Environ. Sci. Technol. 2010, 44, 8957–8963
Using Sludge Fermentation Liquid To Improve Wastewater Short-Cut Nitrification-Denitrification and Denitrifying Phosphorus Removal via Nitrite ZHOUYING JI AND YINGUANG CHEN* State Key Laboratory of Pollution Control and Resources Reuse, School of Environmental Science and Engineering, Tongji University, 1239 Siping Road, Shanghai 200092, China
Received July 27, 2010. Revised manuscript received October 19, 2010. Accepted October 21, 2010.
Wastewater biological nutrient removal (BNR) by short-cut nitrification-denitrification (SCND) and denitrifying phosphorus removal via nitrite (DPRN) has several advantages, such as organic carbon source saving. In this paper, a new method, i.e., by using waste activated sludge alkaline fermentation liquid as BNR carbon source, for simultaneously improving SCND and DPRN was reported. First, the performance of SCND and DPRN with sludge fermentation liquid as carbon source was compared with acetic acid, which was commonly used in literatures. Sludge fermentation liquid showed much higher nitrite accumulation during aerobic nitrification than acetic acid (81.8% versus 40.9%), and the former had significant anoxic denitrification and phosphorus uptake. The soluble phosphorus and total nitrogen removal efficiencies with sludge fermentation liquid were much higher than with acetic acid (97.6% against 73.4% and 98.7% versus 79.2%). Then the mechanisms for sludge fermentation liquid showed higher SCND and DPRN than acetic acid were investigated from the aspects of wastewater composition, microorganisms assayed by 16S rRNA gene clone library, and fluorescence in situ hybridization. More NO2--N accumulated by the use of sludge fermentation liquid was attributed to be more humic acids in the influent, which inhibited nitrite oxidizing bacteria (NOB) more serious than ammonia oxidizing bacteria (AOB), and more AOB but less NOB were observed in the BNR system. The reasons for sludge fermentation liquid BNR system exhibiting greater shortcut denitrifying phosphorus removal were that there were less glycogen accumulating organisms and more phosphorus accumulating organisms and anoxic denitrifying phosphorus removal bacteria with higher nitrite reductase activity.
Introduction Phosphorus and nitrogen removal from wastewater is an effective approach for prevention of eutrophication of water bodies. Biological nutrient removal (BNR) by short-cut nitrification-denitrification (SCND), viz. partial aerobic oxidation of ammonia to nitrite and then the nitrite is anoxically denitrified in the presence of an organic carbon source to N2, has attracted much interest for several advantages * Corresponding author phone: 86-21-65981263; fax: 86-2165986313; e-mail:
[email protected]. 10.1021/es102547n
2010 American Chemical Society
Published on Web 11/05/2010
compared with traditional BNR, such as save 40% organic carbon and 25% oxygen, shorten reaction time, and reduce sludge production (1). Recently, some researchers reported the occurrence of denitrifying polyphosphate accumulating organisms (DNPAOs) capable of utilizing nitrate (NO3-) or nitrite (NO2-) instead of oxygen as the electron acceptor for biological phosphorus removal (2, 3). Denitrifying phosphorus removal via nitrate has become one main research content of biological phosphorus removal for the advantages of organic carbon saving and excessive sludge production reduction with simultaneous denitrification and phosphorus removal. If SCND and denitrifying phosphorus removal via nitrite (DPRN) could be combined together, the removal of nitrogen and phosphorus from a wastewater containing insufficient carbon source would be further improved. However, due to rapid conversion of nitrite to nitrate by nitrite oxidizing bacteria (NOB), it is not easy to achieve a stable SCND in general BNR operation. In the literature the SCND can be achieved by the control of temperature, pH, dissolved oxygen (DO) concentration, etc. (1, 4). As to the denitrifying phosphorus removal via nitrite, in order to maintain a good performance of DPRN, some control strategies should be also implemented, such as the real-time control of pH, ORP, and DO (5). It seems that if we want to achieve both SCND and DPRN in a wastewater treatment plant (WWTP), it is necessary to use a real-time controller, which will increase the investment of equipments. This study reported a new method for achieving stable short-cut nitrification-denitrification and denitrifying phosphorus removal via nitrite (SCND-DPRN) by the use of waste activated sludge alkaline fermentation liquid as BNR carbon source. As in the literature most of the studies on short-cut nitrification-denitrification and denitrifying phosphorus removal used a sodium acetate or acetic acid synthetic wastewater (2, 3, 6, 7); a control reactor with acetic acid as the wastewater-carbon source was also run, and its performance was compared with the sludge fermentation liquid. The mechanisms for sludge fermentation liquid showing more super short-cut nitrification-denitrification and denitrifying phosphorus removal via nitrite were investigated by the methods of key enzyme analysis, 16S rRNA gene clone library, and fluorescence in situ hybridization (FISH).
Materials and Methods Characteristics of the Sludges and the Alkaline Fermentation Liquid of Sludge. Both the seed sludge (used in sequencing batch BNR reactors) and the waste activated sludge (WAS, for sludge alkaline fermentation liquid production) were withdrawn from the secondary sedimentation tank of a municipal WWTP in Shanghai, China. This plant was operated with BNR. The WAS was concentrated by settling at 4 °C for 24 h, and its main characteristics before fermentation are as follows (averages data plus standard deviation of three tests): pH 6.7 ( 0.2, mixed liquor suspended solids (MLSS) 14508 ( 721 mg/L, mixed liquor volatile suspended solids (MLVSS) 10655 ( 527 mg/L, total carbohydrate 1472 ( 73 mg COD/L, and total protein 3101 ( 152 mg COD/L. The fermentation of WAS at alkaline pH and the separation of alkaline fermentation liquid were conducted according to the method described previously (8, 9). After separation the main characteristics of sludge alkaline fermentation liquid are (mg/L): PO43--P (soluble orthophosphate) 8.3 ( 0.4, NH4+-N (ammonia nitrogen) 15.2 ( 0.6, BOD (biochemical oxygen demand) 4035 ( 175, SCOD (soluble chemical oxygen demand) 5705 ( 275, acetic acid VOL. 44, NO. 23, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 1. Comparison of the Removal Efficiencies of Phosphorus, Nitrogen, BOD, Carbohydrate, and Protein in Two SBRsa removal efficiency (%) b
A-SBR F-SBRc
PO43--P
NH4+-N
TN
BOD
carbohydrate
protein
73.4 ( 0.9 97.6 ( 1.3
99.6 ( 1.5 100
79.2 ( 1.1 98.7 ( 1.2
99.5 ( 0.5 95.2 ( 1.4
100 100
100 100
a The data are the averages and their standard deviations in triplicate tests. b The characteristic of synthetic wastewater of A-SBR (mg/L): PO43--P 15.0 ( 0.5, NH4+-N 35.0 ( 1.2, TN 36.0 ( 1.5, and BOD 265 ( 12. c The characteristic of synthetic wastewater of F-SBR (mg/L): PO43--P 15.0 ( 0.5, NH4+-N 35.0 ( 1.2, TN 38.2 ( 1.7, BOD 265 ( 12, soluble carbohydrate (as COD) 5.6 ( 0.3, and soluble protein (as COD) 61.2 ( 1.7.
(as COD) 1206 ( 53, propionic acid (as COD) 508 ( 24, soluble carbohydrate (as COD) 86 ( 4, and soluble protein (as COD) 931 ( 41. Set-up and Operation of Parent Sequencing Batch Reactors (SBRs). Two parent SBRs (A-SBR and F-SBR) were fed respectively with synthetic wastewater of acetic acid and sludge alkaline fermentation liquid and seeded with activated sludge from the above-described WWTP. The SBRs had a working volume of 4.0 L each and were operated under the same condition with three 8 h cycles per day. Each cycle consisted of 100 min anaerobic, 60 min aerobic, 45 min anoxic, 30 min aerobic, 45 min anoxic, 30 min aerobic, 45 min anoxic, and 15 min aerobic period, followed by 70 min settling, 10 min decanting, and 30 min idle phase. These SBRs were maintained at 21 ( 1 °C and controlled at a sludge retention time of approximately 13 days by wasting sludge before the end of the final aerobic phase. The MLSS and MLVSS were 3885 ( 180 and 3108 ( 143 mg/L in F-SBR, which were 3908 ( 169 and 3070 ( 136 mg/L in A-SBR. The DO concentration was about 6.3 ( 0.3 mg/L in aerobic phase. After the settling period, 3.0 L supernatant was discharged and replaced with 3.0 L fresh synthetic wastewater during the next initial 10 min of the anaerobic time. Each of the reactors was constantly mixed with a magnetic stirrer except during the settling, decanting, and idle periods. The influent pH of each reactor was controlled at 7.3 ( 0.2 by adding either 4 M NaOH or 4 M HCl. The synthetic wastewater included stock solutions called “P-water”, “N-water”, “concentrated solution”, and “traceelement solution (adapted from Smolders et al. (10))”, a certain amount of carbon sources, and tap water. The carbon source of A-SBR was pure acetic acid, and that of F-SBR was only sludge alkaline fermentation liquid. The stock solutions, carbon sources, and tap water were added daily to make a final concentration of approximately BOD 265, NH4+-N 35, and PO43--P 15 mg/L in the two SBRs. The characteristics of synthetic wastewater of the two SBRs are listed in Table 1. The “N-water” and “P-water” consisted of (g/L) 133.75 NH4Cl, 65.13 KH2PO4, and 97.26 K2HPO4 · 3H2O, respectively. The volume ratios of “trace-element solution”/synthetic wastewater and “concentrated solution”/synthetic wastewater were respectively 0.5/1000 and 2.5/1000. One liter of “traceelement solution” contained the following: 1.5 g of FeCl3 · 6H2O, 0.15 g of H3BO3, 0.03 g of CuSO4 · 5H2O, 0.18 g of KI, 0.12 g of MnCl2 · 4H2O, 0.06 g of Na2MoO4 · 2H2O, 0.12 g of ZnSO4 · 7H2O, 0.15 g of CoCl2 · 6H2O, and 10 g of ethylenediaminetetraacetic acid. The “concentrated solution” contained (per liter) the following: 20 g of MgSO4 · 7H2O, 9 g of CaCl2 · H2O, 33.94 g of MgCl2 · 6H2O, 25.88 g of peptone, and 4.24 g of yeast extract. Both A-SBR and F-SBR were initially cultured by acetic acid synthetic wastewater with a BOD concentration progressively increasing from around 80 to 265 mg/L over a 30 day period. The beginning ratios of BOD/P and BOD/N were respectively 14.1 and 5.3, which increased to a final value of 17.7 and 7.6 at BOD of 265 mg/L. Then the influent of F-SBR was gradually replaced by sludge fermentation liquid in the coming 21 d. After 60 and 81 d acclimatization, the net 8958
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removal of PO43--P and NH4+-N in A-SBR and F-SBR reached relatively stable, respectively, and then the data were reported. Effect of Main Organic Composition of Sludge Fermentation Liquid on Aerobic Nitrite Accumulation. Acetic acid, propionic acid, protein, and carbohydrate were the main organic composition of sludge alkaline fermentation liquid. In order to understand their effects on NO2--N accumulation under aerobic conditions, the following batch experiments were conducted respectively with synthetic wastewaters of acetic acid, acetic + propionic, acetic + propionic + protein (bovine serum albumin (BSA) was used as a model compound of protein in this study), and acetic + propionic + protein (BSA) + carbohydrate (glucose, a model carbohydrate). The fifth batch reactor, which received sludge fermentation liquid as the carbon source, was also run for comparison. Five min before the end of the final one aerobic phase, 2 L of biomass was taken from F-SBR, centrifuged at 1000 rpm for 5 min to remove the supernatant, washed the biomass three times with 0.9% NaCl solution, and resuspended in tap water with a final volume of 2 L. After the “N water” (see the above description) was added to achieve a final NH4+-N concentration of 11 mg of N/g-MLVSS, the aliquot was divided equally into five reactors with volume of 1 L each. Then different carbon sources were added (mg BOD/g-MLVSS): 85.5 acetic acid (reactor 1), 60.3 acetic plus 25.2 propionic (reactor 2), 32.7 acetic plus 13.7 propionic plus 39.1 BSA (reactor 3), 30.5 acetic plus 13.1 propionic plus 36.6 BSA plus 5.3 glucose (reactor 4), and 85.5 sludge fermentation liquid (reactor 5). The pH of the mixture in each reactor was adjusted to pH 7.3 ( 0.2 by the addition of 4 M NaOH or 4 M HCl. The sludge mixture in all reactors was stirred under anaerobic condition for 100 min and then aerated for 1 h with DO 6.0-6.6 mg/L. All reactors were maintained at 21 ( 1 °C. The changes of NO2--N, NO3--N, and nitrite accumulation ratio with aerobic time were investigated. Effect of Humic Acids and Metal Ions on Aerobic NO2--N Accumulation. Five min before the end of anaerobic phase, 2.4 L of biomass was taken from F-SBR. The mixture was centrifuged at 1000 rpm for 5 min to remove the supernatant, washed three times with 0.9% NaCl solution, and resuspended in tap water with a final volume of 2.4 L before divided equally into three reactors (1 L each). The “N water” (see the above description) was added to each reactor at dosage of 10.0 mg of NH4+-N/g-MLVSS. Then humic acids were added to one reactor (70.8 mg of humic acids/g-MLVSS). ZnSO4 · 7H2O, MnCl2 · 4H2O, and CuSO4 · 5H2O were added to another reactor at dosages of 0.041 mg of Zn/g-MLVSS, 0.0057 mg of Mn/ g-MLVSS, and 0.0014 mg of Cu/g-MLVSS. The third was run for comparison batch reactor. Three reactors were aerated (DO 6.0-6.6 mg/L), mixed with magnetic stirrers, and maintained at 21 ( 1 °C. The time courses of NH4+-N, NO2-N, and NO3--N concentrations were measured. Batch Experiments of PO43--P Uptake. Five min before the end of anaerobic phase, 2.4 L of biomass was taken from A-SBR or F-SBR. The mixture was centrifuged at 1000 rpm for 5 min to remove the supernatant, washed three times with 0.9% NaCl solution, and then resuspended in tap water with a final volume of 2.4 L before divided equally into three
reactors (1 L each). Then NaNO2 and “P water” (see the above description) were added to reactor-1 at a dosage of 8.4 mg of NO2--N/g-MLVSS and 17.3 mg of PO43--P/g-MLVSS, and NaNO3 and “P water” were added to reactor-2 at a dosage of 9.3 mg of NO3--N/g-MLVSS and 17.3 mg of PO43--P/gMLVSS. Only “P water” was added to reactor-3 (17.3 mg of PO43--P/g-MLVSS). N2 gas was introduced to reactor-1 and reactor-2 for 10 min, but reactor-3 was bubbled by oxygen gas with DO 6.0-6.6 mg/L during the entire experiment. All reactors were mixed by a magnetic stirrer and maintained at 21 ( 1 °C. Analytical Methods. The analyses of COD, BOD, total nitrogen (TN), NH4+-N, PO43--P, NO3--N, NO2--N, MLSS, and MLVSS were conducted in accordance with Standard Methods (11). Humic acids were estimated according to ref 12. Protein, carbohydrate, short-chain fatty acids (including acetic, propionic, n-butyric, iso-butyric, n-valeric, and isovaleric acids), poly-3-hydroxybutyrate (P3HB), poly-3-hydroxyvalerate (P3HV), and poly-3-hydroxy-2-methylvalerate (P3H2MV)) were assayed according to the method described in our previous publications (8, 13). The total number of measured acetic, propionic, n-butyric, iso-butyric, n-valeric, and iso-valeric acids and the sum of measured P3HB, P3HV, and P3H2MV were recorded as the total short-chain fatty acids (TSCFA) and polyhydroxyalkanoates (PHA), respectively. The nitrite accumulation ratio (NAR) was calculated as the percentage of NO2--N accounting for the sum of NO2--N and NO3--N. For determining the activity of nitrite reductase (NIR), biomass was taken out of A-SBR and F-SBR at the end of the first aerobic phase, respectively, then washed three times with 0.1 M PBS (phosphate buffer solution, pH 7.8), and resuspended in PBS including 47 mM Na2S2O4 (sodium dithionite) and 0.01 M NaOH. The mixture was sonicated at 20 kHz and 4 °C for 5 min to break down the cells, and the crude extracts were then centrifuged at 12000 rpm and 4 °C for 10 min to remove the unbroken spheroplasts and waste debris. The extracts were kept cold on ice before they were used for enzyme activity assay. The measurement of NIR was according to the method described by Nagaoka et al. (14) with modification. The 2 mL eppendorf centrifuge tube contained 1.38 mL of PBS, 200 µL of Na2S2O4 (47 mM), 20 µL of KNO2 (3.6 mM), 100 µL of methyl viologen (10 mM), and 300 µL of crude enzyme extract, and they were added in the above listed order. The mixture was immediately incubated at 30 °C for 10 min after adding the crude enzyme extract. The NIR activity was expressed by the rate of net NO2--N reduction (µmol NO2--N/(min · g-MLVSS)). The FISH (fluorescence in situ hybridization) technique with 16S rRNA-targeted oligonucleotide probes was employed to monitor the difference of microbial community in A-SBR and F-SBR responsible for nitrogen and phosphorus removal. Table S1 (Supporting Information) lists the oligonucleotides probes, their specificities, and conditions of stringency used for FISH in this study (3, 15, 16). PAO462, PAO651, and PAO846 were applied together (PAOmix, for Rhodocyclus-related PAO) as well as EUB338, EUB338-II, and EUB338-III (EUBmix, for domain Bacteria), GAOQ431, GAOQ989, and GB_G2 (GAOmix, for Candidatus Competibacter Phosphatis), NIT3 and Ntspa662 (for Nitrobacter spp. and genus Nitrospira), and Nso190 and Nso1225 (for ammonia-oxidizing Betaproteobacteria). These probes were commercially synthesized and 5′ labeled with AMAC or TAMRA. The biomass withdrawn from two reactors at the end of anaerobic phase was fixed with 4% freshly prepared paraformaldehyde for 8-10 h at 4 °C. After being rinsed with PBS (pH 7.2), 10 µL samples were immobilized on gelatin coated glass slide and then dehydrated by successive ethanol solution 50%, 75%, 85%, and 98% each for 3 min before dried in the air. Hybridizations on the slide glass were performed according to the standard
method of FISH (17) with slight modifications. Different probes required the addition of different amounts of deionized formamide (DFA, Table S1) to the hybridization buffer (0.9 M NaCl, 20 mM Tris-HCl (pH 7.2), 0.01% SDS (sodium dodecyl sulfate), and 0.2 ng probes). Twenty µL of hybridization buffer was hybridized with the fixed samples, and then the slides were incubated in a prewarmed Boekel InSlide Out Hybridization Oven (Boekel Scientific Inc.) at 46 °C for 2 h. All hybridization experiments were followed by a washing step at 48 °C for 20 min in a washing buffer containing TrisHCl (20 mM, pH 7.2), NaCl (70 mM), EDTA (5 mM), and SDS (0.01%), and the NaCl concentration was fixed by experiments. The washing buffer was removed by rinsing the slides with sterile water and the slides were air-dried. The slides were mounted to avoid bleaching the visual pigments and examined with epifluorescence microscope (Nikon, Japan). Within each field, total sludge cell area targeted by each applied probe was expressed as a percentage of the domain cell area targeted by the bacterial probe EUBmix using the functions provided in the image analyzing software (ImagePro Plus, V6.0, Media Cybernetics); signal intensity threshold was determined by negative control. The total genomic DNA was extracted using the method described by Purkhold et al. (18) with minor modification. Before microbial community analysis, biomass taken from A-SBR and F-SBR at the end of anaerobic phase was centrifuged for 5 min at 10000 rpm, respectively, and the centrifuged precipitate (0.5 g, wet weight) was washed three times with STET buffer (8% Sucrose, 0.1% Tween-20, 50 mM EDTA (pH 8.0), 50 mM Tris (pH 8.0)) and then resuspended in STET buffer. The lysozyme (50 mg/mL) was added to the mixture and incubated for 1 min at 90 °C. Then, 10% SDS and 20 mg/mL proteinase K were added, followed by incubation at 37 °C for 60 min. Five M NaCl and 10% cetyltrimethylammonium bromide were added and incubated at 65 °C for 10 min. After that, Tris-saturated phenol, phenol-chloroform-isoamyl alcohol (25:24:1), and chloroform were respectively used to extract the nucleic acids at 13000 rpm for 10 min, and the extract was then precipitated by incubation with sodium acetate (3 M, pH 5.2) and cold ethanol (100%) overnight at -20 °C. On the next day the mixture was centrifuged at 13000 rpm for 10 min, and the precipitate was washed with cold ethanol (70%), dried at room temperature after centrifugation, and finally resuspended in sterile deionized water. The amount and purity of DNA were determined spectrophotometrically at 260 and 280 nm. The nearly complete 16S rRNA gene fragments of bacteria were amplified using the bacterial primer 27F (5′-AGAGTTTGATCCTGGCTCAG-3′) and 1492R (5′-GGTTACCTTGTTACGACTT3′) (19). The polymerase chain reaction (PCR) amplification was carried out in a total volume of 25 µL containing 10 ng of template DNA, 1 × Ex Taq reaction buffer, 2U Ex Taq polymerase, 3.0 mM MgCl2, 0.2 mM dNTPs, and each primer at a concentration of 0.5 µM by using the StepOne Real-Time PCR System (Applied Biosystems, Foster City, CA). The amplification program consisted of an initial denaturation step (5 min at 94 °C), 30 cycles of denaturation (30 s at 94 °C), annealing (30 s at 55 °C), and extension (60 s at 72 °C), and then a final extension step at 72 °C for 5 min. The amplified DNA was purified using the TaKaRa Agarose Gel DNA Purification Kit (TaKaRa, Japan), ligated to the pMD19-T vector (TaKaRa, Japan), and transformed into E. coli DH5R cells (TaKaRa, Japan) according to the manufacturer’s instructions. Clones were sequenced by Chinese National Human Genome Center (Shanghai, China) using ABI PRISM 3730 automated DNA sequencer (Applied Biosystems, Foster City, CA) and then grouped into operational taxonomy units (OTUs) according to the 97% similarity threshold (20). The closest matching sequences in the GenBank database were searched using the BLAST program (21). Multiple alignments VOL. 44, NO. 23, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 1. Variations of NH4+-N and nitrite accumulation ratio (a), TN and NO3--N (b), NO2--N and PO43--P (c), and P3HB, P3HV, P3H2MV and PHA (d) during one cycle in two SBRs. 0-100 min (anaerobic), 100-160 min (aerobic), 160-205 min (anoxic), 205-235 min (aerobic), 235-280 min (anoxic), 280-310 min (aerobic), 310-355 min (anoxic), and 355-370 min (aerobic). All the standard deviations of triplicate tests are less than 9%. were generated using the ClustalX 2.0 (22), and then the phylogenetic tree was constructed with MEGA 4.0 (23) using the Jukes-Cantor model for the neighbor-joining algorithm (21). Accession Numbers. The nucleotide sequences reported in this paper have been deposited in the GenBank, EMBL, and DDBL nucleotide database under the accession numbers HQ010696-HQ010847.
Results and Discussion Profiles of F-SBR and A-SBR. The transformations of nitrogen (ammonia nitrogen, nitrite, nitrate, NAR, TN), phosphorus, and PHA (P3HB, P3HV, P3H2MV, PHA) during one cycle in two SBRs are shown in Figure 1, and the removal efficiencies of PO43--P, NH4+-N, TN, BOD, and soluble carbohydrate and protein are listed in Table 1. It was observed that in anaerobic time the consumption of all SCFA was completed within the initial 40 min (Figure S1, Supporting Information). As seen in Figure 1, PO43--P release and PHA accumulation were observed in both A-SBR and F-SBR. One important feature of Figure 1 was that F-SBR had a greater aerobic NO2--N but lower aerobic NO3--N accumulation than A-SBR. For 8960
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example, in the first aerobic phase the maximum NO2--N concentration and the corresponding nitrite accumulation ratio in F-SBR were 5.01 mg/g-MLVSS (or 15.48 mg/L) and 81.8%, whereas they were 1.52 mg/g-MLVSS (4.65 mg/L) and 40.9% in A-SBR, respectively. During one cycle 53.0% and 18.9% of the TN were respectively removed via nitrite and nitrate in F-SBR, which were 16.7% and 32.1% in A-SBR. Apparently, the use of sludge fermentation liquid, compared with that of acetic acid, as the carbon source of BNR significantly enhanced the short-cut nitrificationdenitrification. Another feature of Figure 1 was that the concentrations of PO43--P, NO2--N NO3--N, and PHA in F-SBR decreased in the anoxic phases, which indicated that anoxic denitrifying PO43--P removal occurred. For example, there were an average 4.24, 1.11, and 4.21 mg/g-MLVSS of NO2--N, NO3-N, and PO43--P removal and 0.28 mmol-C/g-MLVSS of PHA degradation during the first anoxic phase in F-SBR. However, no significant anoxic PO43--P removal was observed in A-SBR. During one cycle 21.2% PO43--P was removed via anoxic denitrification in F-SBR, but no obvious anoxic PO43--P uptake occurred in A-SBR. Thus, it can be concluded that
the removal of phosphorus and nitrogen in A-SBR was by aerobic phosphorus uptake and normal nitrification and denitrification, but it was achieved mainly by short-cut nitrification-denitrification and denitrifying phosphorus removal in F-SBR. The data in Table 1 showed that the use of sludge fermentation liquid, compared with acetic acid, increased the removal efficiencies of PO43--P and TN without any detectable soluble protein and carbohydrate (the main organic composition of sludge fermentation liquid) left in the effluent (Table 1). In the next text the mechanisms for sludge fermentation liquid showing significantly higher aerobic nitrite accumulation and anoxic denitrifying phosphorus removal via nitrite than acetic acid were investigated. Mechanisms of Sludge Fermentation Liquid Improving Short-Cut Nitrification-Denitrification and Denitrifying Phosphorus Removal via Nitrite. In the literature high concentration (500-3000 mg/L) of short chain fatty acids, such as acetic and propionic acids, have been reported to affect the nitrite formation during nitrification (24, 25). In this study acetic acid, propionic acid, protein, and carbohydrate were the main organic compositions of sludge fermentation liquid, and their effects on aerobic nitrite accumulation were therefore first investigated. According to the batch experimental results (Figure S2, Supporting Information)), the nitrite accumulation ratios with the carbon source of acetic, or acetic + propionic, or acetic + propionic + protein, or acetic + propionic + protein + carbohydrate were almost the same although all of them were lower than that with sludge fermentation liquid. It seems that it was not the main organic composition causing different nitrite accumulation in F-SBR. Several other factors, such as pH, temperature, dissolved oxygen (DO), free ammonia (FA), humic acids (HA), and metal ions, have also been reported in literatures to give impact on nitrite build-up during nitrification (1, 4, 5, 26-28). In this study, the pH increased gradually from 7.47 to 8.25 in A-SBR and from 7.46 to 8.31 in F-SBR (Figure S3, Supporting Information). It seems that two SBRs had almost the same pH variations. Also, two SBRs were operated under the same temperature and DO conditions. Thus, pH, temperature, and DO were not the factors which caused higher NO2--N accumulation in F-SBR. Although FA have been reported to affect nitrite accumulation during nitrification, a significant amount of NO2--N will not be accumulated only after the FA concentration is around 5 mg of NH3/L (26). During the aerobic phases of the current study the maximal FA concentrations, calculated according to Villaverde et al. (4), were only respectively 0.62 and 0.51 mg of NH3/L in A-SBR and F-SBR, which indicated that greater short-cut nitrification
observed in F-SBR was not caused by free ammonia. In addition, Loveless and Painter (27) found that some metal ions, such as Zn2+ and Cu2+, affected the growth of ammonia oxidizing bacteria and nitrite accumulation. In this study although the concentrations of Zn2+, Mn2+, and Cu2+ were respectively 0.041, 0.0057, and 0.0014 mg/g-MLVSS at the end of anaerobic phase in F-SBR, the synthetic wastewater batch experiments showed that the accumulated NO2--N was respectively 2.97 ( 0.18 and 2.78 ( 0.21 mg/g-MLVSS with and without metal ions addition. The presence of small amount of metal ions in sludge fermentation liquid gave little influence on aerobic nitrite accumulation. Some researchers reported that the existence of humic acids (HA) increased NO2--N accumulation during nitrification in biofilm or soils (28). Humic acids mainly include fulvic acids, brown humic acids, and black humic acids. Our previous study showed that there were humic acids in the WAS alkaline fermentation liquid (29). Their concentrations in the influent of F-SBR and A-SBR were respectively 74.8 and 4.2 mg/g-MLVSS and decreased to 70.5 and 0.7 mg/gMLVSS at the end of anaerobic phase. Further batch experiments with synthetic wastewater (the HA concentration was almost the same as that measured at the end of anaerobic phase of F-SBR) showed that the addition of HA caused more NO2--N accumulation (Figure S4, Supporting Information). It has been reported that nitrite oxidizing bacteria (NOB) were more sensitive to humic acids inhibition than ammonia oxidizing bacteria (AOB) (28). The more HA in sludge fermentation liquid SBR the more NO2--N accumulation was therefore observed. The mechanisms of different nitrite accumulation were further investigated using the 16S rRNA gene clone libraries and FISH to describe the structure of microbial communities and the quantities of AOB and NOB in two SBRs. The phylogenetic trees based on the 152 clones obtained from the two 16S rRNA clone library were constructed and shown in Figure S5 and Figure S6 (Supporting Information). The OTUs from the A-SBR clone library were mainly affiliated to the γ- and β-proteobacteria, while in F-SBR clone library β-proteobacteria and Bacteroidetes were the two most important groups. The OTU13 and OTU14, related to the ammonium oxidizers of the Nitrosomonas sp. and the nitrite oxidizers of the Candidatus Nitrotoga arctica, respectively, were observed in F-SBR clone library (18, 30). OTU3 and OTU5 in A-SBR clone library were affiliated to the ammonium oxidizers of the Nitrosospira sp. and the nitrite oxidizers of the Nitrospira sp., respectively (31). The specific ratios of AOB and NOB accounting for the domain bacteria in two SBRs were further investigated with
FIGURE 2. Microscopes of sludges from A-SBR (A1-A3) and F-SBR (F1-F3) at the end of anaerobic phase as visualized by FISH. Ammonia-oxidizing β-Proteobacteria was hybridized with AMCA-labeled Nso1225 and Nso190 (blue), nitrite oxidizing bacteria containing Nitrospira spp. and the genus Nitrobacter was hybridized with TAMARA-labeled Ntspa662 and NIT3 (red), and probe EUBmix specially stained domain Bacteria labeled with 6-FAM (green). Further images analysis showed that ammonia oxidizing and nitrite oxidizing bacteria were 13.8 and 4.2% in F-SBR and 7.4 and 10.2% in A-SBR. VOL. 44, NO. 23, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FISH combined with epifluorescence microscope. The fluorescence micrographs of sludge hybridized with Nso190 and Nso1225 (probes targeted to AOB) and NIT3 and Ntspa662 (probes targeted to NOB) oligonucleotide probes and the fractions of nitrifying bacteria composition (analyzed by Imagine plus 6.0) are demonstrated in Figure 2. The results showed that F-SBR had more AOB than A-SBR (13.8% against 7.4% accounting for the domain bacteria), but the former had less NOB than the latter (4.2% and 10.2%), which might be one reason for greater nitrite accumulation in F-SBR. Finally, the reasons for F-SBR showing higher denitrifying phosphorus removal via nitrite than A-SBR were explored. Bacteria closely related to Rhodocyclus sp., Candidatus Accumulibacter sp., and Candidatus Competibacter phosphatis were observed in both F-SBR and A-SBR clone libraries (Figure S5 and Figure S6, Supporting Information). The Rhodocyclus sp. and Candidatus Accumulibacter sp. were reported as PAOs (phosphate accumulating organisms) in the literature (15), while the Candidatus Competibacter phosphatis are thought to be a representative of GAOs (glycogen accumulating organisms), unfavorable bacteria in biological phosphorus removal systems (3). Moreover, FISH probes targeted to PAOs and GAOs were used to determine the PAOs and GAOs fractions accounting for the domain bacteria in F-SBR and A-SBR, respectively. The fluorescence micrographs of sludge hybridized with PAOmix and GAOmix oligonucleotide probes were demonstrated in Figure S7 (Supporting Information). The images analysis showed that the fractions of PAOs in F-SBR and A-SBR were respectively 58.5% and 37.2%, while the corresponding GAOs were 3.1 and 10.5%. All these observations were in accordance with the trends of phosphorus removal performance in two SBRs. PAOs can be further divided into three types according to Hu et al. (32): PO (those utilizing only O2 as electron acceptor), PON (those utilizing both O2 and NO3--N as electron acceptors, but not NO2--N), and PONn or DNPAOs (those utilizing O2, NO3--N and NO2--N as electron acceptors). PO, PON, and PONn (DNPAOs) were obtained by the batch PO43--P uptake experiments, and their values accounting for total PAOs (PO + PON + PONn) are shown in Table S2 (Supporting Information). It can be seen that the PO fraction in A-SBR (90.8%) was the greatest one, indicating that most of the PAOs only used O2 as the electron acceptor. The DNPAOs percentage in F-SBR, however, was the greatest one (60.7%), which was about 9.2 times higher than that in A-SBR (60.7% versus 6.6%). Thus, F-SBR had greater anoxic denitrifying PO43--P removal than A-SBR. Further investigation showed that the activity of nitrite reductase in F-SBR was almost 3fold of that in A-SBR (4.47 versus 1.45 µmol of NO2--N/ (min · g-MLVSS)), suggesting that the short-cut denitrification ability of F-SBR sludge was much higher than that of A-SBR sludge. It can be easily understood that the use of sludge fermentation liquid as carbon source could achieve higher anoxic denitrifying phosphorus removal via nitrite than acetic acid.
Acknowledgments This work was supported by the National Hi-Tech Research and Development Program (863) (2007AA06Z326), the Foundation of State Key Laboratory of Pollution Control and Resource Reuse (PCRRK09002), and the key projects of National Water Pollution Control and Management of China (2008ZX07315-003).
Supporting Information Available Tables S1-S2 and Figures S1-S7. This material is available free of charge via the Internet at http://pubs.acs.org. 8962
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