Bioprocess Biosyst Eng DOI 10.1007/s00449-017-1844-5
RESEARCH PAPER
Salinity effect on simultaneous nitrification and denitrification, microbial characteristics in a hybrid sequencing batch biofilm reactor Zonglian She1,2,3 • Lan Wu1 • Qun Wang1 • Mengchun Gao1,2 • Chunji Jin1,2 Yangguo Zhao1,2 • Linting Zhao1 • Liang Guo1,2
•
Received: 1 June 2017 / Accepted: 21 September 2017 Springer-Verlag GmbH Germany 2017
Abstract The effect of increasing salinity on nitrogen removal via simultaneous nitrification and denitrification, microbial activities and extracellular polymeric substances (EPS) were investigated in a hybrid sequencing batch biofilm reactor filled with soft combination carriers. In the influent salinity range from 1.0 to 2.0%, average COD, NH4?-N and TN removal efficiencies were higher than 97.1, 97.8 and 86.4% at the steady state. When salinity was increased to 2.5 and 3.0%, ammonium oxidation was obviously inhibited in the reactor. For both suspended sludge (S-sludge) and biofilm, specific ammonium oxidation rate, specific nitrite oxidation rate, specific oxygen uptake rate and dehydrogenase activity reduced with the increase of salinity. The quantity of total EPS increased with the increase of salinity from 1.0 to 2.0%. Generally, humic substances were the dominant composition of EPS in both S-sludge and biofilm, with the percentages of 43.9–54.0 and 43.8–64.6% in total EPS. Keywords High salinity Simultaneous nitrification and denitrification (SND) Microbial activity Extracellular polymeric substances Humic substances
& Liang Guo [email protected] 1
Key Lab of Marine Environment and Ecology, Ocean University of China, Ministry of Education, Qingdao 266100, China
2
College of Environmental Science and Engineering, Ocean University of China, Qingdao 266100, China
3
Shandong Provincial Key Laboratory of Marine Environment and Geological Engineering, Ocean University of China, Qingdao 266100, China
Introduction Wastewaters with high salinity can be generated by chemical, pharmaceutical, petroleum, fish caning, seafood processing, meat packing, tannery, and cheese factories. The use of seawater for toilet flushing can also increase salt levels in sewage [1]. In a biological system for wastewater treatment, high salinity may decrease the cell activity, and even cause cell plasmolysis [2]. In addition, high salt concentration would affect the structure and settling characteristics of sludge flocs, production and composition of extracellular polymeric substances (EPS) of microorganism [3, 4]. Simultaneous nitrification and denitrification (SND) process has several advantages over nitrogen removal through conventional two-step nitrification and denitrification: (a) simplifying the operating procedures, (b) reducing the oxygen requirement and energy consumption [5]. In a SND system, nitrification and denitrification processes occur at the same time in a single reactor. High nitrogen removal efficiency via SND has been reported in both sequencing batch reactor (SBR) and sequencing batch biofilm reactor (SBBR) [6, 7]. Hybrid sequencing batch biofilm reactor (HSBBR) combines advantages of activated sludge, biofilm system and sequencing batch reactor. In a HSBBR, the interaction of the biofilm and S-sludge resulted in a better overall nitrogen removal performance via SND [8]. During aeration phase, nitrification could occur in S-sludge and the outer layer of biofilm where the oxygen concentration was high enough, whereas denitrification could occur in the inner layer of biofilm in which oxygen level was low. Wang et al. [9] reported that high efficiency of SND via nitrite was successfully achieved under limited dissolved oxygen (DO) concentration in a sequence hybrid biological reactor with fixed media.
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To date, a limited number of studies are available in the literature concerning the influences of salinity on the SND process. Corsino et al. [10] reported that simultaneous nitritation and denitritation was successfully sustained at the salinity up to 50 g NaCl L-1 in the treatment of fish canning wastewater by aerobic granular sludge, with a total nitrogen removal efficiency of over 90%. Chen et al. [2] found that under the salinity up to 3.0%, organic matter and nitrogen removal can be achieved simultaneously in a single-stage pressurized biofilm reactor, achieving TN removal efficiency via SND up to 98%. Although the salinity effect on the production of EPS has been studied in a moving bed SBR system based on SND process [4], the influence of high salinity on the performance of nitrogen removal via SND and the microbial characteristics of S-sludge and biofilm in HSBBR are still unclear. The main objectives of this study are (1) to investigate the effect of the increase of influent salinity on the removal of COD and nitrogen through SND in an aerobic HSBBR, (2) to analyze the variations of the specific ammonium oxidation rate (SAOR), specific nitrite oxidation rate (SNIOR), specific nitrate reduction rate (SNARR), specific nitrite reduction rate (SNIRR), specific oxygen uptake rate (SOUR) and dehydrogenase activity (DHA) for S-sludge and biofilm at different salinities, (3) to evaluate the effect of salinity on the EPS production and composition from S-sludge and biofilm.
Materials and methods The HSBBR system and operating conditions A lab-scale aerobic HSBBR (Fig. 1) which contained both S-sludge and attached biofilm was operated for 198 days in
this study. The working volume of HSBBR was 13.3 L. Four soft combination carriers (made of polyvinyl formal fiber and polypropylene hoop) were installed in the reactor, and the space between two carriers was 10 cm. Each carrier had a diameter of 14 cm and a specific surface area of 1236 m2 m-3. The photo of a combined carrier is shown in Fig. 1. The reactor was run with a cycle of 8 h. Each cycle was comprised of five phases: feeding (30 min), aeration (390 min), settling (40 min), decanting (10 min) and idle (10 min). 4.9 L of influent was fed into the reactor during the feeding phase and the same volume of effluent was discharged after the end of the settling phase, resulting in a volumetric exchanging ratio of 36.8% and a hydraulic residence time of 21.7 h. The reactor was operated at fixed temperature of 25.0 ± 1.5 C. The influent pH values were controlled at 7.50–8.02. Oxygen was supplied through two air diffuser stones in the reactor and a flow meter was used to control the airflow rate. The reactor was mixed during the feeding and aeration phases by using a submersible pump placed in the reactor to recirculate liquid. Inflow, outflow, air supply and mixing were controlled by time controllers. Before this experiment commenced, the HSBBR had been operated at low salinity (0.15–1.0%) for 440 days with glucose as carbon source in the influent and SND process had been achieved successfully. In this study, the reactor was fed with synthetic saline wastewater added with sodium acetate (NaAc) as carbon source. The wastewater was prepared with tap water and the composition was as follows (per liter): 0.35 g NaAc, 0.15 g NH4Cl, 0.04 g KH2PO4, 0.2 g NaHCO3 and seawater crystal. The components of seawater crystal solution (1.0% salinity) was as follows (mg L-1): Na? 3266; Cl- 6008; Mg2? 317; SO42- 830; K? 120; Ca2? 100; Zn2? 0.005; Mn2? 0.004; Fe2? 0.04; Co2? 1 9 10-4; Mo6? 1 9 10-3; I- 0.023; Sr? 2.5910-3; Se6? 1.17 9 10-4. Influent salinity was gradually increased from 1.0 to 1.5, 2.0, 2.5 and 3.0%. These five salinity conditions were maintained for 59, 10, 50, 29 and 50 days, respectively. Analytic methods
Fig. 1 Schematic diagram of the HSBBR system and the photo of carrier
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Water samples for the analysis of chemical oxygen demand (COD), ammonium nitrogen (NH4?-N), nitrate nitrogen (NO3--N) and nitrite nitrogen (NO2--N) were centrifuged for 10 min at 4000 rpm, then measured according to standard methods [11]. Total nitrogen (TN) was calculated by the sum of NH4?-N, NO3--N and NO2--N. Total suspended solids (TSS) and volatile suspended solids (VSS) were determined also in accordance with standard methods [11]. Temperature and DO were measured using a DO detector (Oxi 330i, WTW, Germany). pH value was determined by pH probes (PHB-4, China).
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The simultaneous nitrification and denitrification (SND) efficiency was given in Eq. (1). þ SND efficiency ¼ ½1 (NO x;e -N NOx;i -N)=ðNH4;i -N NHþ 4;e -NÞ 100%;
ð1Þ -1 where NOx,e-N (mg L ) is nitrogen in nitrite and nitrate at -1 the end of the aeration phase, NOx,i-N (mg L ) is nitrogen in nitrite and nitrate at the beginning of the aeration phase, -1 NH? 4,i-N (mg L ) is ammonium nitrogen at the beginning -1 of the aeration phase, NH? 4,e-N (mg L ) is ammonium nitrogen at the end of the aeration phase.
Activity tests Batch tests were conducted to evaluate activity of S-sludge and biofilm from the HSBBR. All samples were tested in triplicate. SAOR, SNIOR, SNARR and SNIRR were measured in conical flasks at 25 C while shaking at 150 rpm in a thermostatic water bath oscillators. Salinity in each test was identical with that of the synthetic wastewater fed into the HSBBR. For the measurement of SAOR, NH4Cl was used as nitrogen source to attain approximately 20 mg L-1 of initial nitrogen concentration in bottles. For the test of SNIOR, SNARR and SNIRR, NaNO2, NaNO3 and NaNO2 were added to flasks, respectively, to achieve 10 mg L-1 of initial nitrogen concentration. To determine SNARR and SNIRR, acetate was added as external carbon source to attain an initial COD concentration of 120 mg L-1. For the detection of endogenous SNARR (E-SNARR) and SNIRR (E-SNIRR), no carbon was added. Further details on the tests of nitrifier and denitrifier activities were described in previous study [12]. Specific oxygen uptake rate (SOUR) was determined according to Panswad and Anan [13]. S-sludge or biofilm taken from the HSBBR was mixed with fed wastewater and aerated to oxygen saturation in conical flasks. Then the flasks were sealed. The tests were performed at 25 C while stirring with a magnetic stirrer. DO in each flask was measured by a DO detector. SOUR was calculated from the slope of the DO depletion versus time divided by the VSS in the flask. Dehydrogenase activity (DHA) was tested following the modified method proposed by Klapwuket al. [14]. In this method, DHA is measured based on the reduction of triphenyl tetrazolium chloride (TTC) to triphenyl formazan (TF). The DHA was calculated according to the TF calibration curve and expressed as mg TF produced per g biomass (VSS) and per hour. EPS extraction and analysis Both S-sludge and biofilm samples were collected from the HSBBR for EPS analysis. The extracted soluble EPS (S-
EPS), loosely bound EPS (LB-EPS) and tightly bound EPS (TB-EPS) were analyzed for polysaccharides (PS), proteins (PN), and humic substances (HS). The total EPS content was taken as the sum of S-EPS, LB-EPS and TB-EPS. The extraction and measurement of EPS were performed according to the method described by Zhao et al. [4].
Results and discussion Performance of the HSBBR Figure 2 shows the overall COD, NH4?-N, and TN removal performance during the whole experiment. The excellent COD removal was attained under the five salinities where the removal rate was 95.6–99.6% at steady state (Fig. 2a). This result indicated that the heterotrophic bacteria worked effectively and the increase of salinity had no detrimental effect on the biodegradation of organics in the HSBBR. The ammonium oxidation and SND occurred successfully at salinity lower than 2.0%, with more than 89.9 and 84.4% of NH4?-N and TN removed from HSBBR, respectively (Fig. 2b, e). Similar NH4?-N removal was reported by Bassin et al. [1] whose study showed that good nitrification performance could be obtained at the salinity up to 20 g NaCl L-1. With the increase of salinity to 2.5 and 3.0%, higher effluent NH4?-N concentration was observed, average NH4?-N removal efficiencies dropped dramatically to 57.9 and 37.8%, respectively. These results indicated that autotrophic bacteria for ammonium oxidization were inhibited at higher salt concentrations. Overall, the concentration of NO3--N in the effluent was higher than that of NO2--N at 1.0 and 1.5% salinity, whereas more NO2--N in the outflow was detected when the salinity was increased to 2.0–3.0% (Fig. 2d). Nitrogen removal characteristics in the reactor To gain a better insight into the conversion of nitrogen and COD in the HSBBR, the changes of nitrogen, COD, pH and DO during aeration phase were investigated when the reactor was operated at steady state under 1.0, 2.0 and 3.0% salinities (Fig. 3). As shown in Fig. 3g, most of COD was quickly removed from bulk liquid within the first 0.5 h at 1.0% salinity and 1.5 h at 2.0% salinity, respectively; then COD concentration varied slightly during the late aeration stage. The quick decrease of COD in the early stage of aeration could be attributed to both the absorption and degradation via the S-sludge and biofilm. Some researchers [15, 16] reported that organic matter could be adsorbed rapidly by the
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Fig. 2 Influent and effluent concentration, removal efficiency of COD and nitrogen under different salinities
biofilm or aerobic granular sludge because of the existence of heterotrophic bacteria. For 3.0% salinity, fewer COD was removed in the initial stage of aeration due to the high salinity stress on heterotrophic bacteria. This study has proved that high salinity inhibited the activity of heterotrophic bacteria (As shown in Fig. 4g, h), which could reduce the absorption and utilization of COD by microorganism.
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At 1.0% salinity, ammonium was rapidly oxidized within the initial 2.0 h of the aeration phase with the decrease of NH4?-N concentration from 17.3 to 2.0 mg L-1, obtaining high ammonium oxidation rate (AOR) of 7.86 mg N L-1 h-1 in the reactor (Fig. 3a). The NH4?-N profile had a turning at point A (aerated for 2.0 h) when ammonium was almost completely transformed. TN concentration dropped from 19.0 to 6.7 mg L-1 during the
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Fig. 3 Variations of COD, nitrogen, DO and pH during aeration phase under various salinities (a ammonium turning point, b pH peak point, c pH valley point, d DO breakpoint)
aeration phase, obtaining a SND efficiency of 74.9%. This result implied that the co-existence of nitrifying bacteria and denitrifying bacteria in the hybrid biological reactor made TN removal via SND possible.
At 2.0% salinity, the AOR obtained from the reactor was 4.08 mg N L-1 h-1, which is about half of that at 1.0% salinity, and the nitrite accumulation was slower compared with that at 1.0% salinity (Fig. 3c). Under 2.0% salinity,
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Bioprocess Biosyst Eng Fig. 4 Microbial activities for the S-sludge and biofilm at different salinities
there was almost no NO3--N accumulation and NO3--N level kept below 0.6 mg L-1 during the whole aerobic period, implying that nitrite oxidation was almost completely inhibited. The nitrite accumulation could be observed at 2.0% salinity during the aerobic period, achieving a maximum NO2--N concentration of 5.07 mg L-1 at 4.5 h. This was in accordance with the previous study by Corsino et al. [10] who reported that high salt concentrations lead to the accumulation of nitrite due to the inhibition of nitrite-oxidizing bacteria (NOB). Windey et al. [17] also reported that the halophilic condition in a rotating biological contactor (RBC) was favorite
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to nitritation, since NOB is more negatively affected by high salt concentrations. Furthermore, the continuous decline of TN concentration during the aeration process indicated that denitritation occurred during the whole aeration phase. Based on the much more accumulation of nitrite compared to nitrate and the complete inhibition of NOB activity at 2.0% salinity as shown in Fig. 4b, it could be concluded that short-cut simultaneous nitrification and denitrification via nitrite (i.e., simultaneous nitritation and denitritation) was the predominate process for TN removal in the HSBBR. It is also noticeable that the process of the complete nitrification and denitrification could not be
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eliminated in terms of the existence of nitrate. The presence of an anoxic microzone in the biofilm could be an explanation for the high SND efficiency (90.1%) attained at this salinity, as microzone was favorable to the development of SND [18]. As shown in Fig. 3b and d, during the first 0.5 h for 1.0% salinity and 1.0 h at 2.0% salinity (aeration from the beginning to peak point B on pH curves), pH value increased rapidly, which was perhaps due to the sufficient alkalinity in feed and/or the production of alkalinity via denitrification of nitrate and/or nitrite (remained from previous cycle or carried by sludge and biofilm). After that, pH declined quickly due to the consumption of alkalinity accompanied by the rapid oxidation of ammonium. On the whole, DO level increased continuously during the aeration phase, since the oxygen amount supplied by aeration was higher than that required for the aerobic oxidation of ammonium and COD. Holman and Wareham [19] demonstrated that the depletion of COD and NH4?-N caused a sudden DO increase, which was also observed in the present study at 1.0 and 2.0% salinities. When aeration has been lasted for nearly 2.0 and 4.0 h, the DO profiles had breakpoint at point D (Fig. 3b, d). After this point DO appeared a sudden increase, which could be attributed to the decrease of oxygen demand after ammonium had been almost completely removed. It is notable that when DO reached the breakpoint D, pH got to the valley point C almost at the same time. Therefore, both the pH valley point and DO break point could be used to indicate the end of rapid ammonium oxidation. For 3.0% salinity, the serious inhibition of high salt stress on ammonium oxidation led to high ammonium residue (21.1 mg L-1), and almost no nitrite and nitrate accumulation at the end of aeration (Fig. 3e). Microbial activity at different salinities The effect of salinity on the activities of ammonium oxidizers from the S-sludge and biofilm was similar (Fig. 4a). The increase of salinity led to lower SAOR. At 3.0% salinity, the SAOR for the S-sludge and biofilm declined by 69.3 and 82.2%, respectively, compared with 1.0% salinity. This result explained why NH4?-N accumulation occurred in the HSBBR at 3.0% salinity, as shown in Fig. 2b and Fig. 3e. At 2.0 and 3.0% salinities, the SNIOR of both S-sludge and biofilm was not detected, indicating that the activity of nitrite oxidizers was completely inhibited. No accumulation of nitrate in the reactor at 2.0 and 3.0% salinity (Fig. 3c, e) also demonstrated that nitrite oxidation was completely restrained in the HSBBR. The decrease in the activity of ammonium and nitrite oxidizers found in this study agrees with previous studies [1, 20]. It is worth noting that at 1.0% salinity, the SAOR for both S-sludge
and biofilm were obviously higher than the SNIOR (Fig. 4a, b), implying that nitrite oxidizers were more sensitive to salt stress than ammonium oxidizers. The SAOR of the S-sludge was 0.67–2.40 times higher than that of the biofilm, indicating that the S-sludge showed better ammonium oxidation activity than biofilm and the S-sludge was the main contributor for ammonium oxidation in HSBBR (Fig. 4a). For both the S-sludge and biofilm, the SNARR decreased with the increase of salinity from 1.0 to 2.0%, whereas the SNARR increased when salinity was raised to 3.0% (Fig. 4c). Except the biofilm sample at 1.0% salinity, the SNIRR declined with the salinity increase (Fig. 4d). At the salinities of 2.0 and 3.0%, the SNARR and SNIRR of both the S-sludge and biofilm were generally higher than the SAOR and SNIOR. This result implied that the denitrifier had higher tolerability to high salinity than the ammonium and nitrite oxidizers. It is worth to note that the nitrate reduction activity was fairly high at 2.0 and 3.0% salinities (Fig. 4c), although almost no nitrate was observed in the HSBBR (Fig. 3c, e). Similar result was reported by She et al. [12]. This phenomenon could be perhaps attributed to perfect expression of the denitrification genes under the favorable condition of high nitrate concentration and anaerobic environment in the batch test. Ye et al. [21] indicated that although almost no denitrification was observed in a lab-scale nitrification reactor because of the low concentration of organic matter and high concentration of dissolved oxygen, there were still quite a lot of genes related to denitrification in the reactor. In this study, the SNARR and SNIRR for the S-sludge were 4.94–11.0 and 4.13–10.01 mg Ng-1 VSS h-1, respectively, which were lower than the results in an oxic–anoxic SBR system described by She et al. [12] (8.91–32.96 mg Ng-1 VSS h-1 for SNARR, 11.57–58.51 mg Ng-1 VSS h-1 for SNIRR). This discrepancy was probably caused by the fully aerobic process in this experiment. Previous study demonstrated that the anoxic–aerobic mode had greater denitrification ability than the fully aerobic mode [5]. Wan et al. [22] reported that alternating anoxic feast/aerobic famine regime encouraged heterotrophic growth deeply inside the aggregates and hence promoted denitrification efficiency. The denitrification rates without carbon addition were much lower than those with carbon addition (Fig. 4c–f), due to the rate-limiting step of the internal carbon source consumption. Remarkably, the E-SNIRR was significantly higher than the E-SNARR for both S-sludge and biofilm (Fig. 4e, f), indicating that the activity of nitrate reductase (NAR) is more sensitive to salt stress than that of nitrite reductase (NIR) in the case of no carbon addition. The respiration rates, expressed in terms of the SOUR, showed almost constant when salinity was increased from
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1.0 to 2.0% (Fig. 4g). However, the increase of salinity to 3.0% resulted in a significant drop of SOUR. The SOUR represented respiration activity of the total microorganisms, including heterotrophs and autotrophs. The result showed that the overall activity of microorganisms was significantly inhibited at high salinity of 3.0%. As can be observed in Fig. 4h, the DHA of S-sludge and biofilm decreased with the increasing salinity. When salinity was increased to 3.0%, the DHA was suppressed at a low level (below 0.5 mg TF g-1 VSS h-1). A previous study also reported that microbial DHA was inhibited with increase of salinity [23]. The high activities of SOUR and DHA at the salinities of 1.0 and 2.0% were correlated well with high NH4?-N and TN removals observed at the salinities below 2.0% (Fig. 2). It is clear that the SOUR and DHA reflected the capability of microbial consortia to remove pollutants and can be applied to analyze the influence of salinity on the performance of the biological treatment process [23]. The EPS quantity and compositions Overall, the S-sludge contained more EPS than the biofilm, especially at higher salinities of 2.0 and 3.0% (Fig. 5). The total EPS contents of S-sludge and biofilm were 154.9–381.6 and 152.2–309.2 mg g-1 VSS, respectively. TB-EPS was the major EPS fraction, having the percentages of 45.6–72.3% in total EPS. For both the S-sludge and biofilm, S-EPS, LB-EPS and TB-EPS noticeably increased with the increase of salinity from 1.0 to 2.0%. This phenomenon could be mainly ascribed to plasmolysis and release of intracellular constituents as well as the accumulation of unmetabolized and intermediate products of incomplete degradation of organic substances and microbial produced polymers, caused by the exposure of biomass to a higher salinity [24]. With the further increase of salinity to 3.0%, the contents of S-EPS, LB-EPS and total EPS decreased obviously; however, the TB-EPS contents did not change significantly. S-EPS was an important fraction in total EPS, especially at higher salinity of 2.0 and 3.0%. The S-EPS content determined in the present study Fig. 5 Contents of S-EPS, LBEPS, TB-EPS and total EPS in the sludge and biofilm
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(7.7–85.3 mg g-1 VSS) was quite higher than that obtained by Zhao et al. (3.4–41.3 mg g-1 VSS) [4]. Such discrepancy was probably caused by the thicker biofilm (3–5 mm) developed on the combination carriers in the HSBBR in this experiment, which led to the anoxic or anaerobic metabolisms in the inner layer of the biofilm and enhanced S-EPS production. Drew set al. [25] reported that the concentration of S-EPS rose at very low DO and low nitrate concentrations. Hong et al. [26] also found that the S-EPS concentrations were higher in the anoxic phase than in the aerobic phase. Under the same saline conditions, total EPS amounts detected in this study were higher than those reported by Zhao et al. [4] (150.2–200.6 mg g-1 VSS for S-sludge, 82.2–288.6 mg g-1 VSS for biofilm). The result could be probably attributed to thicker biofilm which was developed in this study. It was reported that thick biofilm tended to produce larger EPS yields, and the EPS yields in the biofilm are proportionally related to the amount of viable biomass present [27]. The biofilm (3–5 mm) growing on the combined carries in the present study was much thicker than that (about 1 mm) growing on the moving carries used by Zhao et al. [4]. Figure 6 shows the contents of HS, PN and PS in EPS fractions. HS was identified as a dominant composition of S-EPS, TB-EPS and total EPS from the S-sludge, accounted for 77.5–90.7, 44.3–59.7 and 43.9–71.3% in S-EPS, TB-EPS and total EPS, respectively. For the biofilm, HS was also a predominant component in S-EPS, TBEPS and total EPS at 1.0 and 2.0% salinities. Such a result could be due to endogenous metabolism, release and decomposition of dead cells, PN and PS, as a response of microorganism to salt addition. The increasing salinity significantly affected HS production. As salinity increased from 1.0 to 2.0%, HS contents in S-EPS, LB-EPS and TBEPS increased by 10.4, 2.3 and 1.0 times for the S-sludge, respectively, and 5.2, 3.0 and 0.5 times for the biofilm. However, when salinity was raised to 3.0%, HS contents declined evidently. This fact was perhaps due to the significant inhibition of the microbial activity at 3.0% salinity (as shown in Fig. 4g, h), which led to the reduction in the
Bioprocess Biosyst Eng Fig. 6 Contents of HS, PN and PS in EPS from the sludge and biofilm
microbial decomposition capacity of dead cells and macromolecular organics (source of HS). PN contents increased with the increase of salinity compared to 1.0% salinity. At 3.0% salinity, PN contents in total EPS from the S-sludge and biofilm increased by 0.50 and 0.45 times, respectively. This result could be attributed to the microorganisms responding to the increasing osmotic pressure [10]. It was reported that large proportion of PN in EPS may be due to the presence of a large quantity of exoenzymes, HS in EPS could play an important role in immobilising these exoenzymes through their reversible
complexation with the enzymes [28]. PS accounted for a small proportion in total EPS (1.8–15.4% for sludge, 2.0–9.0% for biofilm) and existed mainly in TB-EPS.
Conclusions •
High removal of COD and total nitrogen via SND could be obtained in a HSBBR for the treatment of wastewater with 1.0–2.0% salinity, achieving average COD, NH4?-N and TN removal efficiencies of above 97.1,
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•
•
•
97.8 and 86.4% at the steady state. Ammonium oxidation rate at 1.0% salinity was 7.86 mg N L-1 h-1 which was almost two times of that at 2.0% salinity. Overall, high salinity of 3.0% inhibited the activity of ammonium oxidizers, nitrite oxidizers, nitrite reductase and dehydrogenase. Compared to ammonium oxidizers, nitrite oxidizers were more sensitive to salt stress. Salinity increase led to a less influence on the activity of denitrifiers than on that of ammonium oxidizers. For both the S-sludge and biofilm, the quantity of S-EPS, LB-EPS and TB-EPS noticeably increased with the increase of salinity from 1.0 to 2.0%. With the further increase of salinity to 3.0%, the contents of S-EPS and LB-EPS decreased obviously, however, the TB-EPS contents did not change significantly. Overall, the S-sludge contained more total EPS than the biofilm. Especially at 2.0% salinity, total EPS amount in the S-sludge was 381.6 mg g-1 VSS, and the value in the biofilm was 309.2 mg g-1 VSS. TB-EPS was higher than S-EPS and LB-EPS in both the S-sludge and biofilm. HS was the predominant composition in the S-sludge with the percentages of 77.5–90.7, 44.3–59.7 and 43.9–71.3% in S-EPS, TB-EPS and total EPS, respectively, followed by PN with 7.8–21.6, 33.4–46.9 and 24.9–44.1%.
Acknowledgements The work was funded by the National Natural Science Foundation of China (no. 51178437) and the Special Grand National Science & Technology Project of China for Water Pollution Control and Treatment (no. 2014ZX07203-008). Compliance with ethical standards Conflict of interest The authors declare that they have no conflict of interest.
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5. Zhang F, Li P, Chen M, Wu J, Zhu N, Wu P, Chiang P, Hu Z (2015) Effect of operational modes on nitrogen removal and nitrous oxide emission in the process of simultaneous nitrification and denitrification. Chem Eng J 280:549–557 6. Chiu Y-C, Lee L-L, Chang C-N, Chao AC (2007) Control of carbon and ammonium ratio for simultaneous nitrification and denitrification in a sequencing batch bioreactor. Int Biodeterior Biodegrad 59:1–7 7. Jin Y, Ding D, Feng C, Tong S, Suemura T, Zhang F (2012) Performance of sequencing batch biofilm reactors with different control systems in treating synthetic municipal wastewater. Biores Technol 104:12–18 8. Lo IW, Lo KV, Mavinic DS, Shiskowski D, Ramey W (2010) Contributions of biofilm and suspended sludge to nitrogen transformation and nitrous oxide emission in hybrid sequencing batch system. J Environ Sci 22:953–960 9. Wang JL, Peng YZ, Wang SY, Gao YQ (2008) Nitrogen removal by simultaneous nitrification and denitrification via nitrite in a sequence hybrid biological reactor. Chin J Chem Eng 16:778–784 10. Corsino SF, Capodici M, Morici C, Torregrossa M, Viviani G (2016) Simultaneous nitritation–denitritation for the treatment of high-strength nitrogen in hypersaline wastewater by aerobic granular sludge. Water Res 88:329–336 11. Nepa C (2002) Water and wastewater monitoring methods, 4th edn. Chinese Environmental Science Press, Beijing 12. She Z, Zhao L, Zhang X, Jin C, Guo L, Yang S, Zhao Y, Gao M (2016) Partial nitrification and denitrification in a sequencing batch reactor treating high-salinity wastewater. Chem Eng J 288:207–215 13. Panswad T, Anan C (1999) Specific oxygen ammonia and nitrate uptake rates of a biological nutrient removal process treating elevated salinity wastewater. Bioresour Technol 70:237–243 14. Klapwuk A, Drent J, Steenvoorden JHAM (1974) Modified procedure for the TTC-dehydrogenase test in activated-sludge. Water Res 8:121–125 15. Wang J, Gong B, Wang Y, Wen Y, Zhou J, He Q (2017) The potential multiple mechanisms and microbial communities in simultaneous nitrification and denitrification process treating high carbon and nitrogen concentration saline wastewater. Biores Technol 243:708–715 16. Zhao J, Huang J, Guan M, Zhao Y, Chen G, Tian X (2016) Mathematical simulating the process of aerobic granular sludge treating high carbon and nitrogen concentration wastewater. Chem Eng J 306:676–684 17. Windey K, De Bo I, Verstraete W (2005) Oxygen-limited autotrophic nitrification–denitrification (OLAND) in a rotating biological contactor treating high-salinity wastewater. Water Res 39:4512–4520 18. Ding D, Feng C, Jin Y, Hao C, Zhao Y, Suemura T (2011) Domestic sewage treatment in a sequencing batch biofilm reactor (SBBR) with an intelligent controlling system. Desalination 276:260–265 19. Holman JB, Wareham DG (2005) COD, ammonia and dissolved oxygen time profiles in the simultaneous nitrification/denitrification process. Biochem Eng J 22:125–133 20. Moussa MS, Sumanasekera DU, Ibrahim SH, Lubberding HJ, Hooijmans CM, Gijzen HJ, van Loosdrecht MC (2006) Long term effects of salt on activity, population structure and floc characteristics in enriched bacterial cultures of nitrifiers. Water Res 40:1377–1388 21. Ye L, Zhang T, Wang T, Fang Z (2012) Microbial structures, functions, and metabolic pathways in wastewater treatment bioreactors revealed using high-throughput sequencing. Environ Sci Technol 46:13244–13252 22. Wan J, Bessiere Y, Sperandio M (2009) Alternating anoxic feast/ aerobic famine condition for improving granular sludge
Bioprocess Biosyst Eng formation in sequencing batch airlift reactor at reduced aeration rate. Water Res 43:5097–5108 23. Shi K, Zhou W, Zhao H, Zhang Y (2012) Performance of halophilic marine bacteria inocula on nutrient removal from hypersaline wastewater in an intermittently aerated biological filter. Biores Technol 113:280–287 24. Jang D, Hwang Y, Shin H, Lee W (2013) Effects of salinity on the characteristics of biomass and membrane fouling in membrane bioreactors. Biores Technol 141:50–56 25. Drews A, Mante J, Iversen V, Vocks M, Lesjean B, Kraume M (2007) Impact of ambient conditions on SMP elimination and rejection in MBRs. Water Res 41:3850–3858
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RESEARCH PAPER
Salinity effect on simultaneous nitrification and denitrification, microbial characteristics in a hybrid sequencing batch biofilm reactor Zonglian She1,2,3 • Lan Wu1 • Qun Wang1 • Mengchun Gao1,2 • Chunji Jin1,2 Yangguo Zhao1,2 • Linting Zhao1 • Liang Guo1,2
•
Received: 1 June 2017 / Accepted: 21 September 2017 Springer-Verlag GmbH Germany 2017
Abstract The effect of increasing salinity on nitrogen removal via simultaneous nitrification and denitrification, microbial activities and extracellular polymeric substances (EPS) were investigated in a hybrid sequencing batch biofilm reactor filled with soft combination carriers. In the influent salinity range from 1.0 to 2.0%, average COD, NH4?-N and TN removal efficiencies were higher than 97.1, 97.8 and 86.4% at the steady state. When salinity was increased to 2.5 and 3.0%, ammonium oxidation was obviously inhibited in the reactor. For both suspended sludge (S-sludge) and biofilm, specific ammonium oxidation rate, specific nitrite oxidation rate, specific oxygen uptake rate and dehydrogenase activity reduced with the increase of salinity. The quantity of total EPS increased with the increase of salinity from 1.0 to 2.0%. Generally, humic substances were the dominant composition of EPS in both S-sludge and biofilm, with the percentages of 43.9–54.0 and 43.8–64.6% in total EPS. Keywords High salinity Simultaneous nitrification and denitrification (SND) Microbial activity Extracellular polymeric substances Humic substances
& Liang Guo [email protected] 1
Key Lab of Marine Environment and Ecology, Ocean University of China, Ministry of Education, Qingdao 266100, China
2
College of Environmental Science and Engineering, Ocean University of China, Qingdao 266100, China
3
Shandong Provincial Key Laboratory of Marine Environment and Geological Engineering, Ocean University of China, Qingdao 266100, China
Introduction Wastewaters with high salinity can be generated by chemical, pharmaceutical, petroleum, fish caning, seafood processing, meat packing, tannery, and cheese factories. The use of seawater for toilet flushing can also increase salt levels in sewage [1]. In a biological system for wastewater treatment, high salinity may decrease the cell activity, and even cause cell plasmolysis [2]. In addition, high salt concentration would affect the structure and settling characteristics of sludge flocs, production and composition of extracellular polymeric substances (EPS) of microorganism [3, 4]. Simultaneous nitrification and denitrification (SND) process has several advantages over nitrogen removal through conventional two-step nitrification and denitrification: (a) simplifying the operating procedures, (b) reducing the oxygen requirement and energy consumption [5]. In a SND system, nitrification and denitrification processes occur at the same time in a single reactor. High nitrogen removal efficiency via SND has been reported in both sequencing batch reactor (SBR) and sequencing batch biofilm reactor (SBBR) [6, 7]. Hybrid sequencing batch biofilm reactor (HSBBR) combines advantages of activated sludge, biofilm system and sequencing batch reactor. In a HSBBR, the interaction of the biofilm and S-sludge resulted in a better overall nitrogen removal performance via SND [8]. During aeration phase, nitrification could occur in S-sludge and the outer layer of biofilm where the oxygen concentration was high enough, whereas denitrification could occur in the inner layer of biofilm in which oxygen level was low. Wang et al. [9] reported that high efficiency of SND via nitrite was successfully achieved under limited dissolved oxygen (DO) concentration in a sequence hybrid biological reactor with fixed media.
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To date, a limited number of studies are available in the literature concerning the influences of salinity on the SND process. Corsino et al. [10] reported that simultaneous nitritation and denitritation was successfully sustained at the salinity up to 50 g NaCl L-1 in the treatment of fish canning wastewater by aerobic granular sludge, with a total nitrogen removal efficiency of over 90%. Chen et al. [2] found that under the salinity up to 3.0%, organic matter and nitrogen removal can be achieved simultaneously in a single-stage pressurized biofilm reactor, achieving TN removal efficiency via SND up to 98%. Although the salinity effect on the production of EPS has been studied in a moving bed SBR system based on SND process [4], the influence of high salinity on the performance of nitrogen removal via SND and the microbial characteristics of S-sludge and biofilm in HSBBR are still unclear. The main objectives of this study are (1) to investigate the effect of the increase of influent salinity on the removal of COD and nitrogen through SND in an aerobic HSBBR, (2) to analyze the variations of the specific ammonium oxidation rate (SAOR), specific nitrite oxidation rate (SNIOR), specific nitrate reduction rate (SNARR), specific nitrite reduction rate (SNIRR), specific oxygen uptake rate (SOUR) and dehydrogenase activity (DHA) for S-sludge and biofilm at different salinities, (3) to evaluate the effect of salinity on the EPS production and composition from S-sludge and biofilm.
Materials and methods The HSBBR system and operating conditions A lab-scale aerobic HSBBR (Fig. 1) which contained both S-sludge and attached biofilm was operated for 198 days in
this study. The working volume of HSBBR was 13.3 L. Four soft combination carriers (made of polyvinyl formal fiber and polypropylene hoop) were installed in the reactor, and the space between two carriers was 10 cm. Each carrier had a diameter of 14 cm and a specific surface area of 1236 m2 m-3. The photo of a combined carrier is shown in Fig. 1. The reactor was run with a cycle of 8 h. Each cycle was comprised of five phases: feeding (30 min), aeration (390 min), settling (40 min), decanting (10 min) and idle (10 min). 4.9 L of influent was fed into the reactor during the feeding phase and the same volume of effluent was discharged after the end of the settling phase, resulting in a volumetric exchanging ratio of 36.8% and a hydraulic residence time of 21.7 h. The reactor was operated at fixed temperature of 25.0 ± 1.5 C. The influent pH values were controlled at 7.50–8.02. Oxygen was supplied through two air diffuser stones in the reactor and a flow meter was used to control the airflow rate. The reactor was mixed during the feeding and aeration phases by using a submersible pump placed in the reactor to recirculate liquid. Inflow, outflow, air supply and mixing were controlled by time controllers. Before this experiment commenced, the HSBBR had been operated at low salinity (0.15–1.0%) for 440 days with glucose as carbon source in the influent and SND process had been achieved successfully. In this study, the reactor was fed with synthetic saline wastewater added with sodium acetate (NaAc) as carbon source. The wastewater was prepared with tap water and the composition was as follows (per liter): 0.35 g NaAc, 0.15 g NH4Cl, 0.04 g KH2PO4, 0.2 g NaHCO3 and seawater crystal. The components of seawater crystal solution (1.0% salinity) was as follows (mg L-1): Na? 3266; Cl- 6008; Mg2? 317; SO42- 830; K? 120; Ca2? 100; Zn2? 0.005; Mn2? 0.004; Fe2? 0.04; Co2? 1 9 10-4; Mo6? 1 9 10-3; I- 0.023; Sr? 2.5910-3; Se6? 1.17 9 10-4. Influent salinity was gradually increased from 1.0 to 1.5, 2.0, 2.5 and 3.0%. These five salinity conditions were maintained for 59, 10, 50, 29 and 50 days, respectively. Analytic methods
Fig. 1 Schematic diagram of the HSBBR system and the photo of carrier
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Water samples for the analysis of chemical oxygen demand (COD), ammonium nitrogen (NH4?-N), nitrate nitrogen (NO3--N) and nitrite nitrogen (NO2--N) were centrifuged for 10 min at 4000 rpm, then measured according to standard methods [11]. Total nitrogen (TN) was calculated by the sum of NH4?-N, NO3--N and NO2--N. Total suspended solids (TSS) and volatile suspended solids (VSS) were determined also in accordance with standard methods [11]. Temperature and DO were measured using a DO detector (Oxi 330i, WTW, Germany). pH value was determined by pH probes (PHB-4, China).
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The simultaneous nitrification and denitrification (SND) efficiency was given in Eq. (1). þ SND efficiency ¼ ½1 (NO x;e -N NOx;i -N)=ðNH4;i -N NHþ 4;e -NÞ 100%;
ð1Þ -1 where NOx,e-N (mg L ) is nitrogen in nitrite and nitrate at -1 the end of the aeration phase, NOx,i-N (mg L ) is nitrogen in nitrite and nitrate at the beginning of the aeration phase, -1 NH? 4,i-N (mg L ) is ammonium nitrogen at the beginning -1 of the aeration phase, NH? 4,e-N (mg L ) is ammonium nitrogen at the end of the aeration phase.
Activity tests Batch tests were conducted to evaluate activity of S-sludge and biofilm from the HSBBR. All samples were tested in triplicate. SAOR, SNIOR, SNARR and SNIRR were measured in conical flasks at 25 C while shaking at 150 rpm in a thermostatic water bath oscillators. Salinity in each test was identical with that of the synthetic wastewater fed into the HSBBR. For the measurement of SAOR, NH4Cl was used as nitrogen source to attain approximately 20 mg L-1 of initial nitrogen concentration in bottles. For the test of SNIOR, SNARR and SNIRR, NaNO2, NaNO3 and NaNO2 were added to flasks, respectively, to achieve 10 mg L-1 of initial nitrogen concentration. To determine SNARR and SNIRR, acetate was added as external carbon source to attain an initial COD concentration of 120 mg L-1. For the detection of endogenous SNARR (E-SNARR) and SNIRR (E-SNIRR), no carbon was added. Further details on the tests of nitrifier and denitrifier activities were described in previous study [12]. Specific oxygen uptake rate (SOUR) was determined according to Panswad and Anan [13]. S-sludge or biofilm taken from the HSBBR was mixed with fed wastewater and aerated to oxygen saturation in conical flasks. Then the flasks were sealed. The tests were performed at 25 C while stirring with a magnetic stirrer. DO in each flask was measured by a DO detector. SOUR was calculated from the slope of the DO depletion versus time divided by the VSS in the flask. Dehydrogenase activity (DHA) was tested following the modified method proposed by Klapwuket al. [14]. In this method, DHA is measured based on the reduction of triphenyl tetrazolium chloride (TTC) to triphenyl formazan (TF). The DHA was calculated according to the TF calibration curve and expressed as mg TF produced per g biomass (VSS) and per hour. EPS extraction and analysis Both S-sludge and biofilm samples were collected from the HSBBR for EPS analysis. The extracted soluble EPS (S-
EPS), loosely bound EPS (LB-EPS) and tightly bound EPS (TB-EPS) were analyzed for polysaccharides (PS), proteins (PN), and humic substances (HS). The total EPS content was taken as the sum of S-EPS, LB-EPS and TB-EPS. The extraction and measurement of EPS were performed according to the method described by Zhao et al. [4].
Results and discussion Performance of the HSBBR Figure 2 shows the overall COD, NH4?-N, and TN removal performance during the whole experiment. The excellent COD removal was attained under the five salinities where the removal rate was 95.6–99.6% at steady state (Fig. 2a). This result indicated that the heterotrophic bacteria worked effectively and the increase of salinity had no detrimental effect on the biodegradation of organics in the HSBBR. The ammonium oxidation and SND occurred successfully at salinity lower than 2.0%, with more than 89.9 and 84.4% of NH4?-N and TN removed from HSBBR, respectively (Fig. 2b, e). Similar NH4?-N removal was reported by Bassin et al. [1] whose study showed that good nitrification performance could be obtained at the salinity up to 20 g NaCl L-1. With the increase of salinity to 2.5 and 3.0%, higher effluent NH4?-N concentration was observed, average NH4?-N removal efficiencies dropped dramatically to 57.9 and 37.8%, respectively. These results indicated that autotrophic bacteria for ammonium oxidization were inhibited at higher salt concentrations. Overall, the concentration of NO3--N in the effluent was higher than that of NO2--N at 1.0 and 1.5% salinity, whereas more NO2--N in the outflow was detected when the salinity was increased to 2.0–3.0% (Fig. 2d). Nitrogen removal characteristics in the reactor To gain a better insight into the conversion of nitrogen and COD in the HSBBR, the changes of nitrogen, COD, pH and DO during aeration phase were investigated when the reactor was operated at steady state under 1.0, 2.0 and 3.0% salinities (Fig. 3). As shown in Fig. 3g, most of COD was quickly removed from bulk liquid within the first 0.5 h at 1.0% salinity and 1.5 h at 2.0% salinity, respectively; then COD concentration varied slightly during the late aeration stage. The quick decrease of COD in the early stage of aeration could be attributed to both the absorption and degradation via the S-sludge and biofilm. Some researchers [15, 16] reported that organic matter could be adsorbed rapidly by the
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Fig. 2 Influent and effluent concentration, removal efficiency of COD and nitrogen under different salinities
biofilm or aerobic granular sludge because of the existence of heterotrophic bacteria. For 3.0% salinity, fewer COD was removed in the initial stage of aeration due to the high salinity stress on heterotrophic bacteria. This study has proved that high salinity inhibited the activity of heterotrophic bacteria (As shown in Fig. 4g, h), which could reduce the absorption and utilization of COD by microorganism.
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At 1.0% salinity, ammonium was rapidly oxidized within the initial 2.0 h of the aeration phase with the decrease of NH4?-N concentration from 17.3 to 2.0 mg L-1, obtaining high ammonium oxidation rate (AOR) of 7.86 mg N L-1 h-1 in the reactor (Fig. 3a). The NH4?-N profile had a turning at point A (aerated for 2.0 h) when ammonium was almost completely transformed. TN concentration dropped from 19.0 to 6.7 mg L-1 during the
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Fig. 3 Variations of COD, nitrogen, DO and pH during aeration phase under various salinities (a ammonium turning point, b pH peak point, c pH valley point, d DO breakpoint)
aeration phase, obtaining a SND efficiency of 74.9%. This result implied that the co-existence of nitrifying bacteria and denitrifying bacteria in the hybrid biological reactor made TN removal via SND possible.
At 2.0% salinity, the AOR obtained from the reactor was 4.08 mg N L-1 h-1, which is about half of that at 1.0% salinity, and the nitrite accumulation was slower compared with that at 1.0% salinity (Fig. 3c). Under 2.0% salinity,
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there was almost no NO3--N accumulation and NO3--N level kept below 0.6 mg L-1 during the whole aerobic period, implying that nitrite oxidation was almost completely inhibited. The nitrite accumulation could be observed at 2.0% salinity during the aerobic period, achieving a maximum NO2--N concentration of 5.07 mg L-1 at 4.5 h. This was in accordance with the previous study by Corsino et al. [10] who reported that high salt concentrations lead to the accumulation of nitrite due to the inhibition of nitrite-oxidizing bacteria (NOB). Windey et al. [17] also reported that the halophilic condition in a rotating biological contactor (RBC) was favorite
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to nitritation, since NOB is more negatively affected by high salt concentrations. Furthermore, the continuous decline of TN concentration during the aeration process indicated that denitritation occurred during the whole aeration phase. Based on the much more accumulation of nitrite compared to nitrate and the complete inhibition of NOB activity at 2.0% salinity as shown in Fig. 4b, it could be concluded that short-cut simultaneous nitrification and denitrification via nitrite (i.e., simultaneous nitritation and denitritation) was the predominate process for TN removal in the HSBBR. It is also noticeable that the process of the complete nitrification and denitrification could not be
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eliminated in terms of the existence of nitrate. The presence of an anoxic microzone in the biofilm could be an explanation for the high SND efficiency (90.1%) attained at this salinity, as microzone was favorable to the development of SND [18]. As shown in Fig. 3b and d, during the first 0.5 h for 1.0% salinity and 1.0 h at 2.0% salinity (aeration from the beginning to peak point B on pH curves), pH value increased rapidly, which was perhaps due to the sufficient alkalinity in feed and/or the production of alkalinity via denitrification of nitrate and/or nitrite (remained from previous cycle or carried by sludge and biofilm). After that, pH declined quickly due to the consumption of alkalinity accompanied by the rapid oxidation of ammonium. On the whole, DO level increased continuously during the aeration phase, since the oxygen amount supplied by aeration was higher than that required for the aerobic oxidation of ammonium and COD. Holman and Wareham [19] demonstrated that the depletion of COD and NH4?-N caused a sudden DO increase, which was also observed in the present study at 1.0 and 2.0% salinities. When aeration has been lasted for nearly 2.0 and 4.0 h, the DO profiles had breakpoint at point D (Fig. 3b, d). After this point DO appeared a sudden increase, which could be attributed to the decrease of oxygen demand after ammonium had been almost completely removed. It is notable that when DO reached the breakpoint D, pH got to the valley point C almost at the same time. Therefore, both the pH valley point and DO break point could be used to indicate the end of rapid ammonium oxidation. For 3.0% salinity, the serious inhibition of high salt stress on ammonium oxidation led to high ammonium residue (21.1 mg L-1), and almost no nitrite and nitrate accumulation at the end of aeration (Fig. 3e). Microbial activity at different salinities The effect of salinity on the activities of ammonium oxidizers from the S-sludge and biofilm was similar (Fig. 4a). The increase of salinity led to lower SAOR. At 3.0% salinity, the SAOR for the S-sludge and biofilm declined by 69.3 and 82.2%, respectively, compared with 1.0% salinity. This result explained why NH4?-N accumulation occurred in the HSBBR at 3.0% salinity, as shown in Fig. 2b and Fig. 3e. At 2.0 and 3.0% salinities, the SNIOR of both S-sludge and biofilm was not detected, indicating that the activity of nitrite oxidizers was completely inhibited. No accumulation of nitrate in the reactor at 2.0 and 3.0% salinity (Fig. 3c, e) also demonstrated that nitrite oxidation was completely restrained in the HSBBR. The decrease in the activity of ammonium and nitrite oxidizers found in this study agrees with previous studies [1, 20]. It is worth noting that at 1.0% salinity, the SAOR for both S-sludge
and biofilm were obviously higher than the SNIOR (Fig. 4a, b), implying that nitrite oxidizers were more sensitive to salt stress than ammonium oxidizers. The SAOR of the S-sludge was 0.67–2.40 times higher than that of the biofilm, indicating that the S-sludge showed better ammonium oxidation activity than biofilm and the S-sludge was the main contributor for ammonium oxidation in HSBBR (Fig. 4a). For both the S-sludge and biofilm, the SNARR decreased with the increase of salinity from 1.0 to 2.0%, whereas the SNARR increased when salinity was raised to 3.0% (Fig. 4c). Except the biofilm sample at 1.0% salinity, the SNIRR declined with the salinity increase (Fig. 4d). At the salinities of 2.0 and 3.0%, the SNARR and SNIRR of both the S-sludge and biofilm were generally higher than the SAOR and SNIOR. This result implied that the denitrifier had higher tolerability to high salinity than the ammonium and nitrite oxidizers. It is worth to note that the nitrate reduction activity was fairly high at 2.0 and 3.0% salinities (Fig. 4c), although almost no nitrate was observed in the HSBBR (Fig. 3c, e). Similar result was reported by She et al. [12]. This phenomenon could be perhaps attributed to perfect expression of the denitrification genes under the favorable condition of high nitrate concentration and anaerobic environment in the batch test. Ye et al. [21] indicated that although almost no denitrification was observed in a lab-scale nitrification reactor because of the low concentration of organic matter and high concentration of dissolved oxygen, there were still quite a lot of genes related to denitrification in the reactor. In this study, the SNARR and SNIRR for the S-sludge were 4.94–11.0 and 4.13–10.01 mg Ng-1 VSS h-1, respectively, which were lower than the results in an oxic–anoxic SBR system described by She et al. [12] (8.91–32.96 mg Ng-1 VSS h-1 for SNARR, 11.57–58.51 mg Ng-1 VSS h-1 for SNIRR). This discrepancy was probably caused by the fully aerobic process in this experiment. Previous study demonstrated that the anoxic–aerobic mode had greater denitrification ability than the fully aerobic mode [5]. Wan et al. [22] reported that alternating anoxic feast/aerobic famine regime encouraged heterotrophic growth deeply inside the aggregates and hence promoted denitrification efficiency. The denitrification rates without carbon addition were much lower than those with carbon addition (Fig. 4c–f), due to the rate-limiting step of the internal carbon source consumption. Remarkably, the E-SNIRR was significantly higher than the E-SNARR for both S-sludge and biofilm (Fig. 4e, f), indicating that the activity of nitrate reductase (NAR) is more sensitive to salt stress than that of nitrite reductase (NIR) in the case of no carbon addition. The respiration rates, expressed in terms of the SOUR, showed almost constant when salinity was increased from
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1.0 to 2.0% (Fig. 4g). However, the increase of salinity to 3.0% resulted in a significant drop of SOUR. The SOUR represented respiration activity of the total microorganisms, including heterotrophs and autotrophs. The result showed that the overall activity of microorganisms was significantly inhibited at high salinity of 3.0%. As can be observed in Fig. 4h, the DHA of S-sludge and biofilm decreased with the increasing salinity. When salinity was increased to 3.0%, the DHA was suppressed at a low level (below 0.5 mg TF g-1 VSS h-1). A previous study also reported that microbial DHA was inhibited with increase of salinity [23]. The high activities of SOUR and DHA at the salinities of 1.0 and 2.0% were correlated well with high NH4?-N and TN removals observed at the salinities below 2.0% (Fig. 2). It is clear that the SOUR and DHA reflected the capability of microbial consortia to remove pollutants and can be applied to analyze the influence of salinity on the performance of the biological treatment process [23]. The EPS quantity and compositions Overall, the S-sludge contained more EPS than the biofilm, especially at higher salinities of 2.0 and 3.0% (Fig. 5). The total EPS contents of S-sludge and biofilm were 154.9–381.6 and 152.2–309.2 mg g-1 VSS, respectively. TB-EPS was the major EPS fraction, having the percentages of 45.6–72.3% in total EPS. For both the S-sludge and biofilm, S-EPS, LB-EPS and TB-EPS noticeably increased with the increase of salinity from 1.0 to 2.0%. This phenomenon could be mainly ascribed to plasmolysis and release of intracellular constituents as well as the accumulation of unmetabolized and intermediate products of incomplete degradation of organic substances and microbial produced polymers, caused by the exposure of biomass to a higher salinity [24]. With the further increase of salinity to 3.0%, the contents of S-EPS, LB-EPS and total EPS decreased obviously; however, the TB-EPS contents did not change significantly. S-EPS was an important fraction in total EPS, especially at higher salinity of 2.0 and 3.0%. The S-EPS content determined in the present study Fig. 5 Contents of S-EPS, LBEPS, TB-EPS and total EPS in the sludge and biofilm
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(7.7–85.3 mg g-1 VSS) was quite higher than that obtained by Zhao et al. (3.4–41.3 mg g-1 VSS) [4]. Such discrepancy was probably caused by the thicker biofilm (3–5 mm) developed on the combination carriers in the HSBBR in this experiment, which led to the anoxic or anaerobic metabolisms in the inner layer of the biofilm and enhanced S-EPS production. Drew set al. [25] reported that the concentration of S-EPS rose at very low DO and low nitrate concentrations. Hong et al. [26] also found that the S-EPS concentrations were higher in the anoxic phase than in the aerobic phase. Under the same saline conditions, total EPS amounts detected in this study were higher than those reported by Zhao et al. [4] (150.2–200.6 mg g-1 VSS for S-sludge, 82.2–288.6 mg g-1 VSS for biofilm). The result could be probably attributed to thicker biofilm which was developed in this study. It was reported that thick biofilm tended to produce larger EPS yields, and the EPS yields in the biofilm are proportionally related to the amount of viable biomass present [27]. The biofilm (3–5 mm) growing on the combined carries in the present study was much thicker than that (about 1 mm) growing on the moving carries used by Zhao et al. [4]. Figure 6 shows the contents of HS, PN and PS in EPS fractions. HS was identified as a dominant composition of S-EPS, TB-EPS and total EPS from the S-sludge, accounted for 77.5–90.7, 44.3–59.7 and 43.9–71.3% in S-EPS, TB-EPS and total EPS, respectively. For the biofilm, HS was also a predominant component in S-EPS, TBEPS and total EPS at 1.0 and 2.0% salinities. Such a result could be due to endogenous metabolism, release and decomposition of dead cells, PN and PS, as a response of microorganism to salt addition. The increasing salinity significantly affected HS production. As salinity increased from 1.0 to 2.0%, HS contents in S-EPS, LB-EPS and TBEPS increased by 10.4, 2.3 and 1.0 times for the S-sludge, respectively, and 5.2, 3.0 and 0.5 times for the biofilm. However, when salinity was raised to 3.0%, HS contents declined evidently. This fact was perhaps due to the significant inhibition of the microbial activity at 3.0% salinity (as shown in Fig. 4g, h), which led to the reduction in the
Bioprocess Biosyst Eng Fig. 6 Contents of HS, PN and PS in EPS from the sludge and biofilm
microbial decomposition capacity of dead cells and macromolecular organics (source of HS). PN contents increased with the increase of salinity compared to 1.0% salinity. At 3.0% salinity, PN contents in total EPS from the S-sludge and biofilm increased by 0.50 and 0.45 times, respectively. This result could be attributed to the microorganisms responding to the increasing osmotic pressure [10]. It was reported that large proportion of PN in EPS may be due to the presence of a large quantity of exoenzymes, HS in EPS could play an important role in immobilising these exoenzymes through their reversible
complexation with the enzymes [28]. PS accounted for a small proportion in total EPS (1.8–15.4% for sludge, 2.0–9.0% for biofilm) and existed mainly in TB-EPS.
Conclusions •
High removal of COD and total nitrogen via SND could be obtained in a HSBBR for the treatment of wastewater with 1.0–2.0% salinity, achieving average COD, NH4?-N and TN removal efficiencies of above 97.1,
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97.8 and 86.4% at the steady state. Ammonium oxidation rate at 1.0% salinity was 7.86 mg N L-1 h-1 which was almost two times of that at 2.0% salinity. Overall, high salinity of 3.0% inhibited the activity of ammonium oxidizers, nitrite oxidizers, nitrite reductase and dehydrogenase. Compared to ammonium oxidizers, nitrite oxidizers were more sensitive to salt stress. Salinity increase led to a less influence on the activity of denitrifiers than on that of ammonium oxidizers. For both the S-sludge and biofilm, the quantity of S-EPS, LB-EPS and TB-EPS noticeably increased with the increase of salinity from 1.0 to 2.0%. With the further increase of salinity to 3.0%, the contents of S-EPS and LB-EPS decreased obviously, however, the TB-EPS contents did not change significantly. Overall, the S-sludge contained more total EPS than the biofilm. Especially at 2.0% salinity, total EPS amount in the S-sludge was 381.6 mg g-1 VSS, and the value in the biofilm was 309.2 mg g-1 VSS. TB-EPS was higher than S-EPS and LB-EPS in both the S-sludge and biofilm. HS was the predominant composition in the S-sludge with the percentages of 77.5–90.7, 44.3–59.7 and 43.9–71.3% in S-EPS, TB-EPS and total EPS, respectively, followed by PN with 7.8–21.6, 33.4–46.9 and 24.9–44.1%.
Acknowledgements The work was funded by the National Natural Science Foundation of China (no. 51178437) and the Special Grand National Science & Technology Project of China for Water Pollution Control and Treatment (no. 2014ZX07203-008). Compliance with ethical standards Conflict of interest The authors declare that they have no conflict of interest.
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