ 4 -N concentration in bioaugmentation system sharply decreased from day 0 to day 7, whereas concentration in the unbioaugmentation system slightly decreased in the first 5 days. These observations indicated that nitrification was efficiently promoted after adding the special bioaugmentation bacteria (AOB and NOB in this study). The conversion efficiency of nitrogen compounds to nitrite was also higher


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concentration with bioaugmentation was below 8 mg$L–1 after 15 d operation. Compared with bioaugmentation system, the effluent NHþ 4 -N concentration of unbioaugmentation system was a little higher, but also met the Chinese National Discharge Standard of Pollutants for Municipal Wastewater after 17 d operation. Meanwhile, the average removal efficiency of NHþ 4 -N in bioaugmentation system (about 88%) was higher than that in unbioaugmentation system (about 82%). These results suggested that the inoculated cold-adapted AOB could promote the ammonia removal, especially for treating the high ammonia concentration wastewater [4]. Bioaugmentation is a cost-effect method to shorten the start-up time of SBRs and reduce the nitrogen pollution discharge at low temperature. 3.3

Fig. 3 Variations in the NHþ 4 -N and removal efficiencies for synthetic (a) and municipal (b) wastewater treatment in the SBRs

in R2 than in R1. The average effluent NHþ 4 -N –1 concentrations reached 15.2 and 51 mg$L for the bioaugmented and unbioaugmented reactors in the synthetic wastewater system at a low temperature (10°C), respectively. The average NHþ 4 -N removal efficiency increased from 48% in the unbioaugmented system to 85% in the bioaugmented system. After 20 d operation, the effluent NHþ 4 -N concentration of the bioaugmented system met the Chinese National Discharge Standard of Pollutants for Municipal Wastewater, but the unbioaugmented system cannot obtain lower effluent NHþ 4 -N concentration during the whole process [26]. A direct comparison of NHþ 4 -N removal by municipal wastewater SBR operation can be obtained from Fig. 3(b). The average influent concentration of NHþ 4 -N in municipal system was 50 mg$L–1, which was much lower than that in synthetic wastewater. The effluent NHþ 4 -N

Variation tendency of nitrite and nitrate in SBRs

As shown in Fig. 4, the nitrite concentrations increased rapidly in the first 5 days, whereas the nitrate levels were low (even close to 0 mg$L–1) in the bioaugmentation processes for both synthetic and municipal wastewater. The accumulation of nitrite was slower in the unbioaugmented processes (R1 and R3) than in the bioaugmented processes (R2 and R4). As in a previous report, nitrification is completed when the wastewater temperature is maintained at 20°C, the pH is 7.0, and the DO concentration is high [3]. Thus, nitrite does not accumulate in activated sludge processes at common temperature. However, in the present study, the behavior of nitrite was the opposite, and its concentration increased upon inoculation with the cold-adaptive ammonia-oxidizing microorganisms in R2 and R4. One possible reason was that the cold-adapted ammonia-oxidizing bacteria (AOB) were the first to consume oxygen to convert NHþ 4 -N to NO2-N before other microorganisms did [28]. This result confirmed that inoculating AOB in SBRs promotes the conversion of ammonia to nitrite [11,29]. However, NOB was still limited in this bioaugmentation system. Therefore, nitrite accumulation has been found in these SBRs. The inhibition effect of surplus NO2– -N in the bioaugmented reactors (R2 and R4) was avoided by inoculating the cold-adapted NOB into the SBRs after the 10d operation. The nitrite concentration decreased upon the inoculation of NOB. The NO3– -N levels showed a temporal accumulation period from day 17 to day 25. The result showed that the nitration reaction was dominated by the NOB, which were able to sustain the activity at low temperature [11,30]. Meanwhile, the nitrite and nitrate levels continued to decrease and remained at low levels after 26d in the bioaugmentation system. It suggested that the bioaugmentation by adding both cold-adapted AOB and NOB could improve the ammonia removal and avoid the nitrite accumulation effectively at low temperature.

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Fig. 4 Variations in the NO3– -N=NO2– -N concentrations for synthetic (a) and municipal (b) wastewater treatment in the SBRs

3.4 Microbial community structure at the low temperature and different treatment units of SBRs

The results of the PCR-DGGE fingerprints obtained using DNA directly extracted from activated sludge are shown in Fig. 5. The DGGE profile indicates a highly diverse bacterial population, with a number of bands present in all samples although with different intensities. The total number of dominant bands was higher (over 12 dominant bands) in the activated sludge samples from R2 and R4 than from R1 and R3. The microbial community diversity was high in the bioaugmentation systems. Cluster analysis showed that the similarity coefficient exceeded 60% among the samples collected from the R2 and R4 SBRs. The microbial community was relatively stable in the bioaugmented SBRs (R2 and R4) compared with that in

the unbioaugmentation systems (R1 and R3). However, the similarity coefficients were all below 60%, except for those between samples R1 and R3, which were not inoculated with bioaugmentation bacteria. As shown in Fig. 5, the intensity of some bands, such as bands 1, 2, 6, 8, and 10, were very high in all samples, representing the dominant microorganisms in all SBRs. DNA in the dominant bands were collected and sequenced. Bands 1, 2, 6, 8, and 10 were found to be similar to Alishewanella sp. (EU287929), uncultured Nitrospira sp. (HQ658788), uncultured Paracoccus sp. (GQ985497), uncultured Synergistetes bacterium (JQ346774), and uncultured bacterium (FJ672648), respectively. Based on the DGGE fingerprints in Fig. 5, R2 and R4 (bioaugmented) showed a highly diverse bacterial population. Bands 5, 11, and 12 appeared in the R2 and R4 treatment units, and were identified as uncultured Dechloromonas sp. (HQ658773), uncultured beta proteobacterium (AM259327), and uncultured Nitrospira sp. (HQ658787), respectively. Uncultured Dechloromonas sp. (HQ658773) showed similarity to AOB, which may promote nitrifica– – tion from NHþ 4 -N to NO2 -N, then to NO3 -N at low temperatures [29]. The results showed consistency with the NO2– -N test scatter plot (Fig. 4), indicating that nitrite was rapidly amassed at the beginning of the SBR operation. Band 12 was closely similar to a nitrite-oxidizing bacterium that positively contributes to the conversion of nitrite compounds. Comparing the DGGE fingerprints of the stable operation stage on day 30 with those of the initial operation stage on day 5, the microbial community in each treatment unit barely changed, and the bio-diversity as well as intensity remained constant. Only a few non-dominant bands appeared, namely, bands 17, 18, and 20, which were identified as uncultured Clostridium sp. (GQ985493), uncultured Thauera sp. (JN125305), and bacterium enrichment strains (GU476614) in the stable operation of the SBRs. Uncultured Clostridium sp. (GQ985493) and uncultured Thauera sp. (JN125305) may be involved in the conversion of NO2– -N to NO3– -N, considering that these genera are known to participate in such processes [31]. Based on the results of the 16S rDNA sequences, the functional and dominant bacteria in the bioaugmented SBRs (R2 and R4) were successfully determined. In summary, the functional microbial microorganisms of AOB and NOB for nitrification has been determined in the bioaugmentation system, which were similar with uncultured Dechloromonas sp. (HQ658773), uncultured Nitrospira sp. (HQ658787), uncultured Nitrospira sp. (HQ658788), Clostridium sp. (GQ985493) and uncultured Thauera sp. (JN125305).

4

Conclusions

Bioaugmentation using the cold-adapted AOB and NOB embodies superiority at low temperatures in terms of


5.

6.

7.

8.

9.

Fig. 5 DGGE fingerprints showing the microorganism population dynamics in the SBRs

10.

11.

ammonia removal performance. It can accelerate the startup and promote achieving the stable operation of SBR process for ammonia removal. Furthermore, combination AOB and NOB could also avoid the nitrite accumulation effectively. Our findings in this study can help in the further construction of the cold-adapted microbial agent of AOB and NOB for the field application in the future. Acknowledgements This work was supported by grants from the National Creative Research Group from the National Natural Science Foundation of China (No. 51121062), the National Natural Science Foundation of China (Grant Nos. 51108120 and 51178139), and the 4th China Postdoctoral Science Special Foundation (No. 201104430).

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