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Oxygen-limited autotrophic nitrification–denitrification (OLAND) in a rotating biological contactor treating high-salinity wastewater
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Water Research 39 (2005) 4512–4520 www.elsevier.com/locate/watres
Oxygen-limited autotrophic nitrification–denitrification (OLAND) in a rotating biological contactor treating high-salinity wastewater Kim Windey, Inge De Bo, Willy Verstraete Laboratory of Microbial Ecology and Technology (LabMET), Ghent University, Coupure Links 653, B-9000 Ghent, Belgium Received 9 January 2004; received in revised form 2 June 2005; accepted 8 September 2005
Abstract A lab-scale rotating biological contactor (RBC) reactor operated under OLAND conditions was slowly adapted during 178 days to increasing salt concentrations going up to 30 g NaCl L1. The reactor performed well during this experimental period. However, the removal capacity of the reactor was lower under high-salinity conditions. A removal efficiency of 84% was achieved at a N loading rate of 725 mg N L1 d1 and a salt concentration of 30 g L1. The effect of salt shock loading and adaptation to 30 g NaCl L1 on the specific nitritation and anammox activity of the biomass was investigated in short-term batch experiments. A salt shock loading of 30 g L1 caused a 43% decrease in specific nitritation activity and 96% loss of specific anammox activity compared to reference biomass (not exposed to salt). The salt-adapted biomass (3–4 weeks) showed a specific nitritation activity that was 23% lower, and a specific anammox activity that was 58% lower, compared to the reference biomass. Overall, these results demonstrate that the OLAND process can have the potential to treat ammonium-rich brines after adaptation to high salinity. r 2005 Elsevier Ltd. All rights reserved. Keywords: OLAND biofilm reactor; Nitritation; Anammox; Nitrogenous wastewater of high salinity; Salt shock loading
1. Introduction The removal of nitrogen from wastewater has become an important part of the overall treatment process due to the significant impact of nitrogen compounds on the environment and the more stringent legislation on wastewater discharges. Conventionally, nitrogen removal is performed by means of a two-step process. First, ammonium (NH+ 4 ) is oxidized to nitrite (NO2 ) or nitrate (NO3 ) by the aerobic (autotrophic) nitrification process and subsequently reduced to dinitrogen gas (N2) Corresponding author. Tel.: +32 9 264 59 76; fax: +32 9 264 62 48. E-mail address: [email protected] (W. Verstraete).
by the anaerobic (heterotrophic) denitrification process (Focht and Verstraete, 1977). Nowadays, anaerobic digestion plants are being developed that minimize energy consumption, CO2 emission and sludge production. However, these systems typically yield effluents rich in NH+ 4 and poor in biodegradable organic carbon, thereby making them less suitable for biological N removal via the conventional nitrification–denitrification process (Cecchi and Battistoni, 2003). Recently, several new N removal processes have been developed to treat these highly loaded wastewaters. The process in which NH+ is autotrophically oxidized 4 to N2 with NO as the electron acceptor under 2 oxygen-limited conditions is further referred to as oxygen-limited autotrophic nitrification denitrification
0043-1354/$ - see front matter r 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2005.09.002
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(OLAND) (Kuai and Verstraete, 1998). This autotrophic process consumes 63% less oxygen and 100% less biodegradable organic carbon compared to the conventional nitrification–denitrification process and has, therefore, a lower operating cost (Verstraete and Philips, 1998). The OLAND process was first described for a mixed culture of nitrifying bacteria (Kuai and Verstraete, 1998), but was afterwards examined in more detail in a mixed community biofilm of a lab-scale rotating biological contactor (RBC) (Pynaert et al., 2002a, b, 2003, 2004). A mature OLAND biofilm under high NH+ 4 loading rate consists primarily of two major groups of bacteria responsible for autotrophic N removal. The aerobic ammonium oxidizing bacteria (AerAOB, Nitrosomonas sp.) convert NH+ 4 to NO2 with oxygen as the electron acceptor (nitritation) and the anaerobic ammonium oxidizing bacteria (AnAOB, a close relative of Kuenenia stuttgartiensis) subsequently oxidize NH+ with NO 4 2 as the electron acceptor (anammox) (Strous et al., 1998; Pynaert et al., 2003; Wyffels et al., 2003). The overall N removal stoichiometry can be summarized by the following equations (Pynaert et al., 2004): Nitritation :
Anammox :
1:32 NH3 þ 1:98 O2 ! 1:32 NO2 þ 1:32 Hþ þ 1:32 H2 O;
NH3 þ 1:32 NO2 þ Hþ ! 0:26 NO3 þ 1:02 N2 þ 2 H2 O;
N removal :
(1)
(2)
NH3 þ 0:85 O2 ! 0:11 NO3 þ 0:44 N2 þ 0:14 Hþ þ 1:43 H2 O: (3)
High nitrogen-concentrated streams can also contain large amounts of other ions like chloride (fish canning industry, wet lime–gypsum desulphurization process, regeneration liquid from ion exchange columns), sulphate (tannery wastes), etc. High saline concentrations in wastewater are known to have negative effects on nitrogen removal (Ludzack and Noran, 1965; Chen et al., 1971; Panswad and Anan, 1999b). However, Dahl et al. (1997) showed that with an adequate adaptation and slowly increased salt concentration, municipal activated sludge could be adapted to high-salinity wastewater containing up to 33 g NaCl L1. Panswad and Anan (1999a) showed an improvement in the N removal efficiency of a bioreactor treating very highsalinity wastewater (30 g NaCl L1) when seeded with NaCl acclimated cultures. Campos et al. (2002) found that adapted biomass is less sensitive to high saline concentrations. On the contrary, Catalan et al. (1997) observed that when seawater was nitrified, NO 2 accumulated even when marine genera adapted to high saline concentrations were used. Schenk and Hegemann (1995) and Dinc- er and Kargi (1999) also found that nitrite oxidizing bacteria (NOB) were more sensitive to
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salt concentration than AerAOB. Only Hunik et al. (1992, 1993) concluded the opposite. Based on the available literature, it can be concluded that the effects of NaCl on the AerAOB are not well understood and contradictory findings are reported. The effect of salt on the AnAOB has not been studied yet, although one marine AnAOB named Candidatus Scalindua sorokinii was recently discovered in the Black Sea (Kuypers et al., 2003). However, N removal under highly saline conditions was not studied (Kuypers et al., 2003). The objective of this study was to evaluate the feasibility of the OLAND process, operated in a RBC, to treat NH+ 4 -rich wastewater of high salinity. The specific nitritation and anammox activities of the AerAOB and AnAOB, respectively, were investigated before exposure to salt, at a shock loading of 30 g NaCl L1 and after an adaptation period of 3–4 weeks.
2. Materials and methods 2.1. Lab-scale RBC reactor The reactor used in this study is shown in Fig. 1. Its characteristics were as follows: basin length L ¼ 88 cm, width B ¼ 34 cm, height H ¼ 30 cm, water level at 19 cm and total water volume 50 L. Two series of 20 polyvinylchloride discs were mounted (diameter of one disc ¼ 30 cm, thickness ¼ 0.5 cm, 1 cm interspace) on a horizontal shaft, which operated at 2.5 rpm (a low value was chosen so that the O2 concentration is always less than 1 mg L1) with 58% submergence of the disc surface area. The total surface area of the wetted discs and reactor walls was 6.3 m2. In order to increase the specific surface area for biofilm growth, a 5 mm layer of reticulated polyurethane (specific surface area
Fig. 1. Schematic representation of the 50 L lab-scale RBC reactor. Discs are partially (58%) submerged in the wastewater, and rotate round a horizontal axis at 2.5 rpm. Influent is dosed at two points in the reactor and is discharged on the other side.
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71000 m2 m3, Type Filtren TM20, Recticel, Belgium) was attached to one side of every disc. This increased the initial available surface area to about 20 m2. The influent was dosed at two points in the reactor, one at the opposite end of the outlet and one halfway between the two groups of discs, to maximize biomass growth on both groups of discs. The pH in the reactor was measured regularly. 2.2. Synthetic wastewater Synthetic wastewater was fed to the RBC, prepared with tap water and containing NH+ 4 as (NH4)2SO4 and/ or urea (N source), NaHCO3 (C source and buffer) and NaCl (salt). The amounts of NH+ 4 and NaCl varied depending on the applied load. Per gram of NH+ 4 –N in the feed, 7 g of NaHCO3 was supplied. In addition, the wastewater contained 0.07 g L1 of phosphorous as KH2PO4. 2.3. Aerobic batch experiments Aerobic batch experiments were performed in 250 mL Erlenmeyer flasks with 100 mL working volume. From previous batch experiments it was observed that the standard deviation was always less than 8% (Pynaert et al., 2003). Therefore, the indicative batch experiments with reference biomass and shock-loaded biomass were not performed in triplicate. The batch experiment with adapted biomass was performed in triplicate. At the 1 beginning of the test, 0.1 g NH+ and a phosphate 4 –N L buffer with final concentrations of 3.4 g L1 KH2PO4, 4.4 g L1 K2HPO4 and 1.0 g L1 NaHCO3 were supplied to each Erlenmeyer flask. A representative sludge sample was collected from the reactor by scraping off parts of the biofilm (in the whole depth) of several discs. This sludge sample was divided in equal amounts and added to the flasks (2.570.2 g volatile suspended solids (VSS) L1). The flasks were incubated for 6 h at 2872 1C on a shaker (140 rpm) to provide oxygen. The pH was initially adjusted to 7.5770.02 with NaOH (0.1 N). During the experiment, the pH and dissolved oxygen (DO) level were monitored every 1.5 h. Samples of the liquor phase were taken with a sterile syringe and needle, filtered over a 0.45 mm filter (Millipore) and stored at 4 1C for 1 day.
experiments. After aliquots of biofilm sludge were added (2.970.1 g VSS L1), the flasks were closed with butyl rubber stoppers and aluminium caps. The initial pH of all the flasks was 8.0. To obtain anaerobic conditions, the flasks were flushed with N2 gas for 30 min in cycles of 2 min overpressure (70 kPa) and underpressure (90 kPa). Substrate solutions (NH4Cl and NaNO2) were also prepared in serum flasks, closed with butyl rubber stoppers and aluminium caps and flushed with N2 gas. An appropriate volume of each substrate solution was added to the different flasks with a sterile syringe until the final concentration of 0.1 g 1 1 NH+ and 0.1 g NO in each flask was 4 –N L 2 –N L obtained. The flasks were incubated for 6 h at 3072 1C. Samples of the liquor phase were obtained in a similar way as described for the aerobic batch tests. For the same reason as stated for the aerobic experiments, the experiments with the reference and shock-loaded biomass were not performed in triplicate. The batch experiment with the adapted biomass was performed in triplicate. 2.5. Chemical analyses NH+ 4 –N in the effluent of the RBC reactor was determined by the Kjeldahl distillation method as described by Bremner and Keeney (1965). NH+ 4 –N in the batch experiments (aerobic and anaerobic) was determined colorimetrically with Nessler reagent according to standard methods (Greenberg et al., 1992). Both NO 2 –N and NO3 –N were determined using a Methrom 761 Compact Ion Chromatograph (Methrom, Herisau, Switserland) equipped with a conductivity detector. The operational parameters were as follows: column metrosep A supp 5, eluents 1.06 g L1 Na2CO3, flow 0.7 mL min1, sample loop 20 mL. VSS and total suspended solids were determined by drying and weighing according to standard methods (Greenberg et al., 1992). The pH was determined potentiometrically with a digital, portable pH meter (Knick, portamess 751). The DO level was measured with a digital, portable DO meter (Endress-Hauser, COM 381).
3. Results 3.1. Performance of the RBC reactor
2.4. Anaerobic batch experiments Anaerobic batch experiments were performed in 120 mL serum flasks, containing 80 mL of mixed liquor (40 mL headspace). The mixed liquor contained per litre 1 g NaHCO3, 0.05 g KH2PO4, 2 mL trace element solution (Kuai and Verstraete, 1998) with or without salt (NaCl). A representative sludge sample was collected in a similar way as described for the aerobic batch
The RBC reactor performance is depicted in Fig. 2 (Panels I and II). Five periods with different N loading rates can be distinguished (R, A, B, C and D). Parameter settings of the different periods are summarized in Table 1. During the reference period (period R), the average N removal capacity was 885731 mg N L1 d1 (n ¼ 5) at a constant loading rate of 1000 mg N L1 d1. From day 12, the RBC reactor
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the RBC feed was increased from zero to 5 g NaCl L1 (Fig. 2, Panel I). A stepwise increase of 1 g NaCl L1 was applied. This salt addition had no effect on the removal capacity of the reactor. The average N removal capacity
performance was operated at a similar N loading rate and under increasing salt concentrations. From day 12 to day 25 (period A), the N loading rate was 1000 mg N L1 d1, while the salt concentration of
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Fig. 2. Performance of the RBC reactor under changing salt concentration addition. Panel I: loading rate (’) and removal capacity (W) of the reactor (mg N L1 d1) in relation to the salt concentration (g NaCl L1) (}). Panel II: influent NH+ 4 (~) concentration 1 (mg N L1), effluent NH+ 4 (’), NO2 (J) and NO3 (}) concentrations (mg N L ), and N removal percentage (%) (W). Panel III: pH (~) and DO concentrations (}) (mg O2 L1) as a function of time. Vertical arrows indicate process disturbances; on three occasions a lower loading rate was applied due to pump failures.
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Fig. 2. (Continued) Table 1 Characteristics of the RBC reactor operated during periods of different N loading rates and salt concentration range Parameters
Period R
Period A
Period B
Period C
Period D
Time range (d) Feed flow rate (L d1) Hydraulic retention time (d) 1 Feed concentration (mg NH+ 4 –N L ) 1 1 Loading rate (mg N L d ) Salt level (g NaCl L1)
0–11 65 0.77 770 1000 0
12–25 65 0.77 770 1000 0-5
26–81 65 0.77 846 1100 5-10
82–135 55 0.91 769 849 10-30
136–178 55 0.91 476– 657 525–725 30
was 927710 mg N L1 d1 (n ¼ 8), representing a removal percentage of 9371% (n ¼ 8). The NH+ 4 concentration in the effluent was always below 20 mg N L1. The NO 3 concentration in the effluent had an average value of 3775 mg N L1 (n ¼ 8) (Fig. 2, Panel II). Because of this high removal percentage, the N loading rate was further increased to 1100 mg N L1 d1 from day 26 on (period B). On day 33, the salt concentration was increased to 6 g NaCl L1 (Fig. 2, Panel I) and no remarkable loss of activity was measured. The salt concentration was elevated from 6 to 10 g NaCl L1 on day 50. From day 60 on, the reactor performance decreased to a removal capacity of only 641 mg N L1 d1 on day 74 (removal efficiency ¼ 58%). During the whole of period B, no
NO 2 peaks were observed. On day 74, an elevated 1 effluent concentration of 341 mg NH+ and a pH 4 –N L of 8.5 were measured (Fig. 2, Panel III). The low removal percentage can be explained by a free ammonia inhibition. Following this, the N loading rate was temporarily lowered to 849 mg N L1 d1 on day 82 (period C) by means of a lower influent NH+ concentration and 4 by an increase of the hydraulic retention time to 0.9 day (Table 1). During the following days, the removal percentage increased from 57% to 73%. To regain the high removal capacity of the reactor, no extra salt was added to the influent from day 85 to day 91. The N removal percentage of the reactor further increased to 83%. On day 92, the RBC feed was again supplied with 10 g NaCl L1. The removal percentage of the reactor
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first decreased but returned to 86% after 1 week. From day 114 to day 125, the salt concentration was further increased up to 30 g NaCl L1 in steps of 5 g NaCl L1. When a salt concentration of 20 g NaCl L1 was achieved, the N removal capacity started decreasing until the end of period C. The effluent NH+ 4 concentration increased to 201 mg N L1 on day 135, resulting in a removal percentage of 70%. However, no NO 2 peaks were observed in the effluent. On days 126, 127 and 132, indicated by vertical arrows on Fig. 2 (Panel I and II), a lower N loading rate (424 mg N L1 d1) was supplied due to accidental interruption of the feeding. On day 136, it was decided to lower the N loading rate and to subject the biomass to a salt concentration of 30 g NaCl L1. During a time span of 43 days, the performance at increasing loading rates between 525 and 725 mg N L1 d1 was investigated (Fig. 2, Panels I and II). At a N loading rate of 525 mg N L1 d1, a removal percentage of 8170.5% (n ¼ 5) was obtained. The removal capacity of the reactor was however 52% lower compared to the reference period before salt addition to the reactor. The increase of the influent NH+ 4 –N concentration, but temporary decrease of the flow rate, resulted in a N loading rate of 571 mg N L1 d1 and a removal percentage of 8376% (n ¼ 2). An increase of the N loading rate to 625 mg N L1 d1 resulted in a decrease of the N removal percentage to 7479% (n ¼ 6) and hence a N removal capacity of 463742 mg N L1 d1 (n ¼ 6). At a N loading rate of 725 mg N L1 d1 one can distinguish two periods. From day 144 to day 171, a N removal capacity of 485735 mg N L1 d1 (n ¼ 6) or a removal percentage of 6775% (n ¼ 6) was measured. The highest N
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removal capacity, 609715 mg N L1 d1 (n ¼ 5), was achieved from day 172 to day 178, which is 31% lower compared to the reference period before salt addition to the reactor. 3.2. Specific nitritation activity of OLAND biomass Short-term aerobic batch experiments were performed with an OLAND biomass of periods R and D. The biomass of period R was used to measure the reference specific nitritation activity of the AerAOB (no salt) and the specific activity when these AerAOB are exposed to a salt shock loading of 30 g NaCl L1 (Fig. 3: reference biomass R; shock-loaded biomass S). After an exposure period of 23 days to 30 g NaCl L1 (period D), the specific activity of the AerAOB was also measured (Fig. 3: adapted biomass A). During the experiment, the pH decreased with an average value of 0.1370.06. The DO level was always above 3.3 mg O2 L1. The specific nitritation activity of the reference biomass was 60 mg 1 NH+ VSS d1, whereas a decrease of 43% was 4 –N g observed for the shock-loaded biomass, resulting in a 1 specific nitritation activity of 34 mg NH+ 4 –N g 1 VSS d . The adapted biomass showed a specific 1 nitritation activity of 4674 mg NH+ VSS d1. 4 –N g Hence, this was only 24% lower than the specific activity of the reference biomass. In the aerobic batch experiments with reference 1 biomass, a maximum of 7 mg NO was detected 3 –N L (data not shown). When salt was added to shockloaded respectively to adapted biomass, no NO 3 was detected in the liquor phase (detection limit of 0.01 mg 1 NO 3 –N L ).
Specific activity (mg NH4+-N g-1 VSS d-1)
80 70 60 50 40 30 20 10 0
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Fig. 3. Overview of the specific nitritation activity of the AerAOB (light grey bars) and the specific anammox activity of the AnAOB (dark grey bars) determined in batch experiments. R ¼ reference biomass (period R), S ¼ biomass exposed to a salt shock loading of 30 g NaCl L1 (period R), A ¼ biomass adapted to salt (period D) (n ¼ 3).
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3.3. Specific anammox activity of OLAND biomass Anaerobic batch experiments were performed in order to evaluate the specific anammox activity of the AnAOB in the reference, salt shock-loaded and salt-adapted biomass as defined above. The reference specific 1 anammox activity was 77 mg NH+ VSS d1 4 –N g 1 (Fig. 3, R). A shock loading of 30 g NaCl L caused a decrease in specific activity of 95% (Fig. 3, S), resulting 1 in a specific anammox activity of 3 mg NH+ 4 –N g 1 VSS d . The salt-adapted AnAOB (Fig. 3, A) had a 1 specific activity of 3272 mg NH+ VSS d1, 4 –N g which means that adaptation of the AnAOB to salt had occurred. However, the specific activity was still 59% lower than the reference.
4. Discussion The reactor performance was not negatively affected by exposing the OLAND biofilm to salt concentrations up to 6 g NaCl L1. A stimulation of the microbial activity of various other bacterial species at low salt concentrations has been reported in the literature (Chen et al., 2003; Hamoda and Al–Attar, 1995; Ingram, 1940). Chen et al. (2003) reported that a salt concentration below 4.12 g NaCl L1 was in favour of the nitrification rate, but exceeding this level resulted in a decrease. Hamoda and Al-Attar (1995) also stated a stimulation of the activity of microorganisms developed in a freshwater environment below this level. Early studies with Bacillus cereus indicated that endogenous respiration rate of the culture increased with NaCl concentrations up to 10 g L1, but decreased above this level (Ingram, 1940). An explanation for the fluctuations during the period 60–113 days could be that at a salt concentration of 10 g NaCl L1, the AerAOB became initially inhibited. However, after an adaptation period with the higher salt concentration, saline-resistant AerAOB may develop. The imbalance between the amount of freshwater and saline-resistant AerAOB can also be an explanation for the decreased reactor performance at salt levels above 20 g NaCl L1 in period C. Vredenbregt et al. (1997) considered 16.5 g NaCl L1 as a critical level for a freshwater nitrifying culture. Chen et al. (2003) reported a shift in the dominant species of AerAOB from non-saline-resistant species, such as Nitrosomonas europaea lineage and Nitrosomonas eutropha, to saline-resistant species such as Nitrosococcus mobilis lineage when the salt concentration was increased from 16.5 to 30 g NaCl L1. Rapid shifts in salt concentration were reported to have more adverse effects than gradual shifts (Burnett, 1974; Oren et al., 1992; Chen et al., 2003). However, Kincannon and Gaudy (1968) found that a decrease in salt concentration caused more severe negative effects on microorganisms
than an increase in salt concentration. They stated that rapid changes in salt concentration caused immediate release of cellular constituents resulting in an increase in soluble organic carbon. The OLAND process operated at 30 g NaCl L1 can achieve similar N removal percentages as when operated without salt by decreasing the N loading rate with 28% to a value of 725 mg N L1 d1. Further work on the microbial ecology would be necessary to verify whether the OLAND microbial community has undergone a physiological adaptation and/or there has been a substantial shift in microbial species. The halophilic conditions in the RBC reactor were an advantage for the OLAND process, since NOB are more negatively affected by high salt concentrations (Schenk and Hegemann, 1995; Dinc- er and Kargi, 1999). The NOB are unwanted in an OLAND reactor because they oxidize NO 2 to NO3 when sufficient oxygen is present and therefore result in an increase of N components in the effluent. In the aerobic batch experiments with OLAND biomass, no NO 3 was detected when salt was 1 supplied, whereas a maximum of 7 mg NO was 3 –N L detected without salt addition. Hence, the few NOB that were present were further repressed when the reactor feed was supplied with NaCl. During the whole period of salt addition to the reactor, no NO peaks were observed. Hence, 2 the anaerobic ammonium oxidation activity of the OLAND biofilm was not affected by the increase of the salt level. The highest value measured from period R 1 to period D was 16 mg NO on day 18 (period 2 –N L A). This was probably due to the shortage in electron donor, because the NH+ 4 concentration in the effluent 1 was only 3 mg NH+ 4 –N L . Consequently, the ratelimiting step in the OLAND process at elevated salt levels seems to be nitritation because all NO 2 produced was immediately used to further oxidize NH+ 4 anaerobically to N2. However, from the batch experiments it can be concluded that the AnAOB are more influenced by high salt concentrations than the AerAOB, when exposed to a salt shock loading, as well as after an adaptation period of approximately 1 month. It seems that the AnAOB, when encapsulated in a thick biofilm, were better able to withstand high salt levels. It therefore appears that, overall, both groups of bacteria can perform under saline conditions, but further work is necessary to elucidate the physiological basis of their respective salt tolerance mechanisms. Overall, these data indicate that the OLAND process, with careful monitoring of the DO concentration, pH, NH+ 4 –N and NO2 –N concentration within a few months, can become adapted to handle NH+ 4 -rich saline solutions, as often produced by fish canning industries (Mosquera-Corral et al., 2000), the wet lime–gypsum desulphurization process (Vredenbregt et al., 1997) and
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regeneration of ion exchange columns (Semmens and Porter, 1979).
5. Conclusions 1. An OLAND biofilm reactor was exposed to salinity levels up to 30 g NaCl L1 during a time span of 178 days. 2. Autotrophic nitrogen removal of high-salinity wastewater with high nitrogen concentrations can be effectively achieved at loading rates up to 725 mg N L1 d1, reaching N removal efficiencies of 84%. The N removal capacity was 31% lower at a salt level of 30 g NaCl L1 compared to the reference period without salt addition. 3. Short-term batch experiments (both aerobic and anaerobic) were performed to investigate the salt shock loading and adaptation to salt levels of 30 g NaCl L1. A salt shock loading of 30 g L1 caused a 43% decrease in specific nitritation activity and 96% loss of specific anammox activity compared to reference biomass (not exposed to salt). The saltexposed biomass (3–4 weeks) showed a specific nitritation activity that was 23% lower, and a specific anammox activity that was 58% lower, compared to the reference biomass.
Acknowledgements This research was conducted with the support of the Flemish Region and the Flemish Institute for the improvement of scientific–technological research in the industry (IWT-STWW Project 980362).
References Bremner, J.M., Keeney, R.D., 1965. Steam distillation methods for determination of ammonium, nitrate and glycine. Anal. Chem. Acta 32, 485–495. Burnett, W.E., 1974. The effect of salinity variations on the activated sludge process. Water Sewage Work. 121, 37–38. Campos, J.L., Mosquera-Corral, A., Sa´nchez, M., Me´ndez, R., Lema, J.M., 2002. Nitrification in saline wastewater with high ammonia concentration in an activated sludge unit. Water Res. 36 (10), 2555–2560. Catalan, M.A.B., Wang, P.C., Matsumura, M., 1997. Nitrification performance of marine nitrifiers immobilized in polyester- and macro-porous cellulose carriers. J. Ferment. Bioeng. 84 (6), 563–571. Cecchi, F., Battistoni, P., 2003. In: Mata-Alvarez, J. (Ed.), Biomethanization of the Organic Fraction of Municipal Solid Wastes, vol. 323. IWA Publishing, London.
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Chen, M., Canelli, E., Fush, G.W., 1971. Effect of salinity on nitrification in East River. J. Water Pollut. Control Fed. 42, 2474–2481. Chen, G.H., Wong, M.T., Okabe, S., Watanabe, Y., 2003. Dynamic response of nitrifying activated sludge batch culture to increased chloride concentration. Water Res. 37 (13), 3125–3135. Dahl, C., Sund, C., Kristensen, G.H., Vredenbregt, L.H.J., 1997. Combined biological nitrification and denitrification of high-salinity wastewater. Water Sci. Technol. 36 (2–3), 345–352. Dinc- er, A.R., Kargi, F., 1999. Salt inhibition of nitrification and denitrification in saline wastewater. Environ. Technol. 20 (11), 1147–1153. Focht, D.D., Verstraete, W., 1977. Biochemical ecology of nitrification and denitrification. Adv. Microb. Ecol. 8, 135–211. Greenberg, A.E., Clesceri, L.S., Eaton, A.D. (Eds.), 1992. Standard Methods for the Examination of Water and Wastewater. American Public Health Association, American Water Works Association, Water Environment Federation, Washington DC. Hamoda, M.F., Al-Attar, I.M.S., 1995. Effects of high sodiumchloride concentrations on activated-sludge treatment. Water Sci. Technol. 31 (9), 61–72. Hunik, J.H., Meijer, H.J.G., Tramper, J., 1992. Kinetics of Nitrosomonas europaea at extreme substrate, product and salt concentrations. Appl. Microbiol. Biotechnol. 37 (6), 802–807. Hunik, J.H., Meijer, H.J.G., Tramper, J., 1993. Kinetics of Nitrobacter agilis at extreme substrate, product and salt concentrations. Appl. Microbiol. Biotechnol. 40 (2–3), 442–448. Ingram, M., 1940. The influence of sodium chloride and temperature on the endogenous respiration of Bacillus cereus. J. Gen. Physiol. 23, 773–778. Kincannon, D.F., Gaudy, A.F., 1968. Response of biological waste treatment systems to changes in salt concentrations. Biotechnol. Bioeng. 10, 483–496. Kuai, L.P., Verstraete, W., 1998. Ammonium removal by the oxygen-limited autotrophic nitrification–denitrification system. Appl. Environ. Microbiol. 64 (11), 4500–4506. Kuypers, M.M.M., Sliekers, A.O., Lavik, G., Schmid, M., Jørgensen, B.B., Kuenen, J.G., Damste, J.S.S., Strous, M., Jetten, M.S.M., 2003. Anaerobic ammonium oxidation by anammox bacteria in the Black Sea. Nature 422 (6932), 608–611. Ludzack, F.J., Noran, D.K., 1965. Tolerance of high salinities by conventional wastewater treatment processes. J. Water Pollut. Control Fed. 37, 1404–1416. Mosquera-Corral, A., Pampin, R.M., Lema, J.M., 2000. Development of a combined system for the elimination of carbon and nitrogen from waste waters in the fish canning industry. Afinidad 57 (488), 254–260. Oren, A., Gurevich, P., Malkit, A., Henis, Y., 1992. Microbial degradation of pollutants at high salt concentrations. Biodegradation 3, 387–398. Panswad, T., Anan, C., 1999a. Impact of high chloride wastewater on an anaerobic/anoxic/aerobic process with and without inoculation of chloride acclimated seeds. Water Res. 33 (5), 1165–1172.
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Panswad, T., Anan, C., 1999b. Specific oxygen, ammonia, and nitrate uptake rates of a biological nutrient removal process treating elevated salinity wastewater. Bioresour. Technol. 70 (3), 237–243. Pynaert, K., Sprengers, R., Laenen, J., Verstraete, W., 2002a. Oxygen-limited nitrification and denitrification in a labscale rotating biological contactor. Environ. Technol. 23 (3), 353–362. Pynaert, K., Wyffels, S., Sprengers, R., Boeckx, P., Van Cleemput, O., Verstraete, W., 2002b. Oxygen-limited nitrogen removal in a lab-scale rotating biological contactor treating an ammonium-rich wastewater. Water Sci. Technol. 45 (10), 357–363. Pynaert, K., Smets, B.F., Wyffels, S., Beheydt, D., Siciliano, S.D., Verstraete, W., 2003. Characterisation of an autotrophic nitrogen removing biofilm from a highly loaded labscale rotating biological contactor. Appl. Environ. Microbiol. 69 (6), 3626–3635. Pynaert, K., Smets, B.F., Beheydt, D., Verstraete, W., 2004. Start-up of autotrophic nitrogen removing reactors via sequential biocatalyst addition. Environ. Sci. Technol. 38 (4), 1228–1235.
Schenk, H., Hegemann, W., 1995. Nitrification inhibition by high salt concentrations in the aerobic biological treatment of tannery wastewater. GWF—Wasser/Abwasser 136 (9), 465–470. Semmens, M.J., Porter, P.S., 1979. Ammonium removal by ionexchange: using biologically restored regenerant. J. Water Pollut. Control Fed. 51 (12), 2928–2940. Strous, M., Heijnen, J.J., Kuenen, J.G., Jetten, M.S.M., 1998. The sequencing batch reactor as a powerful tool for the study of slowly growing anaerobic ammonium-oxidizing microorganisms. Appl. Microbiol. Biotechnol. 50, 589–596. Verstraete, W., Philips, S., 1998. Nitrification–denitrification processes and technologies in new contexts. Environ. Pollut. 102, 717–726. Vredenbregt, L.H.J., Nielsen, K., Potma, A.A., Kristensen, G.H., Sund, C., 1997. Fluid bed biological nitrification and denitrification in high salinity wastewater. Water Sci. Technol. 36 (1), 93–100. Wyffels, S., Pynaert, K., Boeckx, P., Verstraete, W., Van Cleemput, O., 2003. Identification and quantification of nitrogen removal in a rotating biological contactor by 15N tracer techniques. Water Res. 37, 1252–1259.
Water Research 39 (2005) 4512–4520 www.elsevier.com/locate/watres
Oxygen-limited autotrophic nitrification–denitrification (OLAND) in a rotating biological contactor treating high-salinity wastewater Kim Windey, Inge De Bo, Willy Verstraete Laboratory of Microbial Ecology and Technology (LabMET), Ghent University, Coupure Links 653, B-9000 Ghent, Belgium Received 9 January 2004; received in revised form 2 June 2005; accepted 8 September 2005
Abstract A lab-scale rotating biological contactor (RBC) reactor operated under OLAND conditions was slowly adapted during 178 days to increasing salt concentrations going up to 30 g NaCl L1. The reactor performed well during this experimental period. However, the removal capacity of the reactor was lower under high-salinity conditions. A removal efficiency of 84% was achieved at a N loading rate of 725 mg N L1 d1 and a salt concentration of 30 g L1. The effect of salt shock loading and adaptation to 30 g NaCl L1 on the specific nitritation and anammox activity of the biomass was investigated in short-term batch experiments. A salt shock loading of 30 g L1 caused a 43% decrease in specific nitritation activity and 96% loss of specific anammox activity compared to reference biomass (not exposed to salt). The salt-adapted biomass (3–4 weeks) showed a specific nitritation activity that was 23% lower, and a specific anammox activity that was 58% lower, compared to the reference biomass. Overall, these results demonstrate that the OLAND process can have the potential to treat ammonium-rich brines after adaptation to high salinity. r 2005 Elsevier Ltd. All rights reserved. Keywords: OLAND biofilm reactor; Nitritation; Anammox; Nitrogenous wastewater of high salinity; Salt shock loading
1. Introduction The removal of nitrogen from wastewater has become an important part of the overall treatment process due to the significant impact of nitrogen compounds on the environment and the more stringent legislation on wastewater discharges. Conventionally, nitrogen removal is performed by means of a two-step process. First, ammonium (NH+ 4 ) is oxidized to nitrite (NO2 ) or nitrate (NO3 ) by the aerobic (autotrophic) nitrification process and subsequently reduced to dinitrogen gas (N2) Corresponding author. Tel.: +32 9 264 59 76; fax: +32 9 264 62 48. E-mail address: [email protected] (W. Verstraete).
by the anaerobic (heterotrophic) denitrification process (Focht and Verstraete, 1977). Nowadays, anaerobic digestion plants are being developed that minimize energy consumption, CO2 emission and sludge production. However, these systems typically yield effluents rich in NH+ 4 and poor in biodegradable organic carbon, thereby making them less suitable for biological N removal via the conventional nitrification–denitrification process (Cecchi and Battistoni, 2003). Recently, several new N removal processes have been developed to treat these highly loaded wastewaters. The process in which NH+ is autotrophically oxidized 4 to N2 with NO as the electron acceptor under 2 oxygen-limited conditions is further referred to as oxygen-limited autotrophic nitrification denitrification
0043-1354/$ - see front matter r 2005 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2005.09.002
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(OLAND) (Kuai and Verstraete, 1998). This autotrophic process consumes 63% less oxygen and 100% less biodegradable organic carbon compared to the conventional nitrification–denitrification process and has, therefore, a lower operating cost (Verstraete and Philips, 1998). The OLAND process was first described for a mixed culture of nitrifying bacteria (Kuai and Verstraete, 1998), but was afterwards examined in more detail in a mixed community biofilm of a lab-scale rotating biological contactor (RBC) (Pynaert et al., 2002a, b, 2003, 2004). A mature OLAND biofilm under high NH+ 4 loading rate consists primarily of two major groups of bacteria responsible for autotrophic N removal. The aerobic ammonium oxidizing bacteria (AerAOB, Nitrosomonas sp.) convert NH+ 4 to NO2 with oxygen as the electron acceptor (nitritation) and the anaerobic ammonium oxidizing bacteria (AnAOB, a close relative of Kuenenia stuttgartiensis) subsequently oxidize NH+ with NO 4 2 as the electron acceptor (anammox) (Strous et al., 1998; Pynaert et al., 2003; Wyffels et al., 2003). The overall N removal stoichiometry can be summarized by the following equations (Pynaert et al., 2004): Nitritation :
Anammox :
1:32 NH3 þ 1:98 O2 ! 1:32 NO2 þ 1:32 Hþ þ 1:32 H2 O;
NH3 þ 1:32 NO2 þ Hþ ! 0:26 NO3 þ 1:02 N2 þ 2 H2 O;
N removal :
(1)
(2)
NH3 þ 0:85 O2 ! 0:11 NO3 þ 0:44 N2 þ 0:14 Hþ þ 1:43 H2 O: (3)
High nitrogen-concentrated streams can also contain large amounts of other ions like chloride (fish canning industry, wet lime–gypsum desulphurization process, regeneration liquid from ion exchange columns), sulphate (tannery wastes), etc. High saline concentrations in wastewater are known to have negative effects on nitrogen removal (Ludzack and Noran, 1965; Chen et al., 1971; Panswad and Anan, 1999b). However, Dahl et al. (1997) showed that with an adequate adaptation and slowly increased salt concentration, municipal activated sludge could be adapted to high-salinity wastewater containing up to 33 g NaCl L1. Panswad and Anan (1999a) showed an improvement in the N removal efficiency of a bioreactor treating very highsalinity wastewater (30 g NaCl L1) when seeded with NaCl acclimated cultures. Campos et al. (2002) found that adapted biomass is less sensitive to high saline concentrations. On the contrary, Catalan et al. (1997) observed that when seawater was nitrified, NO 2 accumulated even when marine genera adapted to high saline concentrations were used. Schenk and Hegemann (1995) and Dinc- er and Kargi (1999) also found that nitrite oxidizing bacteria (NOB) were more sensitive to
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salt concentration than AerAOB. Only Hunik et al. (1992, 1993) concluded the opposite. Based on the available literature, it can be concluded that the effects of NaCl on the AerAOB are not well understood and contradictory findings are reported. The effect of salt on the AnAOB has not been studied yet, although one marine AnAOB named Candidatus Scalindua sorokinii was recently discovered in the Black Sea (Kuypers et al., 2003). However, N removal under highly saline conditions was not studied (Kuypers et al., 2003). The objective of this study was to evaluate the feasibility of the OLAND process, operated in a RBC, to treat NH+ 4 -rich wastewater of high salinity. The specific nitritation and anammox activities of the AerAOB and AnAOB, respectively, were investigated before exposure to salt, at a shock loading of 30 g NaCl L1 and after an adaptation period of 3–4 weeks.
2. Materials and methods 2.1. Lab-scale RBC reactor The reactor used in this study is shown in Fig. 1. Its characteristics were as follows: basin length L ¼ 88 cm, width B ¼ 34 cm, height H ¼ 30 cm, water level at 19 cm and total water volume 50 L. Two series of 20 polyvinylchloride discs were mounted (diameter of one disc ¼ 30 cm, thickness ¼ 0.5 cm, 1 cm interspace) on a horizontal shaft, which operated at 2.5 rpm (a low value was chosen so that the O2 concentration is always less than 1 mg L1) with 58% submergence of the disc surface area. The total surface area of the wetted discs and reactor walls was 6.3 m2. In order to increase the specific surface area for biofilm growth, a 5 mm layer of reticulated polyurethane (specific surface area
Fig. 1. Schematic representation of the 50 L lab-scale RBC reactor. Discs are partially (58%) submerged in the wastewater, and rotate round a horizontal axis at 2.5 rpm. Influent is dosed at two points in the reactor and is discharged on the other side.
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71000 m2 m3, Type Filtren TM20, Recticel, Belgium) was attached to one side of every disc. This increased the initial available surface area to about 20 m2. The influent was dosed at two points in the reactor, one at the opposite end of the outlet and one halfway between the two groups of discs, to maximize biomass growth on both groups of discs. The pH in the reactor was measured regularly. 2.2. Synthetic wastewater Synthetic wastewater was fed to the RBC, prepared with tap water and containing NH+ 4 as (NH4)2SO4 and/ or urea (N source), NaHCO3 (C source and buffer) and NaCl (salt). The amounts of NH+ 4 and NaCl varied depending on the applied load. Per gram of NH+ 4 –N in the feed, 7 g of NaHCO3 was supplied. In addition, the wastewater contained 0.07 g L1 of phosphorous as KH2PO4. 2.3. Aerobic batch experiments Aerobic batch experiments were performed in 250 mL Erlenmeyer flasks with 100 mL working volume. From previous batch experiments it was observed that the standard deviation was always less than 8% (Pynaert et al., 2003). Therefore, the indicative batch experiments with reference biomass and shock-loaded biomass were not performed in triplicate. The batch experiment with adapted biomass was performed in triplicate. At the 1 beginning of the test, 0.1 g NH+ and a phosphate 4 –N L buffer with final concentrations of 3.4 g L1 KH2PO4, 4.4 g L1 K2HPO4 and 1.0 g L1 NaHCO3 were supplied to each Erlenmeyer flask. A representative sludge sample was collected from the reactor by scraping off parts of the biofilm (in the whole depth) of several discs. This sludge sample was divided in equal amounts and added to the flasks (2.570.2 g volatile suspended solids (VSS) L1). The flasks were incubated for 6 h at 2872 1C on a shaker (140 rpm) to provide oxygen. The pH was initially adjusted to 7.5770.02 with NaOH (0.1 N). During the experiment, the pH and dissolved oxygen (DO) level were monitored every 1.5 h. Samples of the liquor phase were taken with a sterile syringe and needle, filtered over a 0.45 mm filter (Millipore) and stored at 4 1C for 1 day.
experiments. After aliquots of biofilm sludge were added (2.970.1 g VSS L1), the flasks were closed with butyl rubber stoppers and aluminium caps. The initial pH of all the flasks was 8.0. To obtain anaerobic conditions, the flasks were flushed with N2 gas for 30 min in cycles of 2 min overpressure (70 kPa) and underpressure (90 kPa). Substrate solutions (NH4Cl and NaNO2) were also prepared in serum flasks, closed with butyl rubber stoppers and aluminium caps and flushed with N2 gas. An appropriate volume of each substrate solution was added to the different flasks with a sterile syringe until the final concentration of 0.1 g 1 1 NH+ and 0.1 g NO in each flask was 4 –N L 2 –N L obtained. The flasks were incubated for 6 h at 3072 1C. Samples of the liquor phase were obtained in a similar way as described for the aerobic batch tests. For the same reason as stated for the aerobic experiments, the experiments with the reference and shock-loaded biomass were not performed in triplicate. The batch experiment with the adapted biomass was performed in triplicate. 2.5. Chemical analyses NH+ 4 –N in the effluent of the RBC reactor was determined by the Kjeldahl distillation method as described by Bremner and Keeney (1965). NH+ 4 –N in the batch experiments (aerobic and anaerobic) was determined colorimetrically with Nessler reagent according to standard methods (Greenberg et al., 1992). Both NO 2 –N and NO3 –N were determined using a Methrom 761 Compact Ion Chromatograph (Methrom, Herisau, Switserland) equipped with a conductivity detector. The operational parameters were as follows: column metrosep A supp 5, eluents 1.06 g L1 Na2CO3, flow 0.7 mL min1, sample loop 20 mL. VSS and total suspended solids were determined by drying and weighing according to standard methods (Greenberg et al., 1992). The pH was determined potentiometrically with a digital, portable pH meter (Knick, portamess 751). The DO level was measured with a digital, portable DO meter (Endress-Hauser, COM 381).
3. Results 3.1. Performance of the RBC reactor
2.4. Anaerobic batch experiments Anaerobic batch experiments were performed in 120 mL serum flasks, containing 80 mL of mixed liquor (40 mL headspace). The mixed liquor contained per litre 1 g NaHCO3, 0.05 g KH2PO4, 2 mL trace element solution (Kuai and Verstraete, 1998) with or without salt (NaCl). A representative sludge sample was collected in a similar way as described for the aerobic batch
The RBC reactor performance is depicted in Fig. 2 (Panels I and II). Five periods with different N loading rates can be distinguished (R, A, B, C and D). Parameter settings of the different periods are summarized in Table 1. During the reference period (period R), the average N removal capacity was 885731 mg N L1 d1 (n ¼ 5) at a constant loading rate of 1000 mg N L1 d1. From day 12, the RBC reactor
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the RBC feed was increased from zero to 5 g NaCl L1 (Fig. 2, Panel I). A stepwise increase of 1 g NaCl L1 was applied. This salt addition had no effect on the removal capacity of the reactor. The average N removal capacity
performance was operated at a similar N loading rate and under increasing salt concentrations. From day 12 to day 25 (period A), the N loading rate was 1000 mg N L1 d1, while the salt concentration of
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Fig. 2. Performance of the RBC reactor under changing salt concentration addition. Panel I: loading rate (’) and removal capacity (W) of the reactor (mg N L1 d1) in relation to the salt concentration (g NaCl L1) (}). Panel II: influent NH+ 4 (~) concentration 1 (mg N L1), effluent NH+ 4 (’), NO2 (J) and NO3 (}) concentrations (mg N L ), and N removal percentage (%) (W). Panel III: pH (~) and DO concentrations (}) (mg O2 L1) as a function of time. Vertical arrows indicate process disturbances; on three occasions a lower loading rate was applied due to pump failures.
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10
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Fig. 2. (Continued) Table 1 Characteristics of the RBC reactor operated during periods of different N loading rates and salt concentration range Parameters
Period R
Period A
Period B
Period C
Period D
Time range (d) Feed flow rate (L d1) Hydraulic retention time (d) 1 Feed concentration (mg NH+ 4 –N L ) 1 1 Loading rate (mg N L d ) Salt level (g NaCl L1)
0–11 65 0.77 770 1000 0
12–25 65 0.77 770 1000 0-5
26–81 65 0.77 846 1100 5-10
82–135 55 0.91 769 849 10-30
136–178 55 0.91 476– 657 525–725 30
was 927710 mg N L1 d1 (n ¼ 8), representing a removal percentage of 9371% (n ¼ 8). The NH+ 4 concentration in the effluent was always below 20 mg N L1. The NO 3 concentration in the effluent had an average value of 3775 mg N L1 (n ¼ 8) (Fig. 2, Panel II). Because of this high removal percentage, the N loading rate was further increased to 1100 mg N L1 d1 from day 26 on (period B). On day 33, the salt concentration was increased to 6 g NaCl L1 (Fig. 2, Panel I) and no remarkable loss of activity was measured. The salt concentration was elevated from 6 to 10 g NaCl L1 on day 50. From day 60 on, the reactor performance decreased to a removal capacity of only 641 mg N L1 d1 on day 74 (removal efficiency ¼ 58%). During the whole of period B, no
NO 2 peaks were observed. On day 74, an elevated 1 effluent concentration of 341 mg NH+ and a pH 4 –N L of 8.5 were measured (Fig. 2, Panel III). The low removal percentage can be explained by a free ammonia inhibition. Following this, the N loading rate was temporarily lowered to 849 mg N L1 d1 on day 82 (period C) by means of a lower influent NH+ concentration and 4 by an increase of the hydraulic retention time to 0.9 day (Table 1). During the following days, the removal percentage increased from 57% to 73%. To regain the high removal capacity of the reactor, no extra salt was added to the influent from day 85 to day 91. The N removal percentage of the reactor further increased to 83%. On day 92, the RBC feed was again supplied with 10 g NaCl L1. The removal percentage of the reactor
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first decreased but returned to 86% after 1 week. From day 114 to day 125, the salt concentration was further increased up to 30 g NaCl L1 in steps of 5 g NaCl L1. When a salt concentration of 20 g NaCl L1 was achieved, the N removal capacity started decreasing until the end of period C. The effluent NH+ 4 concentration increased to 201 mg N L1 on day 135, resulting in a removal percentage of 70%. However, no NO 2 peaks were observed in the effluent. On days 126, 127 and 132, indicated by vertical arrows on Fig. 2 (Panel I and II), a lower N loading rate (424 mg N L1 d1) was supplied due to accidental interruption of the feeding. On day 136, it was decided to lower the N loading rate and to subject the biomass to a salt concentration of 30 g NaCl L1. During a time span of 43 days, the performance at increasing loading rates between 525 and 725 mg N L1 d1 was investigated (Fig. 2, Panels I and II). At a N loading rate of 525 mg N L1 d1, a removal percentage of 8170.5% (n ¼ 5) was obtained. The removal capacity of the reactor was however 52% lower compared to the reference period before salt addition to the reactor. The increase of the influent NH+ 4 –N concentration, but temporary decrease of the flow rate, resulted in a N loading rate of 571 mg N L1 d1 and a removal percentage of 8376% (n ¼ 2). An increase of the N loading rate to 625 mg N L1 d1 resulted in a decrease of the N removal percentage to 7479% (n ¼ 6) and hence a N removal capacity of 463742 mg N L1 d1 (n ¼ 6). At a N loading rate of 725 mg N L1 d1 one can distinguish two periods. From day 144 to day 171, a N removal capacity of 485735 mg N L1 d1 (n ¼ 6) or a removal percentage of 6775% (n ¼ 6) was measured. The highest N
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removal capacity, 609715 mg N L1 d1 (n ¼ 5), was achieved from day 172 to day 178, which is 31% lower compared to the reference period before salt addition to the reactor. 3.2. Specific nitritation activity of OLAND biomass Short-term aerobic batch experiments were performed with an OLAND biomass of periods R and D. The biomass of period R was used to measure the reference specific nitritation activity of the AerAOB (no salt) and the specific activity when these AerAOB are exposed to a salt shock loading of 30 g NaCl L1 (Fig. 3: reference biomass R; shock-loaded biomass S). After an exposure period of 23 days to 30 g NaCl L1 (period D), the specific activity of the AerAOB was also measured (Fig. 3: adapted biomass A). During the experiment, the pH decreased with an average value of 0.1370.06. The DO level was always above 3.3 mg O2 L1. The specific nitritation activity of the reference biomass was 60 mg 1 NH+ VSS d1, whereas a decrease of 43% was 4 –N g observed for the shock-loaded biomass, resulting in a 1 specific nitritation activity of 34 mg NH+ 4 –N g 1 VSS d . The adapted biomass showed a specific 1 nitritation activity of 4674 mg NH+ VSS d1. 4 –N g Hence, this was only 24% lower than the specific activity of the reference biomass. In the aerobic batch experiments with reference 1 biomass, a maximum of 7 mg NO was detected 3 –N L (data not shown). When salt was added to shockloaded respectively to adapted biomass, no NO 3 was detected in the liquor phase (detection limit of 0.01 mg 1 NO 3 –N L ).
Specific activity (mg NH4+-N g-1 VSS d-1)
80 70 60 50 40 30 20 10 0
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Fig. 3. Overview of the specific nitritation activity of the AerAOB (light grey bars) and the specific anammox activity of the AnAOB (dark grey bars) determined in batch experiments. R ¼ reference biomass (period R), S ¼ biomass exposed to a salt shock loading of 30 g NaCl L1 (period R), A ¼ biomass adapted to salt (period D) (n ¼ 3).
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3.3. Specific anammox activity of OLAND biomass Anaerobic batch experiments were performed in order to evaluate the specific anammox activity of the AnAOB in the reference, salt shock-loaded and salt-adapted biomass as defined above. The reference specific 1 anammox activity was 77 mg NH+ VSS d1 4 –N g 1 (Fig. 3, R). A shock loading of 30 g NaCl L caused a decrease in specific activity of 95% (Fig. 3, S), resulting 1 in a specific anammox activity of 3 mg NH+ 4 –N g 1 VSS d . The salt-adapted AnAOB (Fig. 3, A) had a 1 specific activity of 3272 mg NH+ VSS d1, 4 –N g which means that adaptation of the AnAOB to salt had occurred. However, the specific activity was still 59% lower than the reference.
4. Discussion The reactor performance was not negatively affected by exposing the OLAND biofilm to salt concentrations up to 6 g NaCl L1. A stimulation of the microbial activity of various other bacterial species at low salt concentrations has been reported in the literature (Chen et al., 2003; Hamoda and Al–Attar, 1995; Ingram, 1940). Chen et al. (2003) reported that a salt concentration below 4.12 g NaCl L1 was in favour of the nitrification rate, but exceeding this level resulted in a decrease. Hamoda and Al-Attar (1995) also stated a stimulation of the activity of microorganisms developed in a freshwater environment below this level. Early studies with Bacillus cereus indicated that endogenous respiration rate of the culture increased with NaCl concentrations up to 10 g L1, but decreased above this level (Ingram, 1940). An explanation for the fluctuations during the period 60–113 days could be that at a salt concentration of 10 g NaCl L1, the AerAOB became initially inhibited. However, after an adaptation period with the higher salt concentration, saline-resistant AerAOB may develop. The imbalance between the amount of freshwater and saline-resistant AerAOB can also be an explanation for the decreased reactor performance at salt levels above 20 g NaCl L1 in period C. Vredenbregt et al. (1997) considered 16.5 g NaCl L1 as a critical level for a freshwater nitrifying culture. Chen et al. (2003) reported a shift in the dominant species of AerAOB from non-saline-resistant species, such as Nitrosomonas europaea lineage and Nitrosomonas eutropha, to saline-resistant species such as Nitrosococcus mobilis lineage when the salt concentration was increased from 16.5 to 30 g NaCl L1. Rapid shifts in salt concentration were reported to have more adverse effects than gradual shifts (Burnett, 1974; Oren et al., 1992; Chen et al., 2003). However, Kincannon and Gaudy (1968) found that a decrease in salt concentration caused more severe negative effects on microorganisms
than an increase in salt concentration. They stated that rapid changes in salt concentration caused immediate release of cellular constituents resulting in an increase in soluble organic carbon. The OLAND process operated at 30 g NaCl L1 can achieve similar N removal percentages as when operated without salt by decreasing the N loading rate with 28% to a value of 725 mg N L1 d1. Further work on the microbial ecology would be necessary to verify whether the OLAND microbial community has undergone a physiological adaptation and/or there has been a substantial shift in microbial species. The halophilic conditions in the RBC reactor were an advantage for the OLAND process, since NOB are more negatively affected by high salt concentrations (Schenk and Hegemann, 1995; Dinc- er and Kargi, 1999). The NOB are unwanted in an OLAND reactor because they oxidize NO 2 to NO3 when sufficient oxygen is present and therefore result in an increase of N components in the effluent. In the aerobic batch experiments with OLAND biomass, no NO 3 was detected when salt was 1 supplied, whereas a maximum of 7 mg NO was 3 –N L detected without salt addition. Hence, the few NOB that were present were further repressed when the reactor feed was supplied with NaCl. During the whole period of salt addition to the reactor, no NO peaks were observed. Hence, 2 the anaerobic ammonium oxidation activity of the OLAND biofilm was not affected by the increase of the salt level. The highest value measured from period R 1 to period D was 16 mg NO on day 18 (period 2 –N L A). This was probably due to the shortage in electron donor, because the NH+ 4 concentration in the effluent 1 was only 3 mg NH+ 4 –N L . Consequently, the ratelimiting step in the OLAND process at elevated salt levels seems to be nitritation because all NO 2 produced was immediately used to further oxidize NH+ 4 anaerobically to N2. However, from the batch experiments it can be concluded that the AnAOB are more influenced by high salt concentrations than the AerAOB, when exposed to a salt shock loading, as well as after an adaptation period of approximately 1 month. It seems that the AnAOB, when encapsulated in a thick biofilm, were better able to withstand high salt levels. It therefore appears that, overall, both groups of bacteria can perform under saline conditions, but further work is necessary to elucidate the physiological basis of their respective salt tolerance mechanisms. Overall, these data indicate that the OLAND process, with careful monitoring of the DO concentration, pH, NH+ 4 –N and NO2 –N concentration within a few months, can become adapted to handle NH+ 4 -rich saline solutions, as often produced by fish canning industries (Mosquera-Corral et al., 2000), the wet lime–gypsum desulphurization process (Vredenbregt et al., 1997) and
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regeneration of ion exchange columns (Semmens and Porter, 1979).
5. Conclusions 1. An OLAND biofilm reactor was exposed to salinity levels up to 30 g NaCl L1 during a time span of 178 days. 2. Autotrophic nitrogen removal of high-salinity wastewater with high nitrogen concentrations can be effectively achieved at loading rates up to 725 mg N L1 d1, reaching N removal efficiencies of 84%. The N removal capacity was 31% lower at a salt level of 30 g NaCl L1 compared to the reference period without salt addition. 3. Short-term batch experiments (both aerobic and anaerobic) were performed to investigate the salt shock loading and adaptation to salt levels of 30 g NaCl L1. A salt shock loading of 30 g L1 caused a 43% decrease in specific nitritation activity and 96% loss of specific anammox activity compared to reference biomass (not exposed to salt). The saltexposed biomass (3–4 weeks) showed a specific nitritation activity that was 23% lower, and a specific anammox activity that was 58% lower, compared to the reference biomass.
Acknowledgements This research was conducted with the support of the Flemish Region and the Flemish Institute for the improvement of scientific–technological research in the industry (IWT-STWW Project 980362).
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