Journal of Bioscience and Bioengineering VOL. xx No. xx, 1e5, 2015 www.elsevier.com/locate/jbiosc
Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors Tulasi Venkata Krishna Mohan,1, * Kadali Renu,1, 2 Yarlagadda Venkata Nancharaiah,1 Pedapati Murali Satya Sai,3 and Vayalam Purath Venugopalan1 Water and Steam Chemistry Division, Bhabha Atomic Research Centre, Kalpakkam 603102, India,1 School of Biosciences and Technology, VIT University, Vellore 632014, India,2 and Waste Immobilization Plant, NRB, Bhabha Atomic Research Centre, Kalpakkam 603102, India3 Received 22 April 2015; accepted 25 May 2015 Available online xxx
A 6-L sequencing batch reactor (SBR) was operated for development of granular sludge capable of denitrification of high strength nitrates. Complete and stable denitrification of up to 5420 mg LL1 nitrate-N (2710 mg LL1 nitrate-N in reactor) was achieved by feeding simulated nitrate waste at a C/N ratio of 3. Compact and dense denitrifying granular sludge with relatively stable microbial community was developed during reactor operation. Accumulation of large amounts of nitrite due to incomplete denitrification occurred when the SBR was fed with 5420 mg LL1 NO3eN at a C/N ratio of 2. Complete denitrification could not be achieved at this C/N ratio, even after one week of reactor operation as the nitrite levels continued to accumulate. In order to improve denitrification performance, the reactor was fed with nitrate concentrations of 1354 mg LL1, while keeping C/N ratio at 2. Subsequently, nitrate concentration in the feed was increased in a step-wise manner to establish complete denitrification of 5420 mg LL1 NO3eN at a C/N ratio of 2. The results show that substrate concentration plays an important role in denitrification of high strength nitrate by influencing nitrite accumulation. Complete denitrification of high strength nitrates can be achieved at lower substrate concentrations, by an appropriate acclimatization strategy. Ó 2015, The Society for Biotechnology, Japan. All rights reserved. [Key words: Carbon to nitrogen ratio; Denitrification; Granular sludge; High strength nitrate; Nitrite accumulation]
Nitrate is one of the most widespread contaminants of natural water resources. During the past many decades, anthropogenic activities have greatly contributed to the increase in nitrate concentrations in surface and groundwater (1e3). Nitrate contamination of aquatic environments is a concern as it has been linked to ecosystem well-being and public health (4). For example, nitrate contamination causes eutrophication of water bodies and bluebaby syndrome in infants (2,5). Due to potential risks on ecosystem and public health, nitrate is listed as a priority pollutant under US Environmental Protection Agency (6). The maximum contaminant level (MCL) for nitrate (NO-3) in drinking water supplies has been set at 45 mg L1 by US EPA, WHO and Bureau of Indian Standards. The European Union has prescribed 50 mg L1 as the limit for nitrate and 0.1 mg L1 for nitrite in drinking water (7). Two main point sources for nitrate contamination are sewage and industrial effluents. High strength nitrate effluents are generated in fertilizer, ammunition, pharmaceutical, metal finishing and nuclear industries (8e10). In nuclear industry, nitric acid is used in various processing steps of nuclear fuel fabrication, dissolution and reprocessing of spent nuclear fuel, resulting in production of low and high strength nitrate-bearing effluents (8,11). The acidic nitrate effluents thus generated are neutralized with hydroxide and stored under safe conditions for treatment and disposal in future.
Biological denitrification for treatment of high strength nitrate effluents has been widely studied, particularly in the case of wastes generated in nuclear fuel cycle operations. It is acknowledged that biological treatment of high strength nitrate wastes typical for industrial effluents is challenging (8,9,11e14). One of the problems commonly encountered during denitrification of high strength nitrates is accumulation of high levels of nitrite (8,11,12,14). Accumulation of high concentrations of nitrite can lead to inhibition of the whole denitrification process (8). Two main factors that could influence nitrite accumulation during denitrification are electron donor supply and composition of the microbial community (15). Nitrate effluents generated in the nuclear industry are often devoid of electron donors. Dosage of electron donor in terms of C/N ratio has been studied for minimizing nitrite accumulation and achieving complete denitrification (12). Regarding the microbial community, it has been reported that biomass in the form of granular sludge may be better for overcoming the inhibitory effects of nitrite concentrations encountered during high strength nitrate denitrification (10). The objectives of this study, therefore, were (i) cultivation of granular sludge for the denitrification of high strength nitrate wastewater and (ii) understanding the effect of C/N ratio on the denitrification process.
MATERIALS AND METHODS * Corresponding author. Tel.: þ91 44 27480203; fax: þ91 44 27480097. E-mail address: [email protected] (T.V. Krishna Mohan).
Seed sludge Activated sludge collected from the outlet of an aeration tank of an operating municipal wastewater treatment plant at Kalpakkam, India was used as
1389-1723/$ e see front matter Ó 2015, The Society for Biotechnology, Japan. All rights reserved. http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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inoculum for the sequencing batch reactors (SBRs). The activated sludge was black in color, contained filamentous microorganisms and exhibited poor settling characteristics (16,17). After bringing to the laboratory, the sludge was washed a few times with deionized water and stored at 4 C. Prior to use, the activated sludge was washed with ultrapure water and was placed on a tissue paper to remove excess water. The wet sludge was weighed and transferred to the SBR. Cultivation of denitrifying granular biomass A glass tank (height: 61 cm, diameter: 15 cm) with 6 L working volume was operated in a sequencing batch mode. SBR operation was chosen as it allows formation and selection of granular sludge. The SBR was inoculated with activated sludge (3 g mixed liquor suspended solids L1) and operated with 24 h cycle period. The SBR cycle consisted of the following: 5 min fill, 23 h reaction, 5 min settle, 10 min decant and 40 min idle. The SBR was fed with simulated nitrate waste prepared in deionized water and consisting of the following ingredients (g L1): MgSO4.7H2O 0.04, KCl 0.02, K2HPO4 0.03, KH2PO4 0.01, CaCl2.2H2O 0.02. Trace elements were supplied by adding 0.1 mL of trace element mix per 1 L of simulated nitrate waste (18). Sodium acetate and sodium nitrate were added to the simulated nitrate waste, as per requirement (Table 1). The SBR was operated at room temperature (w30 C) with 50% volumetric exchange ratio. Mixing was provided at a fixed speed of 100 rpm using a stirrer (Eltek, India). Effluent was drawn from a port located 17 cm above the bottom. The SBR was operated for more than 8 months in order to monitor the stability of denitrification process. Microscopy of denitrifying sludge Morphology of the denitrifying sludge was documented using a DP70 camera (Olympus, Japan) attached to SMZ1000 stereozoom microscope (Nikon, Japan) (16). The denitrifying granular sludge was stained with BacLight bacterial viability kit (Molecular Probes, USA) (19). A 200 mL of BacLight stain mixture (SYTO 9 and propidium iodide) was added to 1.5 mL Eppendorf tube containing the sludge and incubated on shaker set at 100 rpm. After 15 min, the sludge was washed twice with ultrapure water. The stained granules were imaged using a confocal laser scanning microscope TCS SP2 AOBS (Leica Microsystems, Germany) attached to an inverted microscope (Leica DMIRE2). Images were collected using a 63x 1.2 NA water immersion objective (20). Sample was excited using a 488 nm Ar laser line. The fluorescence emission was collected between 500 and 520 nm for SYTO 9 and between 600 and 680 nm for propidium iodide. DNA extraction and PCR-DGGE Activated sludge and denitrifying granular sludge (0.2 g wet weight) collected from the SBR were homogenized by vortexing with glass beads (21). Genomic DNA from the homogenized samples was extracted using DNA isolation kit (Qiagen, Germany). The extracted DNA was verified by 1% (w/ v) agarose gel electrophoresis. The extracted genomic DNA was stored at 50 C. The V3 region of the bacterial 16S rRNA gene was amplified using primers PRBA338f (50 CGCCCGCCGCGCGCGGCGGGCGGGGC GGGGGCACGGGGGGACTCCTACGGGAGGCA GCAG 30 ) and PRUN518r (50 ATTACCGCGGCTGCTGG 30 ) (19). Polymerase chain reaction (PCR) was performed using 50 mL reaction mixture using a Mastercycler gradient thermal cycler (Eppendorf AG, Germany) as described earlier (19). The PCR amplified fragments were separated by denaturing gradient gel electrophoresis (DGGE) using an INGENYphorU system (Ingeny International BV, The Netherlands). The DGGE was performed using 40%e70% urea-formamide denaturant gradient gel [8% (w/v) acrylamide solution (40% acrylamide and bisacrylamide in 37.5:1 ratio)] in TAE buffer (40 mM Tris base, 20 mM acetic acid, 1 mM EDTA, pH 8.0) as described earlier (20,22,23). After electrophoresis, the gel was stained with ethidium bromide (0.1 mg L1) for 30 min and destained for 15 min in ultrapure water. The stained gel was documented with INFINITY gel documentation system (Vilber Lourmat, France) and diversity indices were determined as Ward’s Dice similarity coefficients with the Quantity One version 4.6 (Bio-Rad, USA).
RESULTS AND DISCUSSION Denitrification performance The SBR was inoculated with activated sludge and fed with simulated waste containing 1354 mg NO3eN L1 at a C/N ratio of 3. It may be noted that the starting concentration in the reactor would be 677 mg NO3eN L1, because the SBR retains 50% of waste from the previous batch (50% VER), with no nitrates in it. The pH of simulated nitrate waste was 7.5 before feeding to the reactor. The pH observed in the reactor at the end of cycle was 9.5, the increase being due to denitrification. In subsequent cycles of operation, the pH in the reactor stabilized at w9.5 (Fig. S1). The simulated waste prepared in deionized water and the dissolved oxygen (DO) content before feeding was noted to be 6.0 mg L1. The DO rapidly decreased to below 0.08 mg L1 within a few minutes of addition to the reactor, due to aerobic microbial respiration. The reactor tank was open to atmosphere and the mixing was provided by means of bottom stirring at 100 rpm. Under these conditions, the DO in the SBR cycle period was always below 0.08 mg L1. The reactor operating conditions were conducive for heterotrophic denitrification, observed from cycle 1 onwards (Fig. 1). By keeping the C/N ratio constant at 3, feed nitrate-N was increased to 2710 and then to 5420 mg L1. Complete denitrification was observed at the increased feed nitrate concentrations of 2710 and 5420 mg NO3eN L1. The effluent nitrate and nitrite concentrations were invariably less than 10 mg L1 during long term operation of the SBR, except during a few occasions (Fig. 1). Denitrification of feed NO3eN concentrations of 1354, 2710 and 5420 mg L1 was completed within the first 4, 6 and 8 h, respectively (Fig. S2). Typical denitrification profiles consisted of nitrate removal, accumulation of nitrite and nitrite removal. Nitrite accumulation reached a maximum value of approximately 250 mg L1 NO2eN at 1354 mg L1 nitrate-N in the feed. In order to study the denitrification at reduced substrate concentration, the reactor was subsequently fed with 5420 mg L1 NO3eN at a C/N ratio of 2. The change in C/N ratio caused inhibition of denitrification, resulting in incomplete denitrification and accumulation of nitrite-N as high as 3500 mg L1 (Fig. 2). SBR operation in the subsequent cycles consistently showed the buildup of the high concentrations of nitrite and nitrate. Inhibition of the denitrification process was probably due to the toxicity caused by high nitrite levels.
Analytical procedures Liquid samples collected at regular time intervals during the SBR cycle period were monitored for nitrite and nitrate. Nitrate was analyzed by high performance liquid chromatography (HPLC) (Dionex Ultimate 3000) fitted with an Acclaim OA column using a UVevis detector set at 210 nm (10). Dilute H2SO4 (0.003 N) was prepared in deionized water and used at a flow rate of 0.7 mL min1. Nitrite was estimated spectrophotometrically by reaction with N-(1naphthyl) ethylenediamine dihydrochloride (24). Mixed liquor suspended solids (MLSS) and total organic carbon (TOC) were occasionally analyzed as per standard methods (24).
TABLE 1. The simulated nitrate waste was prepared by varying the amounts of sodium acetate and sodium nitrate. Constituent (g L1)
NaNO3 CH3COONa (3:1)a CH3COONa (2:1)a a
Feed NO3eN (mg L1) 1354
2710
4060
5420
8.22 13.89 9.26
16.43 27.77 18.52
24.86 41.66 27.77
32.86 55.55 37.03
Sodium acetate was added to give a C/N mass ratio of 3 or 2, i.e., 3 or 2 g of carbon for g of nitrogen.
FIG. 1. Effluent nitrate and nitrite levels during 150 days of operation denitrifying sequencing batch reactor. The reactor was fed with simulated nitrate waste containing 1354, 2710, 4060 and 5420 mg L1 NO3eN at a C/N ratio of 3.
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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FIG. 2. Nitrate and nitrite levels in the effluent during break period and post-break period. Denitrifying granular sludge capable of high strength denitrification of up to 5420 mg L1 at C/N ratio of 3 was achieved. Denitrification was severely inhibited when the reactor was fed with 5420 mg L1 NO1 3 at C/N ratio of 2 due to accumulation of high concentrations of nitrite. A step-down procedure was adopted for decreasing feed nitrate concentration up to 1354 mg L1 for re-establishing high strength denitrification.
In order to improve the denitrification performance, the feed C/ N ratio was reverted to 3. However, high nitrite levels continued to accumulate in the SBR, in spite of surplus electron donor input to the reactor. Oh and Silverstein (25) reported accumulation of significant amount of nitrite during low strength denitrification, when the C/N ratio was decreased from 2 to 1. Complete denitrification of low strength denitrification was restored only 3 weeks after reverting the C/N ratio to 2. In the present study, complete denitrification could not be restored in spite of operating for two weeks at C/N of 3, and the nitrite accumulation continued to be exceptionally high. High strength denitrification cannot be directly compared with low strength denitrification as the nitrite levels reach very high concentrations even in the presence of excess electron donor. TOC measurements showed that the effluent still contained about 2000 mg L1 for feed with 5420 mg L1 nitrate at C/N ratio 2. Some unutilized organic carbon was found in the treated water leaving the reactor at the lowest C/N ratio of 1.2 tested (data not shown). This indicates that even though some electron donor is still available at C/N ratio 2, the nitrite levels formed are relatively much higher due to stiff competition between the nitrate and nitrite for electrons. Thus, it is likely that high nitrite levels inhibited high strength denitrification. To re-establish denitrification, a step-down procedure was adopted and the feed nitrate concentration was decreased to 1354 mg NO3eN L1. At this concentration of nitrate, complete denitrification was established within a few days. Subsequently, the feed nitrate level was increased in a step-wise manner (1354, 2710, 4060 and 5420 mg L1) at C/N ratio 2. After 10 days of operation, stable denitrification could be re-established for all the concentrations of nitrate studied (Fig. 2). At this stage, the microorganisms were acclimatized to high strengths of nitrate as well as nitrite. Therefore, the reactor was fed with various nitrate concentrations, increasing in steps, but at a reduced substrate concentration, viz., C/ N ratio of 2. By this strategy, complete and stable denitrification was achieved at C/N ratio of 2 for feed NO3eN concentrations of 1354, 2710, 4060 and 5420 mg L1 (Figs. 3 and S3). Denitrifying granular sludge During the startup, the SBR was operated with a 10 min settling time, to minimize biomass loss from the system. After one week of operation, the settling time was reduced to 5 min to select granular sludge with good settling characteristics. Formation of tiny, irregularly shaped and well
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FIG. 3. Nitrate and nitrite concentrations in the sequencing batch reactor outlet treating simulated nitrate waste containing 1354, 2710, 4060 and 5420 mg L1 NO3eN at a C/N ratio of 2. Reactor was seeded with denitrifying granular sludge developed on feed with C/N ratio of 3.
settling granules was apparent within two weeks of operation. The morphology of granules, which evolved in the reactor during high strength denitrification, is shown in Fig. 4. The sludge predominantly consisted of granules as evident from visual observations, microscopy and settling characteristics. Long rodshaped microorganisms were evident on the surface of granules (Fig. 4). PCR-DGGE revealed clear shifts in the total bacterial community during the course of reactor operation. Some bands were found to be stable throughout the reactor operation, while some new bands became dominant towards the later period of reactor operation (Fig. 5). Certain bands became more intense than the others, possibly due to the enrichment of specific denitrifying strains. Apart from selection pressure imposed for enrichment, nitrite accumulation during high strength denitrification may strongly influenced development of microbial community, as depicted by the appearance and disappearance of bands. In spite of the break in the denitrification process due to change in C/N ratio (day 153 through 174), the microbial community was found to be resilient to the operational perturbations. Several researchers have reported accumulation of nitrite during denitrification of high strength nitrates (8e12,14). Nitrite accumulation during denitrification by mixed liquor activated sludge was explained on the basis of differences in the rates of reduction of nitrate and nitrite at cellular and population level. Nitrate reduction rates are much higher than nitrite reduction rates, causing accumulation of nitrites. Nitrite and nitrate reductases use protons for reduction reactions from periplasm and cytoplasm, respectively. Thus it results in nitrite accumulation in periplasmic space of denitrifying bacteria. Further, at alkaline pH, concentration of protons in the periplasmic space may be limited, leading to inhibition of nitrite reduction (9). The competition for electrons between nitrate and nitrite reductases can also lead to nitrite accumulation (11). At population level, nitrite accumulation was explained on the basis of differences between nitrate respiring bacteria and true denitrifying bacteria (9,11). Nitrate respiring bacteria can use only nitrate as electron acceptor and nitrite is the end product. In the present study, only 4 out of 14 isolates obtained from the denitrifying granular sludge were able to completely reduce nitrate to N2 gas, while all the remaining isolates could only reduce nitrate to nitrite and no further (data not shown). These results point to the possibility that nitrite accumulation could be also due to
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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FIG. 4. Micrographs showing the morphology and microstructure of denitrifying granular sludge developed in the sequencing batch reactor. (A) Morphology of denitrifying granules collected on day 30 of SBR operation. Bar: 1 mm. (B) Scanning electron microscope (SEM) image showing morphology of microorganisms on the surface of denitrifying granule. Scale bar: 1 mm. Panels C and D are confocal images. (C) maximum intensity projection of xy-images obtained from acridine orange stained denitrifying granule. (D) Maximum intensity projection of an overlay green and red channel of multiple xy-images obtained from denitrifying granule stained with BacLight viability stain. Green, SYTO 9 signal; red, propidium iodide signal. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
differences in the populations of nitrate-respiring and true denitrifying bacteria that co-existed in the granular sludge. Practical implications Transient accumulation of nitrite depends on initial nitrate, electron donor concentrations as well as on microbial community composition. Exposure of microbial community to transient nitrite accumulation is inevitable during high
strength nitrate denitrification in sequencing batch reactors. Since nitrite is inhibitory to microorganisms, enrichment of nitrite tolerant denitrifying bacteria is desirable. A step wise increase in initial nitrate concentration at a fixed C/N ratio allows enrichment of nitrite tolerant microbial community. It is evident that an abrupt decrease in electron donor supply will lead to accumulation of higher nitrite levels to which microbial community has been not
FIG. 5. (A) Ethidium bromide stained DGGE gel; (B) dendrogram prepared using UPGAMA clustering method. DGGE analysis of PCR amplified 16S rRNA gene of bacteria from denitrifying sludge collected from the reactor during 6 months of operation. The reactor was inoculated with activated sludge and fed with simulated nitrate waste. Feed nitrate was increased in a step wise manner to achieve denitrification of 5420 mg L1 NO3eN at C/N ratio of 3. Lane labels refer to day of sample collection after reactor startup.
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
VOL. xx, 2015 exposed thus far and can decrease nitrite reduction and then nitrate reduction by inhibiting denitrifying microorganisms. Therefore, proper electron donor dosing of nitrate contaminated water along with enrichment strategy needs to be adopted in order to protect denitrifying community by avoiding inadvertent exposure to high nitrite levels. Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.jbiosc.2015.05.015 References 1. Smith, V. H.: Eutrophication of freshwater and coastal marine ecosystems: a global problem, Environ. Sci. Pollut. R., 10, 126e139 (2003). 2. Camargo, J. A., Alonso, A., and Salamanca, A.: Nitrate toxicity to aquatic animals: a review with new data for freshwater invertebrates, Chemosphere, 58, 1255e1267 (2005). 3. Fenech, C., Rock, L., Nolan, K., Tobin, J., and Morrissey, A.: The potential for a suite of isotope and chemical markers to differentiate sources of nitrate contamination: a review, Water Res., 46, 2023e2041 (2012). 4. Ward, M. H., deKok, T. M., Levallois, P., Brender, J., Gulis, G., Nolan, B. T., and Van Derslice, J.: Workgroup report: drinking-water nitrate and health-recent findings and research needs, Environ. Health Perspect., 113, 1607e1614 (2005). 5. Smith, V. H. and Schindler, D. W.: Eutrophication science: where do we go from here? Trends Ecol. Evol., 24, 201e207 (2009). 6. Morris, J. M., Fallgren, P. H., and Jin, S.: Enhanced denitrification through microbial and steel fuel-cell generated electron transport, Chem. Eng. J., 153, 37e42 (2009). 7. Kesore, K., Janowski, F., and Shaposhnik, V. A.: Highly effective electrodialysis for selective elimination of nitrates from drinking water, J. Membr. Sci., 127, 17e24 (1997). 8. Glass, C., Silverstein, J., and Oh, J.: Inhibition of denitrification in activated sludge by nitrite, Water Environ. Res., 69, 1086e1093 (1997). 9. Glass, C. and Silverstein, J.: Denitrification kinetics of high nitrate concentration water: pH effect on inhibition and nitrate accumulation, Water Res., 32, 831e839 (1998). 10. Nancharaiah, Y. V. and Venugopalan, V. P.: Denitrification of synthetic concentrated nitrate wastes by aerobic granular sludge under anoxic conditions, Chemosphere, 85, 683e688 (2011). 11. Dhamole, P. B., Nair, R. R., D’Souza, S. F., and Lele, S. S.: Denitrification of high strength nitrate waste, Bioresour. Technol., 98, 247e252 (2007). 12. Dhamole, P. B., Nair, R. R., D’Souza, S. F., Pandit, A. B., and Lele, S. S.: Denitrification of high strength nitrate waste from a nuclear industry using
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acclimatized biomass in a pilot scale reactor, Appl. Biochem. Biotechnol., 175, 748e756 (2014). Foglar, L., Briski, F., Sipos, L., and Vukovic, M.: High nitrate removal from synthetic wastewater with the mixed bacterial culture, Bioresour. Technol., 96, 879e888 (2005). Nair, R. R., Dhamole, P. B., Lele, S. S., and D’Souza, S. F.: Biological denitrification of high strength nitrate waste using preadapted denitrifying sludge, Chemosphere, 67, 1612e1617 (2007). Ge, S., Peng, Y., Wang, S., Lu, C., Cao, X., and Zhu, Y.: Nitrite accumulation under constant temperature in anoxic denitrification process: the effects of carbon sources and COD/NO3eN, Bioresour. Technol., 114, 137e214 (2012). Nancharaiah, Y. V., Schwarzenbeck, N., Mohan, T. V. K., Narasimhan, S. V., Wilderer, P. A., and Venugopalan, V. P.: Biodegradation of nitrilotriacetic acid (NTA) and ferric-NTA complex by aerobic microbial granules, Water Res., 40, 1539e1546 (2006). Nancharaiah, Y. V., Joshi, H. M., Mohan, T. V. K., Venugopalan, V. P., and Narasimhan, S. V.: Aerobic granular biomass: a novel biomaterial for efficient uranium removal, Curr. Sci., 91, 503e509 (2006). Nancharaiah, Y. V., Joshi, H. M., Hausner, M., and Venugopalan, V. P.: Bioaugmentation of aerobic microbial granules with Pseudomonas putida carrying TOL plasmid, Chemosphere, 71, 30e35 (2008). Suja, E., Nancharaiah, Y. V., and Venugopalan, V. P.: p-Nitrophenol biodegradation by aerobic microbial granules, Appl. Biochem. Biotechnol., 167, 1569e1577 (2012). Nancharaiah, Y. V., Kiran Kumar Reddy, G., Lalithamanasa, P., and Venugopalan, V. P.: The ionic liquid 1-alkyl-3-methylimidazolium demonstrates comparable antimicrobial and antibiofilm behavior to a cationic surfactant, Biofouling, 28, 1141e1149 (2012). Nancharaiah, Y. V., Wattiau, P., Wuertz, S., Bathe, S., Mohan, S. V., Wilderer, P. A., and Hausner, M.: Dual labeling of Pseudomonas putida with fluorescent proteins for in situ monitoring of conjugal transfer of the TOL plasmid, Appl. Environ. Microbiol., 69, 4846e4852 (2003). Suja, E., Nancharaiah, Y. V., and Venugopalan, V. P.: Biogenic nanopalladium production by self-immobilized granular biomass: application for contaminant remediation, Water Res., 65, 395e401 (2014). Yarlagadda, V. N., Kadali, R., Sharma, N., Sekar, R., and Vayalam Purath, V.: Rapid establishment of p-nitrophenol biodegradation in acetate-fed aerobic granular sludge, Appl. Biochem. Biotechnol., 166, 1225e1235 (2012). American Public Health Association: Standard methods for the examination of water and wastewater, 19th ed. American Public Health Association, New York, USA (1995). Oh, J. and Silverstein, J.: Acetate limitation and nitrite accumulation during denitrification, J. Environ. Eng., 125, 234e242 (1999).
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors Tulasi Venkata Krishna Mohan,1, * Kadali Renu,1, 2 Yarlagadda Venkata Nancharaiah,1 Pedapati Murali Satya Sai,3 and Vayalam Purath Venugopalan1 Water and Steam Chemistry Division, Bhabha Atomic Research Centre, Kalpakkam 603102, India,1 School of Biosciences and Technology, VIT University, Vellore 632014, India,2 and Waste Immobilization Plant, NRB, Bhabha Atomic Research Centre, Kalpakkam 603102, India3 Received 22 April 2015; accepted 25 May 2015 Available online xxx
A 6-L sequencing batch reactor (SBR) was operated for development of granular sludge capable of denitrification of high strength nitrates. Complete and stable denitrification of up to 5420 mg LL1 nitrate-N (2710 mg LL1 nitrate-N in reactor) was achieved by feeding simulated nitrate waste at a C/N ratio of 3. Compact and dense denitrifying granular sludge with relatively stable microbial community was developed during reactor operation. Accumulation of large amounts of nitrite due to incomplete denitrification occurred when the SBR was fed with 5420 mg LL1 NO3eN at a C/N ratio of 2. Complete denitrification could not be achieved at this C/N ratio, even after one week of reactor operation as the nitrite levels continued to accumulate. In order to improve denitrification performance, the reactor was fed with nitrate concentrations of 1354 mg LL1, while keeping C/N ratio at 2. Subsequently, nitrate concentration in the feed was increased in a step-wise manner to establish complete denitrification of 5420 mg LL1 NO3eN at a C/N ratio of 2. The results show that substrate concentration plays an important role in denitrification of high strength nitrate by influencing nitrite accumulation. Complete denitrification of high strength nitrates can be achieved at lower substrate concentrations, by an appropriate acclimatization strategy. Ó 2015, The Society for Biotechnology, Japan. All rights reserved. [Key words: Carbon to nitrogen ratio; Denitrification; Granular sludge; High strength nitrate; Nitrite accumulation]
Nitrate is one of the most widespread contaminants of natural water resources. During the past many decades, anthropogenic activities have greatly contributed to the increase in nitrate concentrations in surface and groundwater (1e3). Nitrate contamination of aquatic environments is a concern as it has been linked to ecosystem well-being and public health (4). For example, nitrate contamination causes eutrophication of water bodies and bluebaby syndrome in infants (2,5). Due to potential risks on ecosystem and public health, nitrate is listed as a priority pollutant under US Environmental Protection Agency (6). The maximum contaminant level (MCL) for nitrate (NO-3) in drinking water supplies has been set at 45 mg L1 by US EPA, WHO and Bureau of Indian Standards. The European Union has prescribed 50 mg L1 as the limit for nitrate and 0.1 mg L1 for nitrite in drinking water (7). Two main point sources for nitrate contamination are sewage and industrial effluents. High strength nitrate effluents are generated in fertilizer, ammunition, pharmaceutical, metal finishing and nuclear industries (8e10). In nuclear industry, nitric acid is used in various processing steps of nuclear fuel fabrication, dissolution and reprocessing of spent nuclear fuel, resulting in production of low and high strength nitrate-bearing effluents (8,11). The acidic nitrate effluents thus generated are neutralized with hydroxide and stored under safe conditions for treatment and disposal in future.
Biological denitrification for treatment of high strength nitrate effluents has been widely studied, particularly in the case of wastes generated in nuclear fuel cycle operations. It is acknowledged that biological treatment of high strength nitrate wastes typical for industrial effluents is challenging (8,9,11e14). One of the problems commonly encountered during denitrification of high strength nitrates is accumulation of high levels of nitrite (8,11,12,14). Accumulation of high concentrations of nitrite can lead to inhibition of the whole denitrification process (8). Two main factors that could influence nitrite accumulation during denitrification are electron donor supply and composition of the microbial community (15). Nitrate effluents generated in the nuclear industry are often devoid of electron donors. Dosage of electron donor in terms of C/N ratio has been studied for minimizing nitrite accumulation and achieving complete denitrification (12). Regarding the microbial community, it has been reported that biomass in the form of granular sludge may be better for overcoming the inhibitory effects of nitrite concentrations encountered during high strength nitrate denitrification (10). The objectives of this study, therefore, were (i) cultivation of granular sludge for the denitrification of high strength nitrate wastewater and (ii) understanding the effect of C/N ratio on the denitrification process.
MATERIALS AND METHODS * Corresponding author. Tel.: þ91 44 27480203; fax: þ91 44 27480097. E-mail address: [email protected] (T.V. Krishna Mohan).
Seed sludge Activated sludge collected from the outlet of an aeration tank of an operating municipal wastewater treatment plant at Kalpakkam, India was used as
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Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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inoculum for the sequencing batch reactors (SBRs). The activated sludge was black in color, contained filamentous microorganisms and exhibited poor settling characteristics (16,17). After bringing to the laboratory, the sludge was washed a few times with deionized water and stored at 4 C. Prior to use, the activated sludge was washed with ultrapure water and was placed on a tissue paper to remove excess water. The wet sludge was weighed and transferred to the SBR. Cultivation of denitrifying granular biomass A glass tank (height: 61 cm, diameter: 15 cm) with 6 L working volume was operated in a sequencing batch mode. SBR operation was chosen as it allows formation and selection of granular sludge. The SBR was inoculated with activated sludge (3 g mixed liquor suspended solids L1) and operated with 24 h cycle period. The SBR cycle consisted of the following: 5 min fill, 23 h reaction, 5 min settle, 10 min decant and 40 min idle. The SBR was fed with simulated nitrate waste prepared in deionized water and consisting of the following ingredients (g L1): MgSO4.7H2O 0.04, KCl 0.02, K2HPO4 0.03, KH2PO4 0.01, CaCl2.2H2O 0.02. Trace elements were supplied by adding 0.1 mL of trace element mix per 1 L of simulated nitrate waste (18). Sodium acetate and sodium nitrate were added to the simulated nitrate waste, as per requirement (Table 1). The SBR was operated at room temperature (w30 C) with 50% volumetric exchange ratio. Mixing was provided at a fixed speed of 100 rpm using a stirrer (Eltek, India). Effluent was drawn from a port located 17 cm above the bottom. The SBR was operated for more than 8 months in order to monitor the stability of denitrification process. Microscopy of denitrifying sludge Morphology of the denitrifying sludge was documented using a DP70 camera (Olympus, Japan) attached to SMZ1000 stereozoom microscope (Nikon, Japan) (16). The denitrifying granular sludge was stained with BacLight bacterial viability kit (Molecular Probes, USA) (19). A 200 mL of BacLight stain mixture (SYTO 9 and propidium iodide) was added to 1.5 mL Eppendorf tube containing the sludge and incubated on shaker set at 100 rpm. After 15 min, the sludge was washed twice with ultrapure water. The stained granules were imaged using a confocal laser scanning microscope TCS SP2 AOBS (Leica Microsystems, Germany) attached to an inverted microscope (Leica DMIRE2). Images were collected using a 63x 1.2 NA water immersion objective (20). Sample was excited using a 488 nm Ar laser line. The fluorescence emission was collected between 500 and 520 nm for SYTO 9 and between 600 and 680 nm for propidium iodide. DNA extraction and PCR-DGGE Activated sludge and denitrifying granular sludge (0.2 g wet weight) collected from the SBR were homogenized by vortexing with glass beads (21). Genomic DNA from the homogenized samples was extracted using DNA isolation kit (Qiagen, Germany). The extracted DNA was verified by 1% (w/ v) agarose gel electrophoresis. The extracted genomic DNA was stored at 50 C. The V3 region of the bacterial 16S rRNA gene was amplified using primers PRBA338f (50 CGCCCGCCGCGCGCGGCGGGCGGGGC GGGGGCACGGGGGGACTCCTACGGGAGGCA GCAG 30 ) and PRUN518r (50 ATTACCGCGGCTGCTGG 30 ) (19). Polymerase chain reaction (PCR) was performed using 50 mL reaction mixture using a Mastercycler gradient thermal cycler (Eppendorf AG, Germany) as described earlier (19). The PCR amplified fragments were separated by denaturing gradient gel electrophoresis (DGGE) using an INGENYphorU system (Ingeny International BV, The Netherlands). The DGGE was performed using 40%e70% urea-formamide denaturant gradient gel [8% (w/v) acrylamide solution (40% acrylamide and bisacrylamide in 37.5:1 ratio)] in TAE buffer (40 mM Tris base, 20 mM acetic acid, 1 mM EDTA, pH 8.0) as described earlier (20,22,23). After electrophoresis, the gel was stained with ethidium bromide (0.1 mg L1) for 30 min and destained for 15 min in ultrapure water. The stained gel was documented with INFINITY gel documentation system (Vilber Lourmat, France) and diversity indices were determined as Ward’s Dice similarity coefficients with the Quantity One version 4.6 (Bio-Rad, USA).
RESULTS AND DISCUSSION Denitrification performance The SBR was inoculated with activated sludge and fed with simulated waste containing 1354 mg NO3eN L1 at a C/N ratio of 3. It may be noted that the starting concentration in the reactor would be 677 mg NO3eN L1, because the SBR retains 50% of waste from the previous batch (50% VER), with no nitrates in it. The pH of simulated nitrate waste was 7.5 before feeding to the reactor. The pH observed in the reactor at the end of cycle was 9.5, the increase being due to denitrification. In subsequent cycles of operation, the pH in the reactor stabilized at w9.5 (Fig. S1). The simulated waste prepared in deionized water and the dissolved oxygen (DO) content before feeding was noted to be 6.0 mg L1. The DO rapidly decreased to below 0.08 mg L1 within a few minutes of addition to the reactor, due to aerobic microbial respiration. The reactor tank was open to atmosphere and the mixing was provided by means of bottom stirring at 100 rpm. Under these conditions, the DO in the SBR cycle period was always below 0.08 mg L1. The reactor operating conditions were conducive for heterotrophic denitrification, observed from cycle 1 onwards (Fig. 1). By keeping the C/N ratio constant at 3, feed nitrate-N was increased to 2710 and then to 5420 mg L1. Complete denitrification was observed at the increased feed nitrate concentrations of 2710 and 5420 mg NO3eN L1. The effluent nitrate and nitrite concentrations were invariably less than 10 mg L1 during long term operation of the SBR, except during a few occasions (Fig. 1). Denitrification of feed NO3eN concentrations of 1354, 2710 and 5420 mg L1 was completed within the first 4, 6 and 8 h, respectively (Fig. S2). Typical denitrification profiles consisted of nitrate removal, accumulation of nitrite and nitrite removal. Nitrite accumulation reached a maximum value of approximately 250 mg L1 NO2eN at 1354 mg L1 nitrate-N in the feed. In order to study the denitrification at reduced substrate concentration, the reactor was subsequently fed with 5420 mg L1 NO3eN at a C/N ratio of 2. The change in C/N ratio caused inhibition of denitrification, resulting in incomplete denitrification and accumulation of nitrite-N as high as 3500 mg L1 (Fig. 2). SBR operation in the subsequent cycles consistently showed the buildup of the high concentrations of nitrite and nitrate. Inhibition of the denitrification process was probably due to the toxicity caused by high nitrite levels.
Analytical procedures Liquid samples collected at regular time intervals during the SBR cycle period were monitored for nitrite and nitrate. Nitrate was analyzed by high performance liquid chromatography (HPLC) (Dionex Ultimate 3000) fitted with an Acclaim OA column using a UVevis detector set at 210 nm (10). Dilute H2SO4 (0.003 N) was prepared in deionized water and used at a flow rate of 0.7 mL min1. Nitrite was estimated spectrophotometrically by reaction with N-(1naphthyl) ethylenediamine dihydrochloride (24). Mixed liquor suspended solids (MLSS) and total organic carbon (TOC) were occasionally analyzed as per standard methods (24).
TABLE 1. The simulated nitrate waste was prepared by varying the amounts of sodium acetate and sodium nitrate. Constituent (g L1)
NaNO3 CH3COONa (3:1)a CH3COONa (2:1)a a
Feed NO3eN (mg L1) 1354
2710
4060
5420
8.22 13.89 9.26
16.43 27.77 18.52
24.86 41.66 27.77
32.86 55.55 37.03
Sodium acetate was added to give a C/N mass ratio of 3 or 2, i.e., 3 or 2 g of carbon for g of nitrogen.
FIG. 1. Effluent nitrate and nitrite levels during 150 days of operation denitrifying sequencing batch reactor. The reactor was fed with simulated nitrate waste containing 1354, 2710, 4060 and 5420 mg L1 NO3eN at a C/N ratio of 3.
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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FIG. 2. Nitrate and nitrite levels in the effluent during break period and post-break period. Denitrifying granular sludge capable of high strength denitrification of up to 5420 mg L1 at C/N ratio of 3 was achieved. Denitrification was severely inhibited when the reactor was fed with 5420 mg L1 NO1 3 at C/N ratio of 2 due to accumulation of high concentrations of nitrite. A step-down procedure was adopted for decreasing feed nitrate concentration up to 1354 mg L1 for re-establishing high strength denitrification.
In order to improve the denitrification performance, the feed C/ N ratio was reverted to 3. However, high nitrite levels continued to accumulate in the SBR, in spite of surplus electron donor input to the reactor. Oh and Silverstein (25) reported accumulation of significant amount of nitrite during low strength denitrification, when the C/N ratio was decreased from 2 to 1. Complete denitrification of low strength denitrification was restored only 3 weeks after reverting the C/N ratio to 2. In the present study, complete denitrification could not be restored in spite of operating for two weeks at C/N of 3, and the nitrite accumulation continued to be exceptionally high. High strength denitrification cannot be directly compared with low strength denitrification as the nitrite levels reach very high concentrations even in the presence of excess electron donor. TOC measurements showed that the effluent still contained about 2000 mg L1 for feed with 5420 mg L1 nitrate at C/N ratio 2. Some unutilized organic carbon was found in the treated water leaving the reactor at the lowest C/N ratio of 1.2 tested (data not shown). This indicates that even though some electron donor is still available at C/N ratio 2, the nitrite levels formed are relatively much higher due to stiff competition between the nitrate and nitrite for electrons. Thus, it is likely that high nitrite levels inhibited high strength denitrification. To re-establish denitrification, a step-down procedure was adopted and the feed nitrate concentration was decreased to 1354 mg NO3eN L1. At this concentration of nitrate, complete denitrification was established within a few days. Subsequently, the feed nitrate level was increased in a step-wise manner (1354, 2710, 4060 and 5420 mg L1) at C/N ratio 2. After 10 days of operation, stable denitrification could be re-established for all the concentrations of nitrate studied (Fig. 2). At this stage, the microorganisms were acclimatized to high strengths of nitrate as well as nitrite. Therefore, the reactor was fed with various nitrate concentrations, increasing in steps, but at a reduced substrate concentration, viz., C/ N ratio of 2. By this strategy, complete and stable denitrification was achieved at C/N ratio of 2 for feed NO3eN concentrations of 1354, 2710, 4060 and 5420 mg L1 (Figs. 3 and S3). Denitrifying granular sludge During the startup, the SBR was operated with a 10 min settling time, to minimize biomass loss from the system. After one week of operation, the settling time was reduced to 5 min to select granular sludge with good settling characteristics. Formation of tiny, irregularly shaped and well
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FIG. 3. Nitrate and nitrite concentrations in the sequencing batch reactor outlet treating simulated nitrate waste containing 1354, 2710, 4060 and 5420 mg L1 NO3eN at a C/N ratio of 2. Reactor was seeded with denitrifying granular sludge developed on feed with C/N ratio of 3.
settling granules was apparent within two weeks of operation. The morphology of granules, which evolved in the reactor during high strength denitrification, is shown in Fig. 4. The sludge predominantly consisted of granules as evident from visual observations, microscopy and settling characteristics. Long rodshaped microorganisms were evident on the surface of granules (Fig. 4). PCR-DGGE revealed clear shifts in the total bacterial community during the course of reactor operation. Some bands were found to be stable throughout the reactor operation, while some new bands became dominant towards the later period of reactor operation (Fig. 5). Certain bands became more intense than the others, possibly due to the enrichment of specific denitrifying strains. Apart from selection pressure imposed for enrichment, nitrite accumulation during high strength denitrification may strongly influenced development of microbial community, as depicted by the appearance and disappearance of bands. In spite of the break in the denitrification process due to change in C/N ratio (day 153 through 174), the microbial community was found to be resilient to the operational perturbations. Several researchers have reported accumulation of nitrite during denitrification of high strength nitrates (8e12,14). Nitrite accumulation during denitrification by mixed liquor activated sludge was explained on the basis of differences in the rates of reduction of nitrate and nitrite at cellular and population level. Nitrate reduction rates are much higher than nitrite reduction rates, causing accumulation of nitrites. Nitrite and nitrate reductases use protons for reduction reactions from periplasm and cytoplasm, respectively. Thus it results in nitrite accumulation in periplasmic space of denitrifying bacteria. Further, at alkaline pH, concentration of protons in the periplasmic space may be limited, leading to inhibition of nitrite reduction (9). The competition for electrons between nitrate and nitrite reductases can also lead to nitrite accumulation (11). At population level, nitrite accumulation was explained on the basis of differences between nitrate respiring bacteria and true denitrifying bacteria (9,11). Nitrate respiring bacteria can use only nitrate as electron acceptor and nitrite is the end product. In the present study, only 4 out of 14 isolates obtained from the denitrifying granular sludge were able to completely reduce nitrate to N2 gas, while all the remaining isolates could only reduce nitrate to nitrite and no further (data not shown). These results point to the possibility that nitrite accumulation could be also due to
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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FIG. 4. Micrographs showing the morphology and microstructure of denitrifying granular sludge developed in the sequencing batch reactor. (A) Morphology of denitrifying granules collected on day 30 of SBR operation. Bar: 1 mm. (B) Scanning electron microscope (SEM) image showing morphology of microorganisms on the surface of denitrifying granule. Scale bar: 1 mm. Panels C and D are confocal images. (C) maximum intensity projection of xy-images obtained from acridine orange stained denitrifying granule. (D) Maximum intensity projection of an overlay green and red channel of multiple xy-images obtained from denitrifying granule stained with BacLight viability stain. Green, SYTO 9 signal; red, propidium iodide signal. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
differences in the populations of nitrate-respiring and true denitrifying bacteria that co-existed in the granular sludge. Practical implications Transient accumulation of nitrite depends on initial nitrate, electron donor concentrations as well as on microbial community composition. Exposure of microbial community to transient nitrite accumulation is inevitable during high
strength nitrate denitrification in sequencing batch reactors. Since nitrite is inhibitory to microorganisms, enrichment of nitrite tolerant denitrifying bacteria is desirable. A step wise increase in initial nitrate concentration at a fixed C/N ratio allows enrichment of nitrite tolerant microbial community. It is evident that an abrupt decrease in electron donor supply will lead to accumulation of higher nitrite levels to which microbial community has been not
FIG. 5. (A) Ethidium bromide stained DGGE gel; (B) dendrogram prepared using UPGAMA clustering method. DGGE analysis of PCR amplified 16S rRNA gene of bacteria from denitrifying sludge collected from the reactor during 6 months of operation. The reactor was inoculated with activated sludge and fed with simulated nitrate waste. Feed nitrate was increased in a step wise manner to achieve denitrification of 5420 mg L1 NO3eN at C/N ratio of 3. Lane labels refer to day of sample collection after reactor startup.
Please cite this article in press as: Krishna Mohan, T. V., et al., Nitrate removal from high strength nitrate-bearing wastes in granular sludge sequencing batch reactors, J. Biosci. Bioeng., (2015), http://dx.doi.org/10.1016/j.jbiosc.2015.05.015
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