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Advanced nitrogen removal from landfill leachate via Anammox system based on Sequencing Biofilm Batch Reactor (SBBR): Effective protection of biofilm
Accepted Manuscript Advanced nitrogen removal from landfill leachate via Anammox system based on Sequencing Biofilm Batch Reactor (SBBR): Effective protection of biofilm Lei Miao, Shuying Wang, Tianhao Cao, Yongzhen Peng, Man Zhang, Zhaoyuan Liu PII: DOI: Reference:
S0960-8524(16)30955-5 http://dx.doi.org/10.1016/j.biortech.2016.06.131 BITE 16753
To appear in:
Bioresource Technology
Received Date: Revised Date: Accepted Date:
11 April 2016 21 June 2016 24 June 2016
Please cite this article as: Miao, L., Wang, S., Cao, T., Peng, Y., Zhang, M., Liu, Z., Advanced nitrogen removal from landfill leachate via Anammox system based on Sequencing Biofilm Batch Reactor (SBBR): Effective protection of biofilm, Bioresource Technology (2016), doi: http://dx.doi.org/10.1016/j.biortech.2016.06.131
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Advanced nitrogen removal from landfill leachate via Anammox system based on Sequencing Biofilm Batch Reactor (SBBR): Effective protection of biofilm Lei Miaoa, Shuying Wanga,*, Tianhao Caoa, Yongzhen Penga,, Man Zhanga, Zhaoyuan Liua a,
*Engineering Research Center of Beijing, Beijing University of Technology, Beijing, PR China
Abstract High levels of organics negatively affect Anammox for treating landfill leachate. To enhance the ability of Anammox to survive against adverse environments, a lab-scale two-stage Anammox system using a Sequencing Biofilm Batch Reactor was applied to treat mature landfill leachate under 35 °C. Over 107 days, with influent total nitrogen (TN) and chemical oxygen demand (COD) concentrations of 3000±100 and 3000±100 mg/L, effluent TN was below 20 mg/L. For extracellular polymeric substance (EPS) of Anammox, slime-EPS and loosely-bound-EPS of floccules were both higher than biofilm, while tight-bound-EPS of biofilm was significantly higher, contributing to biofilm formation. Quantitative microbial analysis showed that as influent COD increased, Anammox gene ratios of biofilm increased from 1.34% to 13.28%; the gene ratios of floccule first increased, then decreased to 3.88%. This indicated that Anammox and heterotrophic bacteria could coexist because of the biofilm, leading to stable nitrogen removal performance, even when organics were present. Keywords: Anammox; landfill leachate; biofilm; EPS; PCR; *Corresponding author Tel/Fax +86 10 67392627
1
Email: [email protected] (S. Wang); [email protected] (L. Miao)
1 Introduction Landfill leachate is a kind of wastewater, containing large amounts of organics and inorganics, with a particularly high ammonia concentration (Renou et al., 2008). Landfill leachate treatments are usually classified as physical/chemical and biological treatments (Hu et al., 2016; Renou et al., 2008). The physical/chemical treatment (e.g. coagulation-flocculation) is usually used as the pre-treatment or advanced treatment step due to its high cost and secondary pollution (Zhang et al., 2016). Biological processes, based on active sludge, are usually used to remove nitrogen and organics from the leachate due to their simplicity and cost-effectiveness (Zhu et al., 2013). For early-age or middle-age leachate, a conventional nitrification/denitrification process or modified process (such as the process coupling of two-stage up-flow anaerobic sludge bed (UASB) and anoxic/oxic reactor (A/O), or the endogenous denitrification process), is used to remove nitrogen and organics because of the high carbon to nitrogen (C/N) ratio (Peng et al., 2008; Miao et al., 2015). However, it is difficult to achieve high nitrogen removal from mature leachate with a low C/N ratio (usually <3), because the conventional nitrification/denitrification process requires additional external carbon sources. This requirement makes the treatment systems complicated and expensive (Kulikowska et al., 2008). Anaerobic ammonium oxidation (Anammox), is a novel cost-effective process with great potential (Mulder et al., 1995; Kartal et al., 2010). Compared with the conventional nitrification/denitrification process, Anammox does not need external 2
carbon sources, and the power consumption for Anammox can be reduced substantially (Shi, et al., 2013). Therefore, it can effectively treat mature landfill leachate (Anfruns, et al., 2013). However, Anammox cannot work alone, because it requires nitrite substrate supplied by nitritation. As such, Anammox is usually combined with nitritation to treat wastewater. Nevertheless, some factors adversely affect Anammox, including high nitrite and organic concentrations. For example, Strous et al. (1999) demonstrated that Anammox was inhibited by nitrite concentrations up to 100 mg/L. Thus, it is important to avoid feeding in a short time for Anammox process treating the wastewater with high substrate content. Further, organics can significantly influence Anammox (Tang et al., 2010). Molinuevo et al. (2009) observed that a COD concentration up to 292 mg/L could completely inhibit Anammox. Ni et al. (2012) studied artificial wastewater, finding that a low COD concentration did not affect nitrogen removal; however, a high COD concentration did suppress Anammox activity and reduced the bacteria population when the COD exceeded 400 mg/L. Miao et al. (2014) used a three-stage Anammox system to treat landfill leachate, finding that Anammox was not significantly inhibited when the biodegradable COD was less than 150 mg/L. Therefore, applying Anammox requires controlling the influent COD concentration. Furthermore, for Anammox to be effective in treating real wastewater, the system must make full use of the organics in the wastewater and enhance nitrogen removal performance through the synergy between Anammox and denitrification. Because Anammox bacteria grow slowly, it is important to enrich the Anammox 3
bacteria during the start-up of the Anammox system. Compared with the activated sludge process, the biofilm technology associated with Anammox covers a smaller footprint, and has more microbial populations (Van Hulle et al., 2010), a stronger impact resistance (Nicolella et al., 2000; Han et al., 2012) (such as oxygen and high nitrite concentration), and a more stable ecological system when treating wastewater. In addition, biofilm systems are more efficient and more widely applicable in nitrogen removal systems than activated sludge (Tsushima et al., 2007; van der Star et al., 2007). Therefore, in this study a two-stage Sequencing Biofilm Batch Reactor (SBBR) system coupled with biofilm was applied to treat mature landfill leachate and achieve a fast start-up. The extracellular polymeric substances (EPS) and microbial population of the Anammox SBR during different phases were measured to assess microbiological changes. 2 Materials and Methods 2.1 Experimental setup and operational procedure The two-stage SBBR system used in this study, was comprised of a nitritation SBR (SBRni) and Anammox SBBR (ASBBR). The system was built using polymethyl methacrylate with a total working volume capacity of 20 L, divided into equal parts of 10 L each for the two system components (SBRni and ASBBR). The SBRni was equipped with pH meters, mechanical stirrers, and air diffusers; the ASBBR was equipped with a pH meter and a mechanical stirrer. Aeration intensity was maintained at 100 L/h during the aeration period. The operational temperatures for the SBRni and ASBBR were maintained at 25 and 35 °C, respectively, using a 4
temperature controller. Figure 1 Table 1 presents the experimental procedure. There were four phases during the start-up and steady operation of the two-stage SBBR system. During Phase 1, synthetic wastewater was used. Carriers were added to the ASBBR to form the biofilm and promote biofilm culturing. The carriers were cubic sponges (1 cm×1 cm×1 cm) made of polyurethane, and the packing ratio was 10%. During Phase 2 and 3, the SBRni effluent and real mature landfill leachate, respectively, were gradually added into the ASBBR. During Phase 4, the ASBBR was in steady operation; the ASBBR influent was the mixed effluent of SBRni and raw mature landfill leachate with a ratio of NO2−-N/NH4+-N of 1.3-1.5 (Figure 1). Table 1 In the two-stage SBBR system, the SBRni was operated using the traditional approach: filling-aeration-settle-decant. ASBBR was operated using the modified approach: continuous feeding (5 h) and stirring until the reaction was completed. The terminal point of Anammox was determined by the pH profile. The operational cycle time of each stage was controlled at 24 hours. The exchange volumetric rates of SBRni and ASBBR were both 50%. 2.2 Influent and seed sludge During the biofilm formation phase (Phase 1), the ASBBR system was fed with synthetic wastewater, containing KH2PO4 (10 mg/L), CaCl2·2H2O (5.6 mg/L), MgSO4·7H2O (300 mg/L), KHCO3 (1250 mg/L), trace element solution I (EDTA 5
5000 mg/L and FeSO4 5000 mg/L), and trace element solution II (EDTA 1000 mg/L, H3BO4 14 mg/L, MnCl2·4H2O 990 mg/L, CuSO4·5H2O 250 mg/L, ZnSO4·7H2O 430 mg/L, NiCl2·6H2O 190 mg/L, NaSeO4·10H2O 210 mg/L, NaMoO4·2H2O 220 mg/L) (Miao et al., 2014). NaNO2 and NH4Cl solutions were added to supply nitrite and ammonium for Anammox activities. Influent pH was controlled at 7.5 ± 0.2. After start-up, mature landfill leachate collected from the Liulitun Municipal Solid Waste (MSW) Sanitation Landfill Site (Beijing, China) was used as the feeding solution (Table 2). Table 2 The mixed liquor suspended solid (MLSS) of the SBRni was 3000±100 mg/L. The excess sludge collected from the existing lab-scale Anammox SBR, treating synthetic wastewater, was used as the inoculums for the ASBBR. The inoculums of ASBBR were the mixture of granule and floccules; the MLSS was 2000±100 mg/L after inoculation. 2.3 Analytical methods The pH and temperature were monitored using a pH/Oxi 340i analyzer (WTW Company, Germany). MLSS, NH4+-N, NO3−-N, NO2--N, and COD were measured using the standard methods (APHA, 1995). 2.4 EPS extraction and analysis EPS is typically categorized as slime EPS (S-EPS), loosely-bound EPS (LB-EPS), and tight-bound EPS (TB-EPS). The sludge samples of biofilm and floccule in the Anammox SBBR were collected during the steady state phase. The sludge was 6
directly centrifuged at 2000 g for 15 min, and the supernatant was collected as the S-EPS. The residual was re-suspended to its original volume using a buffer solution with a pH of 7, containing 1.3 mM Na3PO4, 2.7 mM NaH2PO4, 6 mM NaCl, and 0.7 mM KCl. The suspension was centrifuged again at 5000 g for 15 min, and the supernatant was collected as the LB-EPS. Then, the residual was re-suspended again with a buffer solution to the original volume; the solution was then treated using ultrasound at 20 kHz and 480 W for 2 min. The extracted solution was centrifuged at 20,000 g for 20 min, and the supernatant was collected as the TB-EPS. Protein (PN), polysaccharose (PC), and nucleic acid were measured using Lowry-Folin methods, Phenol-Sulfuric acid method, and diphenylamine methods, respectively (Li et al., 2014). The three-dimensional excitation-emission matrix (3D-EEM) spectra of all EPS samples were measured using an EEM fluorescence spectrophotometer (LS-55, Perkin Elmer, USA). In this study, the 3D-EEM spectra were collected with corresponding scanning excitation spectra from 220 to 550 nm at 1 nm increments by varying the emission wavelength from 200 to 550 nm at 10 nm increments. 2.5 Calculations Equations (1) and (2) were used to calculate the nitrogen load rate (NLR) (kgN/m3·d) and nitrogen removal rate (NRR) (kgN/m3·d), respectively. mgN
NRR gVSS·h =
mg
-
mg
-
mg
mg
-
mg
-
mg
NH4+ –Ninf L +NO2 –Ninf L +NO3 –Ninf L -NH4+ –Neff L -NO2 –Neff L -NO3 –Neff L (1) g th ∙MLVSSL
7
#$ !" & ( % ·d =
)*+ – ,-. %$/ + 012 – ,-. %$/ + 0&2 – ,-. %$/ × 4567 × 102: 1 4%& ∙ ;
<
2
where NH4+-Ninf, NO2--Ninf, NO3--Ninf and NH4+-Neff, NO2--Neff, NO3--Neff are the nitrogen concentrations of ASBBR influent and effluent (mg/L), respectively; Vinf is the volume of ASBBR influent (L); V is the working volume of ASBBR (L); MLVSS is the mixed liquor volatile suspended solid concentration of ASBBR (gVSS); and t is the reaction time of ASBBR (h). 2.6 DNA isolation and PCR Anammox bacterial populations were quantified and correlated with nitrogen removal efficiency in the Anammox system. The biofilm and floccule activated sludge samples from the ASBBR were collected during Phases 1, 2, 3, and 4. The DNA was extracted from the sludge sample using a FastDNA SPIN Kit for soil (Bio 101, Vista, CA). A nested PCR assay was conducted to amplify the Anammox 16S rRNA gene. The all-bacterial amplification was carried out using a 341f-543r (Koike et al., 2007) with a thermal profile of 95 °C for 10 min, followed by 25 cycles of 30 s at 95 °C, 30 s at 55 °C, and 30 s at 72 °C. The Anammox amplification was carried out using an Amx368f-Amx820r primers31 (Schmid et al., 2005) with a thermal profile of 96 °C for 10 min followed by 25 cycles of 1 min at 96 °C, 1 min at 52 °C, and 1 min at 72 °C. The PCR reagents (50 µL) contained 10× PCR buffer (5 µL), dNTPs (4 µL, 2.5 mmol/L), Ex Taq polymerase (0.5 µL, 2.5U, Takara, Dalian, China), each primer (1 µL, 10 mmol /L), DNA template (1.25 µL, 1–10 ng), and ddH2O (37.25 µL) (Zhu et al., 2011). 8
2.7 Quantitative PCR The abundances of all the bacterial and Anammox DNAs were determined by real-time PCR using an MX3000P Real-Time PCR system (Stratagene, La Jolla, CA) equipped with the fluorescent dye SYBR-Green approach. The primers for all the bacteria and Anammox in the real-time PCR were 341f-543r and Amx368f-Amx820r. The amplification was performed in the 20 µL reaction mixtures, containing SYBR Green exTaq (10 µL, Takara, Dalian, China), ROX Reference Dye 50 (0.3 µL), each primer (0.3 µL, 10 mmol/L), and the DNA template (2 µL, 1–10 ng). The program contains the following steps: 3 min at 95 °C, followed by 40 cycles of 30 s at 95 °C, 30 s at 59 °C, and 30 s at 72 °C.
3 Results and Discussion 3.1 Performance of the two-stage SBBR system In the two-stage SBBR system, nitrite accumulation remained above 95% in SBRni, which treated the mature landfill leachate. The start-up of ASBBR lasted for 107 days (Phase 1, 2, and 3). After Phase 1, the SBRni was combined with the ASBBR to test the two stage process. 3.1.1 Start-up of the two-stage Anammox system Phase 1 lasted 33 days; carriers were added into the ASBBR to form the biofilm and promote biofilm culturing. During the first 14 days of Phase 1, the influent total inorganic nitrogen (TIN) was maintained at 1500±200 mg/L (Figure 2). The influent nitrite concentration reached 850±50 mg/L, which could significantly inhibit 9
Anammox bacteria. However, the ASBBR feeding mode was then changed to a 5 h continuous feeding mode, reducing the level at which high nitrite concentrations inhibited Anammox bacteria. The NO2-N/NH4-N remained at 1.30±0.5 (Figure 2), similar to the theoretical value of 1.32. The NO3-N/NH4-N was 0.19±0.2 (Figure 2), lower than the theoretical value (0.26). This is because microorganisms secreted EPS; denitrifying bacteria then used this EPS as a carbon source to remove nitrate (Ni et al., 2012). Figure 2 The effluent TIN concentration remained at 130±20 mg/L and TIN removal exceeded 90% (Figure 2), indicating good performance of the ASBBR. Therefore, starting on day 15 th, the influent TIN concentration was increased to 1800±100 mg/L, and on day 26th, the concentration was increased further to 2200±200 mg/L (Figure 2). The TIN removal held above 90%, and the NO2-N/NH4-N and NO3-N/NH4-N remained at 1.40±0.5 and 0.20±0.2, respectively. These levels are similar to the levels seen during the first 14 days (Figure 2). The reaction time was then prolonged to 15 h and 16 h, respectively (not shown in Figure 2). Finally, the nitrogen loading rate (NLR) and nitrogen removal rate (NRR) reached 1.65 kgN/(m3·d) and 39.06 mgN/(gVSS·h), respectively. The sludge concentration in the ASBBR was also measured. The MLSS of the flocculent sludge and biofilm reached approximately 3900±100 mg/L and 570±50 mg/L, respectively. These results indicate the biofilm had been successfully cultured in the carriers. Figure 3 10
Starting on day 34th (Phase 2), SBRni effluent was gradually added into the ASBBR, thus providing influent nitrite to the ASBBR. To help the Anammox bacteria adapt to the landfill leachate, only 10% of the ASBBR influent nitrite initially came from the SBRni effluent. This proportion was then gradually increased to 20%, 50% and 100% (Figure 3). Because the SBRni effluent nitrite concentration was limited (<800 mg/L), the influent TIN concentration of ASBBR decreased as the ratio of SBRni effluent increased. At the end of Phase 2, the influent TIN concentration of the ASBBR was maintained at 800 mg/L, with an effluent TIN concentration of 100 mg/L. There were almost no biodegradable organics in the SBRni effluent after aeration (influent COD of ASBBR was only 200 mg/L). Therefore, no significant denitrification was observed and Anammox was not inhibited in Phase 2. Compared with Phase 1, the NO2-N/NH4-N and NO3-N/NH4-N remained stable at 1.35±0.5 and 0.18±0.2, respectively. However, the reaction time increased from 14 h to 20 h, because the Anammox bacteria were affected by the landfill leachate. At the end of Phase 2, the reaction time returned to 14 h, suggesting the Anammox bacteria had adapted to the SBRni effluent. During Phase 2, the NLR and NRR decreased to 0.69 kgN/(m3·d) and 14.58 mgN/(gVSS·h), respectively, along with the decrease in influent TIN concentration. Furthermore, during Phase 2, the effluent COD of ASBBR was a little higher than the influent COD concentration. This may be because as the landfill leachate was added, some Anammox bacteria broke apart and some microorganisms secreted EPS. Figure 4 11
After the ASBBR adapted to the incoming SBRni effluent, raw mature landfill leachate was gradually added to the ASBBR (Phase 3, from day 76th). To reduce the impact of raw landfill leachate on the Anammox bacteria, only 50% of the ammonia in the ASBBR influent was supplied by the raw mature landfill leachate. When the raw landfill leachate was added, the influent organics also increased. This led to a gradual decrease in the effluent TIN of ASBBR. Until day 93 rd, the effluent TIN of ASBBR decreased to 40 mg/L (Figure 4). The NO2-N/NH4-N remained unchanged (1.35±0.5); in contrast, the NO3-N/NH4-N fell significantly, to approximately 0.1. This is because while the biodegradable organic concentration in the raw landfill leachate was low (<300 mg/L), the denitrifying bacteria also used the organics to reduce nitrate to N2. From day 94th, 100% of the ammonia in the ASBBR influent was supplied by the raw mature landfill leachate. With 100% of landfill leachate added, the effluent TN dropped further, reaching 20 mg/L (Figure 4). Compared with Phase 2, the TN removal efficiency improved significantly (reached above 95%). The NO3-N/NH4-N fell below 0.1, suggesting that the Anammox was not significantly inhibited by the influent organics. In contrast, nitrogen removal was improved by the appropriate denitrification. During Phase 3, the reaction time initially only increased 2 h, and then returned to 14 h (Figure 4). This suggests that the Anammox bacteria were enriched in the added carriers, enhancing the bacteria’s resistance to the adverse environment. Furthermore, the Anammox gene copies numbers and ratios (Table 4) also supported the conclusion above. The biofilm carriers helped Anammox bacteria coexist with the 12
denitrifiers, improving nitrogen removal through the combination of Anammox and denitrification. Finally, the MLSS of the floccules and biofilm in ASBBR reached approximately 4100±100 mg/L and 630±50 mg/L, respectively (The pictures of floccule sludge and biofilm were shown in Figure S1). The NLR and NRR were maintained at 0.51 kgN/(m3·d) and 12.08 mgN/(gVSS·h), respectively, with the TN removal of above 95%. Anfruns et al. (2013) used partial nitritation coupled with Anammox SBR process to treat landfill leachate, an NLR of 0.32 kgN/(m3·d) with only nitrogen removal of 84% was achieved. Miao et al. (2014) applied three-stage Anammox SBR system to treat landfill leachate; only about 90% of TN was removed. Compared with other studies which applied Anammox to treat real wastewater, the two-stage Anammox SBBR system achieved a better performance of TN removal. Furthermore, the difference in the value between influent and effluent COD of ASBBR in this study was maintained at 200 mg/L (Figure 4). Molinuevo et al. (2009) found
that
when
COD
concentration
of
pig
manure
effluent
after
UASB-post-digestion reached above 237 mg/L, Anammox was inhibited and the denitrification became the dominant nitrogen removal process, which was similar with the conclusion in this study. When comparing with the Anammox process coupled with granular sludge (the Anammox system collapsed when influent biodegradable COD exceeded 150 mg/L (Miao et al., 2014)), the Anammox system with biofilm tolerated higher COD concentrations. This could indicate that the added carriers enriched the Anammox bacteria and supported coexistence between Anammox bacteria and denitrifiers. The conclusion could also be demonstrated using 13
the quantitative microbial analysis of the Anammox process (discussed in Part 3.3, Table 4)). Compared with the Anammox gene numbers of the inoculum (9.45×10 7 gene copies/(g·dry·sludge), Table 4), the Anammox gene numbers in the biofilm finally reached 2.63×109 gene copies/(g·dry· sludge) (Table 4), indicating the Anammox bacteria have been enriched. Therefore, to remove nitrogen from landfill leachate, adding the biofilm carriers into the Anammox system helped stabilize the system and improved nitrogen removal. 3.1.2 Anammox SBBR performance in a typical cycle during the steady phase After the system was stable, ASBBR performance in a typical cycle during steady phase (Phase 4) was monitored (Figure 5). The ASBBR was set in a continuous feeding mode for 5 hours, reducing the degree to which the influent’s high nitrite concentration inhibited the Anammox bacteria (the inhibition level was about 100 mg/L, Strous et al. 1999). Therefore, the feeding was accompanied by an Anammox reaction, and the average hourly TN removal in remained above 100 mg (Figure 5). The pH during the feeding time decreased slightly (from 8.11 to 7.89), also supporting high activity by the Anammox bacteria. Additionally, the nitrate in the ASBBR only increased slightly as the COD decreased during the first 5 hours. This indicates that denitrification was involved in the nitrogen removal process, as the influent’s biodegradable organics increased gradually with feeding. After feeding, the maximum ammonia and nitrite concentrations of ASBBR reached approximately 79 mg/L and 92 mg/L, respectively, suggesting the Anammox bacteria were not significantly inhibited. Furthermore, the 14
maximum ammonia and nitrite concentrations also led to a high driving force of reaction. Therefore, the TN removal in the first hour after feeding reached its maximum (about 386 mg) (Figure 5). During the following 6 hours, the hourly TN removal rate changed little, remaining at approximately 200 mg/L until the 12 th h. Between the 5th hour and 12th h, the nitrate concentration also varied little, suggesting the Anammox was accompanied with denitrification. Until the 12th h, the nitrite concentration decreased to 2 mg/L, with a remaining ammonia level of 8 mg/L (Figure 5). Figure 5 Over the following 2 hours, the nitrite concentration remained stable, while the ammonia and nitrate concentrations decreased. This is because the nitrate was reduced to nitrite, and was removed with the remaining ammonia by the Anammox. Because only a little ammonia and nitrate remained during this period, the hourly TN removal decreased to 100 mg. Until the 14th h, the ammonia concentration was close to 0 mg/L; meanwhile, a change point occurred in the pH profile. During the reaction time after feeding (Figure 5), the pH continued to rise until the 14th h, when it began to decline. This signaled the completion of Anammox. Therefore, the pH value can be used to identify the terminal point of Anammox. The difference of 200 mg/L between influent and effluent COD enhanced the denitrification and led to a final decrease in effluent nitrate concentration to 12 mg/L. 3.2 Analysis of EPS from Anammox SBBR EPS is one kind of macromolecule organics excreted by microorganisms; they 15
include carbohydrates, proteins, nucleic acids, and humic acids. The presence of EPS supports the survival of many microorganisms under different circumstances (Liu et al., 2004). For example, it could enrich the nutriment in the external environment for microorganisms and promote the formation and stability of microbial community structure (Flemming et al., 2010). In this study, EPS from sludge floccules and the biofilm in the ASBBR during Phase 4 were extracted and analyzed (Table 3). Compared with floccules in the ASBBR, the total EPS content of biofilm was higher. For S-EPS and LB-EPS, the EPS of the floccules exceeded that of the biofilm; however, TB-EPS of biofilm was much higher than the TB-EPS of floccules (225.77 mg/gVSS and 117.71 mg/gVSS, respectively). This contributed to biofilm formation. An analysis of EPS composition showed that carbohydrates made up a significant proportion in both floccules and biofilm (73.38% and 71.03%, respectively), followed by protein and nucleic acid. This finding was consistent with Jin et al. (2014). The large amount of carbohydrate found in S-EPS, LB-EPS, and TB-EPS facilitated cell adhesion on the biofilm surface, supporting its formation (Zhang et al., 2014). Protein was mainly found in S-EPS and TB-EPS, particularly in the S-EPS of floccule and TB-EPS of biofilm. This was also similar to the results of Zhang et al., (2014). These results suggest that extracellular proteins varied between the floccule and biofilm. The nucleic acid content of EPS was the lowest, below 10%, in both floccule and biofilm. Table 3 Researchers have proposed that the ratio of protein and carbohydrate (PN/PC 16
ratio) in EPS may be a good indicator of sludge stability: the higher the PN/PC ratio, the poorer the stability (Jin et al., 2013). The PN/PC ratio of the floccule was higher than the biofilm for S-EPS, nevertheless, the opposite result was seen for TB-EPS. Overall, compared with an Anammox system with synthetic wastewater (Jin et al., 2014), the PN/PC ratio of EPS in ASBBR was lower, for the floccule and the biofilm. This suggests that the ASBBR had good sludge stability, possibly due to the acclimation of Anammox system to real landfill leachate. Figure 6 Figure 6 shows the 3D-EEM fluorescence spectra of each EPS fractions from floccule and biofilm in the ASBBR. Two peaks were clearly observed from the 3D-EEM fluorescence spectra of the five samples (SEPS-floccule, LBEPS- floccule, LBEPS-biofilm, TBEPS- floccule, and TBEPS-biofilm), except the SEPS-biofilm. However, the positions of the peaks varied across all samples, which could indicate different substances in EPS (Chen et al., 2003). Peak A was observed at the excitation/emission wavelengths (Ex/Em) of 340-345 nm/430-435nm, indicating a humic acid-like substance. Peak B was identified at the Ex/Em of 230-240 nm/455-460 nm, which was considered to be a fulvic acid-like substance (Figure 6 (a) and (b)). The humic acid-like substance and fulvic acid-like substance detected in SEPS may have come from the raw mature leachate. Peak C was identified at the Ex/Em of 220-230 nm/355-360 nm, also considered to be a fulvic acid-like substance (Figure 6 (c) and (d)). It can be concluded that there were no large differences between the substances 17
in S-EPS and LB-EPS of the floccule and biofilm. In contrast, the positions of the peaks changed significantly in TB-EPS of the floccule and biofilm. Peak A, B and C were not observed, while Peak D and E were identified at the Ex/Em of 270-280 nm/345-350 nm and the Ex/Em of 220-230 nm/345-350 nm, respectively (Figure 6 (e) and (f)). Peak D was related to soluble microbial byproduct-like substances, while Peak E was related to aromatic proteins-like substances. It suggested that TB-EPS, which was closest to the cell, contains many protein-like substances and byproduct secreted by microorganisms. As such, it plays an important role in protecting the microorganisms.
3.3 Quantitative microbial analysis of Anammox A quantitative microbial analysis of the Anammox process assessed the variation in functional Anammox microorganisms in the SBBR on day 0, 33, 75, and 107. During the initial phase of ASBBR, the average Anammox gene ratio of the inoculum was 7.55% (Table 4). When the carriers were added, the average Anammox gene ratios in biofilm and floccules reached 1.34% and 5.22% (Table 4), respectively. This suggested the Anammox adhered to the carriers, gradually forming the biofilm. Table 4 During Phase 2, the SBRni effluent was added to the ASBBR, and the average Anammox gene ratios in biofilm and floccule increased to 5.49% and 8.48%, respectively. This suggested that the low levels of nonbiodegradable substances did not significantly inhibit Anammox. Compared with the floccules, the growth of 18
Anammox gene ratio in the biofilm was larger. This indicated that the biofilm could effectively mitigate adverse effects from the external environment (e.g. landfill leachate). Finally during Phase 3, the Anammox gene ratios in biofilm and floccule changed when the raw landfill leachate was added. The Anammox ratio of biofilm significantly increased to 13.28%, whereas the ratio of floccule decreased to 3.88%. This was the result of heterotrophic bacteria growth. Nevertheless, with the help of the biofilm, Anammox adhered more effectively to the carriers. The results suggest that due to the biofilm of Anammox, the Anammox and heterotrophic bacteria could coexist in the system, leading to stable nitrogen removal performance even when organics were present.
4 Conclusions A two-stage Anammox SBBR system was used to treat leachate for 107 days. With influent ammonia and COD concentrations of 3000±100 mg/L and 3000±100 mg/L, respectively, effluent TN remained below 20 mg/L, with 95% of TN removed. EPS results during the stable phase showed that S-EPS and LB-EPS of floccules were both higher than biofilm, while TB-EPS of biofilm was significantly higher. The quantitative microbial analysis showed that as influent COD increased, the gene ratios of floccules first increased, and then decreased to 3.88%. The Anammox gene ratios of biofilm increased from 1.34% to 13.28% due to the biofilm protection.
19
Acknowledgements Financial support of Natural Science Foundation of China (51478013) and Funding Projects of Beijing Municipal Commission of Education are gratefully acknowledged.
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[7]Hu, L., Zeng, G.M., Chen, G.Q., Dong, H.R., Liu, Y.T., Wan, J., Chen, A.W., Guo, Z., Yan, M., Wu, H.P., Yu, Z.G., 2016. Treatment of landfill leachate using immobilized Phanerochaete chrysosporium loaded with nitrogen-doped TiO2 nanoparticles. J. Hazard. Mater. 301, 106-118 [8] Jin, R.C., Ma, C., Yu, J.J., 2013. Performance of an Anammox UASB reactor at high load and low ambient temperature, Chem. Eng. J. 232, 17-25. [9] Jin, R., Zhang, Q.Q., Zhang, Z.Z., Liu, J.H., Yang, B.E., Guo, L.X., Wang, H.Z., 2014. Bio-augmentation for mitigating the impact of transient oxytetracycline shock on anaerobic ammonium oxidation (ANAMMOX) performance. Bioresour. Technol. 163, 244-253. [10] Kartal, B., Kuenen, J.G., van Loosdrecht, M.C.M., 2010. Sewage treatment with anammox. Science 328, 702–703. [11] Koike, S., Krapac, I.G., Oliver, H.D., Yannarell, A.C., Chee-Sanford, J.C., Aminov, R.I., Mackie, R.I. R-mackie Uiuc, 2007. Monitoring and source tracking of tetracycline resistance genes in lagoons and groundwater adjacent to swine production facilities over a 3-year period. Appl. Environ. Microbiol. 73, 4813-4823. [12] Kulikowska, D. and Klimiuk, E., 2008. The effect of landfill age on municipal leachate composition. Bioresour. Technol. 99, 5981-5985. [13] Li, X.L., Peng, Y.Z., Ren, N.Q., Li, B.K., Chai, T.Z., Zhang, L., 2014. Effect of temperature on short chain fatty acids (SCFAs) accumulation and microbiological transformation in sludge alkaline fermentation with Ca(OH)2 adjustment. Water Res. 61, 34-45. 21
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[22] Peng, Y., Zhang, S.J., Zeng, W., Zheng, S.W., Mino, T., Satoh, H., 2008. Organic removal by denitritation and methanogenesis and nitrogen removal by nitritation from landfill leachate. Water Res. 42, 883-892. [23] Renou, S., Givaudan, J.G., Poulain, S., Dirassouyan, F., Moulin, P., 2008. Landfill leachate treatment: Review and opportunity. J. Hazard. Mater. 150, 468-493. [24] Schmid, M.C., Maas, B., Dapena, A., van de Pas-Schoonen, K., van de Vossenberg, J., Kartal, B., van Niftrik, L., Schmidt, I., Cirpus, I., Kuenen, J.G., Wagner, M., Damsté, J.S.S., Kuypers, M., Revsbech, N.P., Mendez, R., Jetten, M.S.M., Strous, M., 2005. Biomarkers for in situ detection of anaerobic ammonium-oxidizing (anammox) bacteria. Appl. Environ. Microbiol. 71, 1677-1684. [25] Shi, Y., Hu, S.H., Lou, J.Q., Lu, P.L., Keller, J., Yuan, Z.G., 2013. Nitrogen Removal from Wastewater by Coupling Anammox and Methane-Dependent Denitrification in a Membrane Biofilm Reactor. Environ. Sci. Technol. 47, 11577-11583 [26] Strous, M., Kuenen, J.G.and Jetten, M., 1999. Key physiology of anaerobic ammonium oxidation. Appl. Environ. Microbiol. 65, 3248-3250. [27] Tang, C.J, Zheng, P., Wang, C.H., Mahmood, Q., 2010. Suppression of anaerobic ammonium oxidizers under high organic content in high-rate Anammox UASB reactor. Bioresour. Technol. 101, 1762-1768. [28] Tsushima, I., Ogasawara, Y., Kindaichi, T., Satoh, H., Okabe, S., 2007. Development of high-rate anaerobic ammonium-oxidizing (anammox) biofilm reactors, Water Res. 41, 1623–1634. 23
[29] Van der Star, W.R.L., Abma, W.R., Blommers, D., Mulder, J.W., Tokutomi, T., Strous, M., Picioreanu, C., Van Loosdrecht, M.C.M., 2007. Startup of reactors for anoxic ammonium oxidation: experiences from the first full-scale anammox reactor in Rotterdam, Water Res. 41, 4149–4163. [30] Van Hulle, S.W.H., Vandeweyer, H.J.P., Meesschaert, B.D., Vanrolleghem, P.A., Dejans, P., Dumoulin, A., 2010. Engineering aspects and practical application of autotrophic nitrogen removal from nitrogen rich streams. Chem. Eng. J. 162, 1–20. [31] Zhang, P., Fang, F., Chen, Y.P., Shen, Y., Zhang, W., Yang, J.X., Li, C., Guo, J.S., Liu, S.Y., Huang, Y., Li, S., Gao, X., Yan, P., 2014. Composition of EPS fractions from suspended sludge and biofilm and their roles in microbial cell aggregation. Chemosphere 117, 59-65. [32] Zhang, C., Lai, C., Zeng, G.M., Huang, D.L., Yang, C.P., Wang, Y., Zhou, Y.Y., Cheng, M., 2016. Efficacy of carbonaceous nanocomposites for sorbing ionizable antibiotic sulfamethazine from aqueous solution. Water Res. 95, 103-112. [33] Zhu, G.B., Wang S.Y., Wang Y., Wang C.X., Risgaard-Petersen, N., Jetten, M.S.M, Yin, C.Q., 2011. Anaerobic ammonia oxidation in a fertilized paddy soil. ISME J. 5, 1905-1912. [34] Zhu, R.L., Wang S.Y., Li, J., Wang, k., Miao, L., Ma, B., Peng, Y.Z., 2013. Biological nitrogen removal from landfill leachate using anaerobic-aerobic process: Denitritation via organics in raw leachate and intracellular storage polymers of microorganisms. Bioresour. Technol. 128, 401-408.
24
Figure 1 Schematic diagram of two-stage SBBR process treating mature landfill leachate Figure 2 ASBBR Performance during Phase 1 Figure 3 ASBBR Performance during Phase 2 Figure 4 ASBBR Performance during Phase 3 Figure 5 Variations of nitrogen and COD in a typical ASBBR cycle during Phase 4 Figure 6 EEM fluorescence spectra of EPS extracted from floccule and biofilm in ASBBR（ （ a: SEPS-floccule, b: SEPS-biofilm, c: LBEPS- floccule, d: LBEPS-biofilm, e: TBEPS- floccule, f: TBEPS-biofilm） ） Table 1 Experimental procedure Table 2 Main characteristics of the leachate Table 3 EPS concentrations and PN/PC ratios of floccules and biofilm in the ASBBR during the Phase 4 Table 4 Abundance of Anammox bacteria in ASBBR during different phases
25
Figure 1 Schematic diagram of two-stage SBBR process treating mature landfill leachate
26
Phase 1
3.0
3000
100 90
-
+
TINeff
NO3 -N/NH4 -N
2500 80 2000
TIN removal efficiency
1500
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70 60 50 40
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500
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20 10
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Figure 2 ASBBR Performance during Phase 1
27
0
TN removal efficiency (%)
+
NO2 -N/NH4 -N
TIN (mg/L)
2.0
-
TINinf
-
+
NO2 -N/NH4 -N, NO3 -N/NH4 -N(-)
2.5
Phase 2 +
-
+
NO2 -N/NH4 -N
TINeff
NO3 -N/NH4 -N
TIN removal
1800 110 1500 100 1200 90
1.5
900
80
1.0
600
70
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50
TIN (mg/L)
2.0
26 24
40 CODinf
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350 Reaction Time
300
CODeff
250
22
200
20
150
18 16
100
14
50
12
40
50 Time (day)
60
70
Figure 3 ASBBR Performance during Phase 2
28
0
COD (mg/L)
+ -
0.0 28
Reaction Time (h)
TN removal (%)
2.5
-
TINinf
-
+
NO2 -N/NH4 -N, NO3 -N/NH4 -N (-)
3.0
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NO2 -N/NH4 -N
TNeff
NO3 -N/NH4 -N TN removal
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90 800 600
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80 70 60
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Reaction Time (hour)
-
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CODeff Reaction Time
24
1000
20
800
16
600 400
12 80
85
90 Time (day)
95
100
Figure 4 ASBBR Performance during Phase 3
29
105
200
TN removal (%)
+
TNinf
-
2.0
-
COD (mg/L)
+
100
2.5
TN (mg/L)
NO2 -N/NH4 -N, NO3 -N/NH4 -N (-)
Phase 3
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NH4 -N
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80
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600 60
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300 20
200 100 0
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COD (mg/L)
COD
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100 TN removal
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Figure 5 Variations of nitrogen and COD in a typical ASBBR cycle during Phase 4
30
Figure 6 EEM fluorescence spectra of EPS extracted from floccule and biofilm in ASBBR （a: SEPS-floccule, b: SEPS-biofilm, c: LBEPS- floccule, d: LBEPS-biofilm, e: TBEPS- floccule, f: TBEPS-biofilm） ）
31
Experimental Phases Phase 1 Phase 2
Phase 3
Phase 4
Table 1 Experimental procedure Substances Concentrations of Influent Composition ASBBR influence (mg/L) of ASBBR NH4 + NO2COD synthetic wastewater The mixture of synthetic wastewater and effluent of SBRni The mixture of raw mature landfill leachate and effluent of SBRni The mixture of raw mature landfill leachate and effluent of SBRni
32
500-1000
600-1400
——
300-600
400-800
100-200
200-350
250-450
500-1000
200-250
250-300
800-1000
Table 2 Main characteristics of the leachate Compounds Concentration COD(mg/L) 3000±1000 BOD5(mg/L) 100±50 + NH4 -N(mg/L) 3000±100 NO3--N(mg/L) 1±0.5 NO2 -N(mg/L) 1±0.5 Alkalinity(mg/L) 7000±100 pH 8.0±0.2
33
Table 3 EPS concentrations and PN/PC ratios of floccules and biofilm in the ASBBR during the Phase 4
S-EPS
LB-EPS
TB-EPS
PN/PC
Sample Protein(mg/gVSS) Carbohydrate(mg/gVSS) Nucleic acid(mg/gVSS) Protein(mg/gVSS) Carbohydrate(mg/gVSS) Nucleic acid(mg/gVSS) Protein(mg/gVSS) Carbohydrate(mg/gVSS) Nucleic acid(mg/gVSS) S-EPS LB-EPS TB-EPS
Floccule-ASBBR 42.15 118.55 20.10 8.04 109.26 2.31 34.78 79.00 3.93 0.36 0.07 0.44
34
Biofilm-ASBBR 14.54 132.08 3.09 6.38 102.45 0.52 98.32 109.86 17.60 0.11 0.06 0.89
Table 4 Abundance of Anammox bacteria in ASBBR during different phases
Anammox gene All bacteria gene Samples numbers（gene numbers（gene -1 copies•g dry sludge） copies•g-1 dry sludge） 9.45×10 7 1.25×10 9 Inoculum Biofilm 2.45×10 8 1.82×1010 Day 33rd 8 Floccule 7.59×10 1.45×1010 Biofilm 9.61×10 8 1.75×1010 Day 75th Floccule 1.09×10 9 1.28×1010 9 Biofilm 2.63×10 1.98×1010 th Day 107 Floccule 3.81×10 8 9.82×10 9
35
Average Anammox gene ratio 7.55% 1.34% 5.22% 5.49% 8.48% 13.28% 3.88%
36
Highlights Anammox was enriched in carriers to enhance ability against adverse environment. As influent TN was 3000 mg/L, effluent TN was below 20 mg/L with TN removal of 95%. The ratio of NO3−:NH4 + fell below 0.1 in Anammox process treating real leachate. TB-EPS of biofilm and floccule were 46% and 28%, respectively, helped form biofilm. Anammox gene ratio of biofilm increased to 13.28%, yet floccule decreased to 3.88%.
37
S0960-8524(16)30955-5 http://dx.doi.org/10.1016/j.biortech.2016.06.131 BITE 16753
To appear in:
Bioresource Technology
Received Date: Revised Date: Accepted Date:
11 April 2016 21 June 2016 24 June 2016
Please cite this article as: Miao, L., Wang, S., Cao, T., Peng, Y., Zhang, M., Liu, Z., Advanced nitrogen removal from landfill leachate via Anammox system based on Sequencing Biofilm Batch Reactor (SBBR): Effective protection of biofilm, Bioresource Technology (2016), doi: http://dx.doi.org/10.1016/j.biortech.2016.06.131
This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
Advanced nitrogen removal from landfill leachate via Anammox system based on Sequencing Biofilm Batch Reactor (SBBR): Effective protection of biofilm Lei Miaoa, Shuying Wanga,*, Tianhao Caoa, Yongzhen Penga,, Man Zhanga, Zhaoyuan Liua a,
*Engineering Research Center of Beijing, Beijing University of Technology, Beijing, PR China
Abstract High levels of organics negatively affect Anammox for treating landfill leachate. To enhance the ability of Anammox to survive against adverse environments, a lab-scale two-stage Anammox system using a Sequencing Biofilm Batch Reactor was applied to treat mature landfill leachate under 35 °C. Over 107 days, with influent total nitrogen (TN) and chemical oxygen demand (COD) concentrations of 3000±100 and 3000±100 mg/L, effluent TN was below 20 mg/L. For extracellular polymeric substance (EPS) of Anammox, slime-EPS and loosely-bound-EPS of floccules were both higher than biofilm, while tight-bound-EPS of biofilm was significantly higher, contributing to biofilm formation. Quantitative microbial analysis showed that as influent COD increased, Anammox gene ratios of biofilm increased from 1.34% to 13.28%; the gene ratios of floccule first increased, then decreased to 3.88%. This indicated that Anammox and heterotrophic bacteria could coexist because of the biofilm, leading to stable nitrogen removal performance, even when organics were present. Keywords: Anammox; landfill leachate; biofilm; EPS; PCR; *Corresponding author Tel/Fax +86 10 67392627
1
Email: [email protected] (S. Wang); [email protected] (L. Miao)
1 Introduction Landfill leachate is a kind of wastewater, containing large amounts of organics and inorganics, with a particularly high ammonia concentration (Renou et al., 2008). Landfill leachate treatments are usually classified as physical/chemical and biological treatments (Hu et al., 2016; Renou et al., 2008). The physical/chemical treatment (e.g. coagulation-flocculation) is usually used as the pre-treatment or advanced treatment step due to its high cost and secondary pollution (Zhang et al., 2016). Biological processes, based on active sludge, are usually used to remove nitrogen and organics from the leachate due to their simplicity and cost-effectiveness (Zhu et al., 2013). For early-age or middle-age leachate, a conventional nitrification/denitrification process or modified process (such as the process coupling of two-stage up-flow anaerobic sludge bed (UASB) and anoxic/oxic reactor (A/O), or the endogenous denitrification process), is used to remove nitrogen and organics because of the high carbon to nitrogen (C/N) ratio (Peng et al., 2008; Miao et al., 2015). However, it is difficult to achieve high nitrogen removal from mature leachate with a low C/N ratio (usually <3), because the conventional nitrification/denitrification process requires additional external carbon sources. This requirement makes the treatment systems complicated and expensive (Kulikowska et al., 2008). Anaerobic ammonium oxidation (Anammox), is a novel cost-effective process with great potential (Mulder et al., 1995; Kartal et al., 2010). Compared with the conventional nitrification/denitrification process, Anammox does not need external 2
carbon sources, and the power consumption for Anammox can be reduced substantially (Shi, et al., 2013). Therefore, it can effectively treat mature landfill leachate (Anfruns, et al., 2013). However, Anammox cannot work alone, because it requires nitrite substrate supplied by nitritation. As such, Anammox is usually combined with nitritation to treat wastewater. Nevertheless, some factors adversely affect Anammox, including high nitrite and organic concentrations. For example, Strous et al. (1999) demonstrated that Anammox was inhibited by nitrite concentrations up to 100 mg/L. Thus, it is important to avoid feeding in a short time for Anammox process treating the wastewater with high substrate content. Further, organics can significantly influence Anammox (Tang et al., 2010). Molinuevo et al. (2009) observed that a COD concentration up to 292 mg/L could completely inhibit Anammox. Ni et al. (2012) studied artificial wastewater, finding that a low COD concentration did not affect nitrogen removal; however, a high COD concentration did suppress Anammox activity and reduced the bacteria population when the COD exceeded 400 mg/L. Miao et al. (2014) used a three-stage Anammox system to treat landfill leachate, finding that Anammox was not significantly inhibited when the biodegradable COD was less than 150 mg/L. Therefore, applying Anammox requires controlling the influent COD concentration. Furthermore, for Anammox to be effective in treating real wastewater, the system must make full use of the organics in the wastewater and enhance nitrogen removal performance through the synergy between Anammox and denitrification. Because Anammox bacteria grow slowly, it is important to enrich the Anammox 3
bacteria during the start-up of the Anammox system. Compared with the activated sludge process, the biofilm technology associated with Anammox covers a smaller footprint, and has more microbial populations (Van Hulle et al., 2010), a stronger impact resistance (Nicolella et al., 2000; Han et al., 2012) (such as oxygen and high nitrite concentration), and a more stable ecological system when treating wastewater. In addition, biofilm systems are more efficient and more widely applicable in nitrogen removal systems than activated sludge (Tsushima et al., 2007; van der Star et al., 2007). Therefore, in this study a two-stage Sequencing Biofilm Batch Reactor (SBBR) system coupled with biofilm was applied to treat mature landfill leachate and achieve a fast start-up. The extracellular polymeric substances (EPS) and microbial population of the Anammox SBR during different phases were measured to assess microbiological changes. 2 Materials and Methods 2.1 Experimental setup and operational procedure The two-stage SBBR system used in this study, was comprised of a nitritation SBR (SBRni) and Anammox SBBR (ASBBR). The system was built using polymethyl methacrylate with a total working volume capacity of 20 L, divided into equal parts of 10 L each for the two system components (SBRni and ASBBR). The SBRni was equipped with pH meters, mechanical stirrers, and air diffusers; the ASBBR was equipped with a pH meter and a mechanical stirrer. Aeration intensity was maintained at 100 L/h during the aeration period. The operational temperatures for the SBRni and ASBBR were maintained at 25 and 35 °C, respectively, using a 4
temperature controller. Figure 1 Table 1 presents the experimental procedure. There were four phases during the start-up and steady operation of the two-stage SBBR system. During Phase 1, synthetic wastewater was used. Carriers were added to the ASBBR to form the biofilm and promote biofilm culturing. The carriers were cubic sponges (1 cm×1 cm×1 cm) made of polyurethane, and the packing ratio was 10%. During Phase 2 and 3, the SBRni effluent and real mature landfill leachate, respectively, were gradually added into the ASBBR. During Phase 4, the ASBBR was in steady operation; the ASBBR influent was the mixed effluent of SBRni and raw mature landfill leachate with a ratio of NO2−-N/NH4+-N of 1.3-1.5 (Figure 1). Table 1 In the two-stage SBBR system, the SBRni was operated using the traditional approach: filling-aeration-settle-decant. ASBBR was operated using the modified approach: continuous feeding (5 h) and stirring until the reaction was completed. The terminal point of Anammox was determined by the pH profile. The operational cycle time of each stage was controlled at 24 hours. The exchange volumetric rates of SBRni and ASBBR were both 50%. 2.2 Influent and seed sludge During the biofilm formation phase (Phase 1), the ASBBR system was fed with synthetic wastewater, containing KH2PO4 (10 mg/L), CaCl2·2H2O (5.6 mg/L), MgSO4·7H2O (300 mg/L), KHCO3 (1250 mg/L), trace element solution I (EDTA 5
5000 mg/L and FeSO4 5000 mg/L), and trace element solution II (EDTA 1000 mg/L, H3BO4 14 mg/L, MnCl2·4H2O 990 mg/L, CuSO4·5H2O 250 mg/L, ZnSO4·7H2O 430 mg/L, NiCl2·6H2O 190 mg/L, NaSeO4·10H2O 210 mg/L, NaMoO4·2H2O 220 mg/L) (Miao et al., 2014). NaNO2 and NH4Cl solutions were added to supply nitrite and ammonium for Anammox activities. Influent pH was controlled at 7.5 ± 0.2. After start-up, mature landfill leachate collected from the Liulitun Municipal Solid Waste (MSW) Sanitation Landfill Site (Beijing, China) was used as the feeding solution (Table 2). Table 2 The mixed liquor suspended solid (MLSS) of the SBRni was 3000±100 mg/L. The excess sludge collected from the existing lab-scale Anammox SBR, treating synthetic wastewater, was used as the inoculums for the ASBBR. The inoculums of ASBBR were the mixture of granule and floccules; the MLSS was 2000±100 mg/L after inoculation. 2.3 Analytical methods The pH and temperature were monitored using a pH/Oxi 340i analyzer (WTW Company, Germany). MLSS, NH4+-N, NO3−-N, NO2--N, and COD were measured using the standard methods (APHA, 1995). 2.4 EPS extraction and analysis EPS is typically categorized as slime EPS (S-EPS), loosely-bound EPS (LB-EPS), and tight-bound EPS (TB-EPS). The sludge samples of biofilm and floccule in the Anammox SBBR were collected during the steady state phase. The sludge was 6
directly centrifuged at 2000 g for 15 min, and the supernatant was collected as the S-EPS. The residual was re-suspended to its original volume using a buffer solution with a pH of 7, containing 1.3 mM Na3PO4, 2.7 mM NaH2PO4, 6 mM NaCl, and 0.7 mM KCl. The suspension was centrifuged again at 5000 g for 15 min, and the supernatant was collected as the LB-EPS. Then, the residual was re-suspended again with a buffer solution to the original volume; the solution was then treated using ultrasound at 20 kHz and 480 W for 2 min. The extracted solution was centrifuged at 20,000 g for 20 min, and the supernatant was collected as the TB-EPS. Protein (PN), polysaccharose (PC), and nucleic acid were measured using Lowry-Folin methods, Phenol-Sulfuric acid method, and diphenylamine methods, respectively (Li et al., 2014). The three-dimensional excitation-emission matrix (3D-EEM) spectra of all EPS samples were measured using an EEM fluorescence spectrophotometer (LS-55, Perkin Elmer, USA). In this study, the 3D-EEM spectra were collected with corresponding scanning excitation spectra from 220 to 550 nm at 1 nm increments by varying the emission wavelength from 200 to 550 nm at 10 nm increments. 2.5 Calculations Equations (1) and (2) were used to calculate the nitrogen load rate (NLR) (kgN/m3·d) and nitrogen removal rate (NRR) (kgN/m3·d), respectively. mgN
NRR gVSS·h =
mg
-
mg
-
mg
mg
-
mg
-
mg
NH4+ –Ninf L +NO2 –Ninf L +NO3 –Ninf L -NH4+ –Neff L -NO2 –Neff L -NO3 –Neff L (1) g th ∙MLVSSL
7
#$ !" & ( % ·d =
)*+ – ,-. %$/ + 012 – ,-. %$/ + 0&2 – ,-. %$/ × 4567 × 102: 1 4%& ∙ ;
<
2
where NH4+-Ninf, NO2--Ninf, NO3--Ninf and NH4+-Neff, NO2--Neff, NO3--Neff are the nitrogen concentrations of ASBBR influent and effluent (mg/L), respectively; Vinf is the volume of ASBBR influent (L); V is the working volume of ASBBR (L); MLVSS is the mixed liquor volatile suspended solid concentration of ASBBR (gVSS); and t is the reaction time of ASBBR (h). 2.6 DNA isolation and PCR Anammox bacterial populations were quantified and correlated with nitrogen removal efficiency in the Anammox system. The biofilm and floccule activated sludge samples from the ASBBR were collected during Phases 1, 2, 3, and 4. The DNA was extracted from the sludge sample using a FastDNA SPIN Kit for soil (Bio 101, Vista, CA). A nested PCR assay was conducted to amplify the Anammox 16S rRNA gene. The all-bacterial amplification was carried out using a 341f-543r (Koike et al., 2007) with a thermal profile of 95 °C for 10 min, followed by 25 cycles of 30 s at 95 °C, 30 s at 55 °C, and 30 s at 72 °C. The Anammox amplification was carried out using an Amx368f-Amx820r primers31 (Schmid et al., 2005) with a thermal profile of 96 °C for 10 min followed by 25 cycles of 1 min at 96 °C, 1 min at 52 °C, and 1 min at 72 °C. The PCR reagents (50 µL) contained 10× PCR buffer (5 µL), dNTPs (4 µL, 2.5 mmol/L), Ex Taq polymerase (0.5 µL, 2.5U, Takara, Dalian, China), each primer (1 µL, 10 mmol /L), DNA template (1.25 µL, 1–10 ng), and ddH2O (37.25 µL) (Zhu et al., 2011). 8
2.7 Quantitative PCR The abundances of all the bacterial and Anammox DNAs were determined by real-time PCR using an MX3000P Real-Time PCR system (Stratagene, La Jolla, CA) equipped with the fluorescent dye SYBR-Green approach. The primers for all the bacteria and Anammox in the real-time PCR were 341f-543r and Amx368f-Amx820r. The amplification was performed in the 20 µL reaction mixtures, containing SYBR Green exTaq (10 µL, Takara, Dalian, China), ROX Reference Dye 50 (0.3 µL), each primer (0.3 µL, 10 mmol/L), and the DNA template (2 µL, 1–10 ng). The program contains the following steps: 3 min at 95 °C, followed by 40 cycles of 30 s at 95 °C, 30 s at 59 °C, and 30 s at 72 °C.
3 Results and Discussion 3.1 Performance of the two-stage SBBR system In the two-stage SBBR system, nitrite accumulation remained above 95% in SBRni, which treated the mature landfill leachate. The start-up of ASBBR lasted for 107 days (Phase 1, 2, and 3). After Phase 1, the SBRni was combined with the ASBBR to test the two stage process. 3.1.1 Start-up of the two-stage Anammox system Phase 1 lasted 33 days; carriers were added into the ASBBR to form the biofilm and promote biofilm culturing. During the first 14 days of Phase 1, the influent total inorganic nitrogen (TIN) was maintained at 1500±200 mg/L (Figure 2). The influent nitrite concentration reached 850±50 mg/L, which could significantly inhibit 9
Anammox bacteria. However, the ASBBR feeding mode was then changed to a 5 h continuous feeding mode, reducing the level at which high nitrite concentrations inhibited Anammox bacteria. The NO2-N/NH4-N remained at 1.30±0.5 (Figure 2), similar to the theoretical value of 1.32. The NO3-N/NH4-N was 0.19±0.2 (Figure 2), lower than the theoretical value (0.26). This is because microorganisms secreted EPS; denitrifying bacteria then used this EPS as a carbon source to remove nitrate (Ni et al., 2012). Figure 2 The effluent TIN concentration remained at 130±20 mg/L and TIN removal exceeded 90% (Figure 2), indicating good performance of the ASBBR. Therefore, starting on day 15 th, the influent TIN concentration was increased to 1800±100 mg/L, and on day 26th, the concentration was increased further to 2200±200 mg/L (Figure 2). The TIN removal held above 90%, and the NO2-N/NH4-N and NO3-N/NH4-N remained at 1.40±0.5 and 0.20±0.2, respectively. These levels are similar to the levels seen during the first 14 days (Figure 2). The reaction time was then prolonged to 15 h and 16 h, respectively (not shown in Figure 2). Finally, the nitrogen loading rate (NLR) and nitrogen removal rate (NRR) reached 1.65 kgN/(m3·d) and 39.06 mgN/(gVSS·h), respectively. The sludge concentration in the ASBBR was also measured. The MLSS of the flocculent sludge and biofilm reached approximately 3900±100 mg/L and 570±50 mg/L, respectively. These results indicate the biofilm had been successfully cultured in the carriers. Figure 3 10
Starting on day 34th (Phase 2), SBRni effluent was gradually added into the ASBBR, thus providing influent nitrite to the ASBBR. To help the Anammox bacteria adapt to the landfill leachate, only 10% of the ASBBR influent nitrite initially came from the SBRni effluent. This proportion was then gradually increased to 20%, 50% and 100% (Figure 3). Because the SBRni effluent nitrite concentration was limited (<800 mg/L), the influent TIN concentration of ASBBR decreased as the ratio of SBRni effluent increased. At the end of Phase 2, the influent TIN concentration of the ASBBR was maintained at 800 mg/L, with an effluent TIN concentration of 100 mg/L. There were almost no biodegradable organics in the SBRni effluent after aeration (influent COD of ASBBR was only 200 mg/L). Therefore, no significant denitrification was observed and Anammox was not inhibited in Phase 2. Compared with Phase 1, the NO2-N/NH4-N and NO3-N/NH4-N remained stable at 1.35±0.5 and 0.18±0.2, respectively. However, the reaction time increased from 14 h to 20 h, because the Anammox bacteria were affected by the landfill leachate. At the end of Phase 2, the reaction time returned to 14 h, suggesting the Anammox bacteria had adapted to the SBRni effluent. During Phase 2, the NLR and NRR decreased to 0.69 kgN/(m3·d) and 14.58 mgN/(gVSS·h), respectively, along with the decrease in influent TIN concentration. Furthermore, during Phase 2, the effluent COD of ASBBR was a little higher than the influent COD concentration. This may be because as the landfill leachate was added, some Anammox bacteria broke apart and some microorganisms secreted EPS. Figure 4 11
After the ASBBR adapted to the incoming SBRni effluent, raw mature landfill leachate was gradually added to the ASBBR (Phase 3, from day 76th). To reduce the impact of raw landfill leachate on the Anammox bacteria, only 50% of the ammonia in the ASBBR influent was supplied by the raw mature landfill leachate. When the raw landfill leachate was added, the influent organics also increased. This led to a gradual decrease in the effluent TIN of ASBBR. Until day 93 rd, the effluent TIN of ASBBR decreased to 40 mg/L (Figure 4). The NO2-N/NH4-N remained unchanged (1.35±0.5); in contrast, the NO3-N/NH4-N fell significantly, to approximately 0.1. This is because while the biodegradable organic concentration in the raw landfill leachate was low (<300 mg/L), the denitrifying bacteria also used the organics to reduce nitrate to N2. From day 94th, 100% of the ammonia in the ASBBR influent was supplied by the raw mature landfill leachate. With 100% of landfill leachate added, the effluent TN dropped further, reaching 20 mg/L (Figure 4). Compared with Phase 2, the TN removal efficiency improved significantly (reached above 95%). The NO3-N/NH4-N fell below 0.1, suggesting that the Anammox was not significantly inhibited by the influent organics. In contrast, nitrogen removal was improved by the appropriate denitrification. During Phase 3, the reaction time initially only increased 2 h, and then returned to 14 h (Figure 4). This suggests that the Anammox bacteria were enriched in the added carriers, enhancing the bacteria’s resistance to the adverse environment. Furthermore, the Anammox gene copies numbers and ratios (Table 4) also supported the conclusion above. The biofilm carriers helped Anammox bacteria coexist with the 12
denitrifiers, improving nitrogen removal through the combination of Anammox and denitrification. Finally, the MLSS of the floccules and biofilm in ASBBR reached approximately 4100±100 mg/L and 630±50 mg/L, respectively (The pictures of floccule sludge and biofilm were shown in Figure S1). The NLR and NRR were maintained at 0.51 kgN/(m3·d) and 12.08 mgN/(gVSS·h), respectively, with the TN removal of above 95%. Anfruns et al. (2013) used partial nitritation coupled with Anammox SBR process to treat landfill leachate, an NLR of 0.32 kgN/(m3·d) with only nitrogen removal of 84% was achieved. Miao et al. (2014) applied three-stage Anammox SBR system to treat landfill leachate; only about 90% of TN was removed. Compared with other studies which applied Anammox to treat real wastewater, the two-stage Anammox SBBR system achieved a better performance of TN removal. Furthermore, the difference in the value between influent and effluent COD of ASBBR in this study was maintained at 200 mg/L (Figure 4). Molinuevo et al. (2009) found
that
when
COD
concentration
of
pig
manure
effluent
after
UASB-post-digestion reached above 237 mg/L, Anammox was inhibited and the denitrification became the dominant nitrogen removal process, which was similar with the conclusion in this study. When comparing with the Anammox process coupled with granular sludge (the Anammox system collapsed when influent biodegradable COD exceeded 150 mg/L (Miao et al., 2014)), the Anammox system with biofilm tolerated higher COD concentrations. This could indicate that the added carriers enriched the Anammox bacteria and supported coexistence between Anammox bacteria and denitrifiers. The conclusion could also be demonstrated using 13
the quantitative microbial analysis of the Anammox process (discussed in Part 3.3, Table 4)). Compared with the Anammox gene numbers of the inoculum (9.45×10 7 gene copies/(g·dry·sludge), Table 4), the Anammox gene numbers in the biofilm finally reached 2.63×109 gene copies/(g·dry· sludge) (Table 4), indicating the Anammox bacteria have been enriched. Therefore, to remove nitrogen from landfill leachate, adding the biofilm carriers into the Anammox system helped stabilize the system and improved nitrogen removal. 3.1.2 Anammox SBBR performance in a typical cycle during the steady phase After the system was stable, ASBBR performance in a typical cycle during steady phase (Phase 4) was monitored (Figure 5). The ASBBR was set in a continuous feeding mode for 5 hours, reducing the degree to which the influent’s high nitrite concentration inhibited the Anammox bacteria (the inhibition level was about 100 mg/L, Strous et al. 1999). Therefore, the feeding was accompanied by an Anammox reaction, and the average hourly TN removal in remained above 100 mg (Figure 5). The pH during the feeding time decreased slightly (from 8.11 to 7.89), also supporting high activity by the Anammox bacteria. Additionally, the nitrate in the ASBBR only increased slightly as the COD decreased during the first 5 hours. This indicates that denitrification was involved in the nitrogen removal process, as the influent’s biodegradable organics increased gradually with feeding. After feeding, the maximum ammonia and nitrite concentrations of ASBBR reached approximately 79 mg/L and 92 mg/L, respectively, suggesting the Anammox bacteria were not significantly inhibited. Furthermore, the 14
maximum ammonia and nitrite concentrations also led to a high driving force of reaction. Therefore, the TN removal in the first hour after feeding reached its maximum (about 386 mg) (Figure 5). During the following 6 hours, the hourly TN removal rate changed little, remaining at approximately 200 mg/L until the 12 th h. Between the 5th hour and 12th h, the nitrate concentration also varied little, suggesting the Anammox was accompanied with denitrification. Until the 12th h, the nitrite concentration decreased to 2 mg/L, with a remaining ammonia level of 8 mg/L (Figure 5). Figure 5 Over the following 2 hours, the nitrite concentration remained stable, while the ammonia and nitrate concentrations decreased. This is because the nitrate was reduced to nitrite, and was removed with the remaining ammonia by the Anammox. Because only a little ammonia and nitrate remained during this period, the hourly TN removal decreased to 100 mg. Until the 14th h, the ammonia concentration was close to 0 mg/L; meanwhile, a change point occurred in the pH profile. During the reaction time after feeding (Figure 5), the pH continued to rise until the 14th h, when it began to decline. This signaled the completion of Anammox. Therefore, the pH value can be used to identify the terminal point of Anammox. The difference of 200 mg/L between influent and effluent COD enhanced the denitrification and led to a final decrease in effluent nitrate concentration to 12 mg/L. 3.2 Analysis of EPS from Anammox SBBR EPS is one kind of macromolecule organics excreted by microorganisms; they 15
include carbohydrates, proteins, nucleic acids, and humic acids. The presence of EPS supports the survival of many microorganisms under different circumstances (Liu et al., 2004). For example, it could enrich the nutriment in the external environment for microorganisms and promote the formation and stability of microbial community structure (Flemming et al., 2010). In this study, EPS from sludge floccules and the biofilm in the ASBBR during Phase 4 were extracted and analyzed (Table 3). Compared with floccules in the ASBBR, the total EPS content of biofilm was higher. For S-EPS and LB-EPS, the EPS of the floccules exceeded that of the biofilm; however, TB-EPS of biofilm was much higher than the TB-EPS of floccules (225.77 mg/gVSS and 117.71 mg/gVSS, respectively). This contributed to biofilm formation. An analysis of EPS composition showed that carbohydrates made up a significant proportion in both floccules and biofilm (73.38% and 71.03%, respectively), followed by protein and nucleic acid. This finding was consistent with Jin et al. (2014). The large amount of carbohydrate found in S-EPS, LB-EPS, and TB-EPS facilitated cell adhesion on the biofilm surface, supporting its formation (Zhang et al., 2014). Protein was mainly found in S-EPS and TB-EPS, particularly in the S-EPS of floccule and TB-EPS of biofilm. This was also similar to the results of Zhang et al., (2014). These results suggest that extracellular proteins varied between the floccule and biofilm. The nucleic acid content of EPS was the lowest, below 10%, in both floccule and biofilm. Table 3 Researchers have proposed that the ratio of protein and carbohydrate (PN/PC 16
ratio) in EPS may be a good indicator of sludge stability: the higher the PN/PC ratio, the poorer the stability (Jin et al., 2013). The PN/PC ratio of the floccule was higher than the biofilm for S-EPS, nevertheless, the opposite result was seen for TB-EPS. Overall, compared with an Anammox system with synthetic wastewater (Jin et al., 2014), the PN/PC ratio of EPS in ASBBR was lower, for the floccule and the biofilm. This suggests that the ASBBR had good sludge stability, possibly due to the acclimation of Anammox system to real landfill leachate. Figure 6 Figure 6 shows the 3D-EEM fluorescence spectra of each EPS fractions from floccule and biofilm in the ASBBR. Two peaks were clearly observed from the 3D-EEM fluorescence spectra of the five samples (SEPS-floccule, LBEPS- floccule, LBEPS-biofilm, TBEPS- floccule, and TBEPS-biofilm), except the SEPS-biofilm. However, the positions of the peaks varied across all samples, which could indicate different substances in EPS (Chen et al., 2003). Peak A was observed at the excitation/emission wavelengths (Ex/Em) of 340-345 nm/430-435nm, indicating a humic acid-like substance. Peak B was identified at the Ex/Em of 230-240 nm/455-460 nm, which was considered to be a fulvic acid-like substance (Figure 6 (a) and (b)). The humic acid-like substance and fulvic acid-like substance detected in SEPS may have come from the raw mature leachate. Peak C was identified at the Ex/Em of 220-230 nm/355-360 nm, also considered to be a fulvic acid-like substance (Figure 6 (c) and (d)). It can be concluded that there were no large differences between the substances 17
in S-EPS and LB-EPS of the floccule and biofilm. In contrast, the positions of the peaks changed significantly in TB-EPS of the floccule and biofilm. Peak A, B and C were not observed, while Peak D and E were identified at the Ex/Em of 270-280 nm/345-350 nm and the Ex/Em of 220-230 nm/345-350 nm, respectively (Figure 6 (e) and (f)). Peak D was related to soluble microbial byproduct-like substances, while Peak E was related to aromatic proteins-like substances. It suggested that TB-EPS, which was closest to the cell, contains many protein-like substances and byproduct secreted by microorganisms. As such, it plays an important role in protecting the microorganisms.
3.3 Quantitative microbial analysis of Anammox A quantitative microbial analysis of the Anammox process assessed the variation in functional Anammox microorganisms in the SBBR on day 0, 33, 75, and 107. During the initial phase of ASBBR, the average Anammox gene ratio of the inoculum was 7.55% (Table 4). When the carriers were added, the average Anammox gene ratios in biofilm and floccules reached 1.34% and 5.22% (Table 4), respectively. This suggested the Anammox adhered to the carriers, gradually forming the biofilm. Table 4 During Phase 2, the SBRni effluent was added to the ASBBR, and the average Anammox gene ratios in biofilm and floccule increased to 5.49% and 8.48%, respectively. This suggested that the low levels of nonbiodegradable substances did not significantly inhibit Anammox. Compared with the floccules, the growth of 18
Anammox gene ratio in the biofilm was larger. This indicated that the biofilm could effectively mitigate adverse effects from the external environment (e.g. landfill leachate). Finally during Phase 3, the Anammox gene ratios in biofilm and floccule changed when the raw landfill leachate was added. The Anammox ratio of biofilm significantly increased to 13.28%, whereas the ratio of floccule decreased to 3.88%. This was the result of heterotrophic bacteria growth. Nevertheless, with the help of the biofilm, Anammox adhered more effectively to the carriers. The results suggest that due to the biofilm of Anammox, the Anammox and heterotrophic bacteria could coexist in the system, leading to stable nitrogen removal performance even when organics were present.
4 Conclusions A two-stage Anammox SBBR system was used to treat leachate for 107 days. With influent ammonia and COD concentrations of 3000±100 mg/L and 3000±100 mg/L, respectively, effluent TN remained below 20 mg/L, with 95% of TN removed. EPS results during the stable phase showed that S-EPS and LB-EPS of floccules were both higher than biofilm, while TB-EPS of biofilm was significantly higher. The quantitative microbial analysis showed that as influent COD increased, the gene ratios of floccules first increased, and then decreased to 3.88%. The Anammox gene ratios of biofilm increased from 1.34% to 13.28% due to the biofilm protection.
19
Acknowledgements Financial support of Natural Science Foundation of China (51478013) and Funding Projects of Beijing Municipal Commission of Education are gratefully acknowledged.
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Figure 1 Schematic diagram of two-stage SBBR process treating mature landfill leachate Figure 2 ASBBR Performance during Phase 1 Figure 3 ASBBR Performance during Phase 2 Figure 4 ASBBR Performance during Phase 3 Figure 5 Variations of nitrogen and COD in a typical ASBBR cycle during Phase 4 Figure 6 EEM fluorescence spectra of EPS extracted from floccule and biofilm in ASBBR（ （ a: SEPS-floccule, b: SEPS-biofilm, c: LBEPS- floccule, d: LBEPS-biofilm, e: TBEPS- floccule, f: TBEPS-biofilm） ） Table 1 Experimental procedure Table 2 Main characteristics of the leachate Table 3 EPS concentrations and PN/PC ratios of floccules and biofilm in the ASBBR during the Phase 4 Table 4 Abundance of Anammox bacteria in ASBBR during different phases
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Figure 1 Schematic diagram of two-stage SBBR process treating mature landfill leachate
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105
200
TN removal (%)
+
TNinf
-
2.0
-
COD (mg/L)
+
100
2.5
TN (mg/L)
NO2 -N/NH4 -N, NO3 -N/NH4 -N (-)
Phase 3
3.0
800 +
NH4 -N
900
80
850
TN removal (mg)
600 60
500
change point
400 40
300 20
200 100 0
8.8
8.6
8.4 800
750
8.2
pH (-)
pH
COD (mg/L)
COD
-
NO3 -N
Nitrogen concentration (mg/L)
-
NO2 -N
700
100 TN removal
8.0
700 7.8 650
7.6
0 0
1
2
3
4
5
6
7
8
Time (hour)
9
10
11
12
13
14
15
600
7.4
Figure 5 Variations of nitrogen and COD in a typical ASBBR cycle during Phase 4
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Figure 6 EEM fluorescence spectra of EPS extracted from floccule and biofilm in ASBBR （a: SEPS-floccule, b: SEPS-biofilm, c: LBEPS- floccule, d: LBEPS-biofilm, e: TBEPS- floccule, f: TBEPS-biofilm） ）
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Experimental Phases Phase 1 Phase 2
Phase 3
Phase 4
Table 1 Experimental procedure Substances Concentrations of Influent Composition ASBBR influence (mg/L) of ASBBR NH4 + NO2COD synthetic wastewater The mixture of synthetic wastewater and effluent of SBRni The mixture of raw mature landfill leachate and effluent of SBRni The mixture of raw mature landfill leachate and effluent of SBRni
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500-1000
600-1400
——
300-600
400-800
100-200
200-350
250-450
500-1000
200-250
250-300
800-1000
Table 2 Main characteristics of the leachate Compounds Concentration COD(mg/L) 3000±1000 BOD5(mg/L) 100±50 + NH4 -N(mg/L) 3000±100 NO3--N(mg/L) 1±0.5 NO2 -N(mg/L) 1±0.5 Alkalinity(mg/L) 7000±100 pH 8.0±0.2
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Table 3 EPS concentrations and PN/PC ratios of floccules and biofilm in the ASBBR during the Phase 4
S-EPS
LB-EPS
TB-EPS
PN/PC
Sample Protein(mg/gVSS) Carbohydrate(mg/gVSS) Nucleic acid(mg/gVSS) Protein(mg/gVSS) Carbohydrate(mg/gVSS) Nucleic acid(mg/gVSS) Protein(mg/gVSS) Carbohydrate(mg/gVSS) Nucleic acid(mg/gVSS) S-EPS LB-EPS TB-EPS
Floccule-ASBBR 42.15 118.55 20.10 8.04 109.26 2.31 34.78 79.00 3.93 0.36 0.07 0.44
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Biofilm-ASBBR 14.54 132.08 3.09 6.38 102.45 0.52 98.32 109.86 17.60 0.11 0.06 0.89
Table 4 Abundance of Anammox bacteria in ASBBR during different phases
Anammox gene All bacteria gene Samples numbers（gene numbers（gene -1 copies•g dry sludge） copies•g-1 dry sludge） 9.45×10 7 1.25×10 9 Inoculum Biofilm 2.45×10 8 1.82×1010 Day 33rd 8 Floccule 7.59×10 1.45×1010 Biofilm 9.61×10 8 1.75×1010 Day 75th Floccule 1.09×10 9 1.28×1010 9 Biofilm 2.63×10 1.98×1010 th Day 107 Floccule 3.81×10 8 9.82×10 9
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Average Anammox gene ratio 7.55% 1.34% 5.22% 5.49% 8.48% 13.28% 3.88%
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Highlights Anammox was enriched in carriers to enhance ability against adverse environment. As influent TN was 3000 mg/L, effluent TN was below 20 mg/L with TN removal of 95%. The ratio of NO3−:NH4 + fell below 0.1 in Anammox process treating real leachate. TB-EPS of biofilm and floccule were 46% and 28%, respectively, helped form biofilm. Anammox gene ratio of biofilm increased to 13.28%, yet floccule decreased to 3.88%.
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