Introduction
Biological phosphorus removal (BPR) is most economical and sustainable wastewater phosphorus (P) removal strategy [1–3]. After the excessive P uptake and polyphosphate (polyP) synthesis of phosphorus accumulating organisms (PAOs), the P removal from wastewater is achieved by discharging P-riched excess sludge [3,4], in which the aerobic/anoxic synthesis and storage of intracellular polyP was considered as a key biochemical link in the formation of P-riched sludge [1,4]. On the other hand, it had been revealed that extracellular polymer substances (EPS) were a P reservoir of BPR sludge, of which the total phosphorus (TP) content accounted for 10%–70% of the sludge TP content [5–9]. Moreover, extracellular polyP presented the anaerobic-decomposition/ aerobic-synthesis characteristics [10–12]. Therefore, the combined actions of PAOs and EPS were involved in BPR process [12]. However, there were significant differences in the previous reports about the percentages of the extracellular TP content to the sludge TP content [5–9], ascribed to the different extraction methods and sludge sources. Thereby, the relative magnitude of EPS in the P storage of BPR sludge was still ambiguous.
The EPS are mainly comprised of polysaccharide, protein, humic acid and DNA [13], in which the electronegative groups (e.g. carboxyl, hydroxyl, aldehyde, ester, amino and sulfhydryl) are adsorbed or bounded with metal cations (e.g. K+, Mg2+ and Ca2+) through ionic bonding, electrostatic adsorption or bridging action [13]. The main P species in the EPS of BPR sludge are orthophosphate (orthoP) and polyP [6,9], which are also bounded with metal cations through ionic bonds, to form active polyP (e.g. K-polyP, Mg-polyP), inert polyP (e.g. Ca-polyP) and phosphate precipitation (e.g. MgNH4PO4, Ca5 (PO4)3 (OH)) [7,14]. Therefore, the interactions among EPS macromolecules, metal elements and phosphates should be involved in the P adsorption of EPS. However, there were few reports about the effects of the EPS yield, composition and properties, and the metal element content and type on the extracellular P content, thus the understand about the P adsorption effect of EPS was not in-depth.
There were significantly differences in weakening or destroying the actions or forces in EPS with different extraction methods [13], such as heating, cation exchange resin (CER), EDTA and formaldehyde-NAOH, etc., thus the extraction efficiencies of EPS and extracellular P were obviously distinguished [6,9,15,16]. Additionally, the degrees of extracellular polyP hydrolysis induced by acid or alkaline were different during the extraction processes with different extraction methods, such as the H2SO4, NaOH and formaldehyde-NaOH, affecting the content and average chain length of extracted extracellular polyP [17]. Ultrasonic-CER is a reliable and efficient method to extract EPS and extracellular P [6,17], with the following technical features: i) avoiding polyP hydrolysis with low frequency (e.g., 21 kHz) ultrasonic or CER; ii) non chemical contamination; iii) avoiding cell damage/lysis through the optimization of extract parameters; iv) Improving CER extraction efficiency obviously through ultrasonic dispersion. Furthermore, the outer loosely bound and inner tightly bound EPS (i.e. LB-EPS and TB-EPS) could be fractionally extracted by ultrasonic-CER method, which had been used to clearly elucidate the migration and transformation characteristics of extracellular P [11,12]. However, the extraction efficiency of hydrophobic EPS may be relatively lower with ultrasonic-CER method, and thus the adsorption of P by EPS may be underestimated. On the other hand, the P in distinguishable outer-layer EPS of BPR sludge flocs had been in-situ observed by scanning electron microscope combined with energy dispersive spectrometer (SEM-EDS) [18]. If the residual EPS and extracellular P after the extraction of LB-EPS and TB-EPS are observed by SEM-EDS, the P adsorption effect evaluation of EPS could be more accurate.
The influences of sludge sources on the EPS and extracellular P contents could be attributed to the differences in environmental factors and operating parameters (i.e., solid residence time (SRT), dissolved oxygen (DO), etc.) [19–22], but the observations about the influences of them on the P adsorption effect of EPS were few. SRT is an important design and operation parameter [1]. For example, considering the biological nutrient removal (BNR) or nitrification, the SRTs of BPR systems should be more than 10 d [23]. The SRTs of BPR systems coupled with membrane bioreactor (MBR) were usually more than 20 d [5,24]. However, the BPR performance might be worse when increasing SRT, because of the decrease of excess sludge discharge [21,25]. On the other hand, the sludge TP content could be enhanced with the increase of SRT [5]. The successful operation of BPR systems with the SRT greater than 20 d had been reported [5,24,26]. At present, the influences of SRT on BPR performance were not completely clear, mainly because: i) the observations about the effect of SRT on PAOs dominance were few and even contradictory [21,23,27]; ii) the improvement of P storage capacity of sludge with the increase of SRT is controversial [5,24,26–28], and in particular, the influence of SRT on the P adsorption effect of EPS was still unclear. EPS mainly came from microbial metabolic secretion and lysis [13], of which the yield, composition and properties are affected by microbial community structure and metabolic activity. In addition, EPS are biodegradable and can be carbon sources utilized by microorganisms [13], thus affecting microbial community structure. Thereby, it is necessary to deeply investigate the influence of SRT on the P adsorption of EPS from the perspective of the yield, composition, and properties of EPS as well as microbial community structure, to comprehensively understand the effect of SRT on BPR performance.
In the study, the 3 lab-scale anaerobic/aerobic sequencing batch reactors (A/O-SBR) respectively with SRTs of 12 d, 18 d, and 25 d were adopted. The LB-EPS, TB-EPS and bacterial cells in the P -rich sludge sampled at the end of aerobic stage were fractionally extracted by the ultrasonic-CER method. The influence of SRT on the P adsorption of EPS were deeply studied from the perspective of the yield, composition, and properties of LB-EPS and TB-EPS, and the metal element content and type as well as the microbial community structure. The main aims were that i) the relative magnitude of EPS in the P storage of BPR sludge was objectively evaluated by the P contents in LB-EPS and TB-EPS extracts and the subsequent SEM-EDS observation, and ii) the influence of SRT on the P adsorption effect of EPS was elaborated through analyzing the P and metal elements adsorption properties by LB-EPS and TB-EPS and discussing the key microbial populations and their effects on the extracellular polyP content and average chain length.
Materials and Methods
2.1. Control and Operation of Reactors
The sodium acetate and sodium propionate with the COD ratio of 3:1 were applied as the carbon sources for the 3 lab-scale anaerobic-oxic sequence batch reactors (A/O-SBR), with the influent COD concentration of 800 mg/L and the COD:N:P of 100:4:5. The 3 reactors were operated for two 12-h cycles per day, consisting of instantaneous influent, 3.5-h anaerobic period, 7.5-h aerobic period, 50-min settling period, 5-min decanting period and 5-min idle period. The water temperatures of the reactors were 20±1°C, the pH ranges were 7.4–8.8, and the DO concentrations at the end of aerobic stage were 2.5–4.0 mg/L. Initially, the SRTs of the 3 reactors were all controlled at about 18 d. Subsequently, the mixing liquids of 3 reactor were homogeneously mixed and even distributed, when was recorded as day 0. Thereafter, the 3 reactors were tuned to operate at the SRTs of 12 d, 18 d and 25 d, respectively, labeled as SBR-1, SBR-2, and SBR-3. The effluent TP concentrations, suspended solid (SS) concentrations, volatile suspended solid (VSS) concentrations, solid volume index (SVI) and sludge TP contents were monitored regularly. After about 75 d of monitoring the indicators, the 3 reactors all reached relative steady states. As shown in Table 1, the observed operational processes of the 3 reactors were divided into the transition phase (Phase I, day 0-day 74) and steady phase (Phase II, day 75-day 140). During the Phase II, the fraction extraction of the LB-EPS, TB-EPS, and bacterial cells, and the characterization of the extracted LB-EPS and TB-EPS as well as and the analysis of microbial community structure were carried out.
2.2. Extraction and Physicochemical Analysis of LB-EPS, TB-EPS and Intracellular P
According to the methods detailed by the previous studies [6,12], the LB-EPS and TB-EPS in sludge were fractionally extracted through the 21 kHz ultrasonic (JY90-II, ScientzBioscience Co., Inc., Ningbo, China) with power density of 1 W·mL−1 and the 001 × 7 Type CER (20–40 mesh, Suqing Co., Jiangsu, China) with stirring rate of 550 rmp. After extracting TB-EPS, the centrifugated precipitates were regarded as bacterial cells, which was re-suspended with cryogenic ultrapure water (2±1°C). Subsequently, the intracellular P was extracted by the ultrasonic with high power density of 10 W·mL−1, and then the high-speed cryo-centrifugation at 20,000×g was carried out.
After the digestion with alkaline potassium persulfate, the sludge TP content, and the TP contents (in mg·g−1 VSS) of LB-EPS and TB-EPS were determined by vanadium molybdate blue colorimetric method [29], and then intracellular TP content was calculated by differential subtraction method. The P species in the extracts of LB-EPS, TB-EPS and intracellular P were analyzed by liquid-phase 31P nuclear magnetic resonance (NMR) spectrometry (600 MHz NMR, Agilent Technologies, USA), in which the contents of different species P and the average polyP chain length were calculated according to the peak areas [17]. After the digestion with HNO3-H2O2, the K, Mg, Ca, and Fe contents in the sludge, LB-EPS extract and bacterial cell suspensions were determined by inductively coupled plasma-Optical emission spectrometry (5100 ICP-OES, Agilent Technologies, USA) [30], and then the K, Mg, Ca, and Fe contents in TB-EPS extract were calculated by differential subtraction method. The TOC concentrations of the LB-EPS and TB-EPS extracts were analyzed using a TOC analyzer (Multi N/C 2100S, Analytik Jena AG, Germany), characterizing the contents of LB-EPS and TB-EPS (in mg·g−1 VSS). The polysaccharide contents of LB-EPS and TB-EPS extracts were determined by anthrone sulfate method [31]. The three-dimensional fluorescence spectra (3DEEM) of LB-EPS and TB-EPS extracts were analyzed with a fluorescence spectrophotometer (FP-6500, JASCO, Japan).
2.3. Relative Magnitudes of Ionic Bonds and Hydrogen Bonds in LB-EPS and TB-EPS
The relative deflocculation rates of the sludges under the 2 kinds of liquid-phase environmental conditions (i.e., 0.5 mmol·L−1 EDTA and 6 mmol·L−1 urea) were investigated through modifying the method of Liao et al. [32], to characterize the relative magnitudes of ionic bonds and hydrogen bonds in EPS (including LB-EPS and TB-EPS). The sludges were sampled at the end of aerobic phase, and then the VSS concentrations were adjusted to 2000±50 mg/L by decanting the sedimentation supernatant or replenishing the centrifugal supernatant. The regulated sludge suspensions were divided into two equal parts. One was directly used for measuring the relative deflocculation rate, and the other was dispersed by 21 kHz ultrasonic with power density of 1 W·mL−1 for 3 min, followed by testing the relative deflocculation rate, in which the disperse parameters were the same as those of LB-EPS extraction. The variations of particle size distribution of sludges from the 3 reactors after ultrasonic were shown in Table S1. During the sludge deflocculation trial, a control group was supplemented with an equal volume of ultrapure water, while a treatment group was supplemented with an equal volume of 1 mmol·L−1 EDTA or 12 mmol·L−1 urea solutions, corresponding to the VSS concentration of approximately 1000 mg/L, and the EDTA concentration of 0.5 mmol·L−1 solution or the urea concentration of 6 mmol·L−1. Subsequently, both the control and treatment groups were oscillated in an ice-water bath at 200 rpm, while the median particle size (D50) variations of sludge flocs in the 2 groups were recorded by a laser diffraction particle sizer (SALD-2300, Shimadzu (Hong Kong) Ltd., China). The relative deflocculation rate was calculated as shown as Eq. (1).
2.4. Microscopic Morphology Observation and Energy Spectrum Analysis During the Fractional Extraction Process
A field emission scanning electron microscope (Quanta TEG 250, FEI, USA) equipped with an Apollo XLT probe X-ray spectrometer (Octane SDD, EDAX, USA) was used to observe the microscopic morphology of sludge during the process of fractional extracting LB-EPS and TB-EPS, and to analyze the elemental compositions in the residual TB-EPS. The samples were sharply cooled by liquid nitrogen, followed by freeze-dry. The dried samples were spray coated with gold, using a magnetron sputtering vacuum coater (Q 150RS, QUORUM, UK). Finally, the treated samples were observed in randomly selected fields of view at an accelerating voltage of 30 kV and a magnification of ×20,000 times.
2.5. 16s rRNA Gene Amplification and Sequencing
DNA templates were extracted from sludge samples at the end of aerobic phase on day 75 and day 135, with a soil genomic DNA extraction kit (DP336, Tiangen Biochemical Technology, China). Subsequently, DNA templates were amplified by a polymerase chain reaction (PCR) system (GeneAmp 9700, Thermo Fisher Scientific, Singapore), in which the amplification primers were the universal primers of 338F (ACTCCTACGGGGAGGCAGCAG) and 806R (GGACTACHVGGGTWTCTAAT) for the V3–V4 region [33]. The amplified product was paired-end sequenced on the Illumina MiSeq PE300 platform by Shanghai Majorbio Bio-pharm Technology Co., ltd. The bioinformatics analysis was performed on the Majorbio Cloud Platform (www.i-sanger.com). The effective sequences at identity ≥ 97% were clustered into the same operational taxonomic units (OTUs) by UPARSE (v.11), after reassembling and filtrating the raw sequences. The taxonomy of each OTU was analyzed by RDP Classifier (v.2.13) against the 16S rRNA database (Silva v.138). Principal component analysis was carried out through Bray-curtis distance to assess the microbial community discrepancy. Additionally, the heat map of the relative abundances of of putative PAOs and GAOs OTUs was drawn using Origin software (OriginPro 2021b SR1, OriginLab, USA).
Results and Discussion
3.1. P Content, Distribution and Species in Sludge
From Table 1, during the steady phase (i.e., Phase II), the P removal efficiencies of the SBR-1, SBR-2 and SBR-3 reactors were 96.1 ± 0.7%, 95.8 ± 0.5% and 90.0 ± 3.7%, respectively. The P removal efficiencies of the 2 reactors with the SRTs of 12 d and 18 d were close, while that with the SRT of 25 d was slightly lower. At the end of aerobic stage, the sludge TP contents in the 3 reactors were 93.5±6.6, 103.6±5.8 and 128.3±5.8 mg P·g−1 VSS, respectively. Compared with the SRTs of 12 d and 18 d, the P storage of sludge with the SRT of 25 d was significantly improved.
From Fig. 1, the P of sludge was mainly distributed in EPS, of which the TP contents accounted for 63.7%–68.1% of the sludge TP contents. Furthermore, the P of EPS was mainly distributed in TB-EPS, of which the TP contents were 57.2%–65.0% of the sludge TP contents. The TP contents in LB-EPS were low, of which the main species was orthoP. There was little polyP with the short average chain length of 2.2–4.0 in LB-EPS. The TP contents in LB-EPS were decreased with the increase of SRT, while those in TB-EPS were increased. The main P species in TB-EPS was polyP, of which the contents and average chain lengths were increased with the increase of SRT. The TP contents in TB-EPS for the SBR-1, SBR-2 and SBR-3 sludges were 53.5 ± 5.1, 60.9 ± 7.8 and 83.3 ± 3.6 mg P·g−1 VSS, respectively, with the extracellular polyP contents of 50.3 ± 6.9, 55.4 ± 8.9 and 81.7 ± 4.2 mg P·g−1 VSS and the average chain lengths of 199.1 ± 2.4, 207.8 ± 8.1 and 227.5 ± 11.6. Additionally, the TP contents in bacterial cells were 33.9 ± 6.6, 35.2± 3.8 and 40.9 ± 3.0 mg P·g−1 VSS, respectively, with the intracellular polyP contents of 32.5 ± 6.7, 33.2 ± 4.3 and 40.1 ± 3.5 mg P·g−1 VSS and the average chain lengths of 203.1 ± 8.6, 221.9 ± 2.6 and 229.7 ± 8.8. Compared with the SRTs of 12 d and 18 d, the significant P storage enhancement of sludge with the SRT of 25 d was mainly attributed to the increases of the extracellular TP content, polyP content and its average chain length in TB-EPS.
3.2. Content, Distribution, Compositions and Molecular Interactions in EPS
From Fig. 2, the TB-EPS contents in sludge from the SBR-1, SBR-2 and SBR-3 reactors were 90.3±4.2, 98.5±2.8 and 122.8±5.0 mg TOC·g−1 VSS, respectively, accounting for 91.2% – 95.2% of the EPS contents, indicating that TB-EPS was the main component of EPS. With the increase of SRT, the contents of EPS and TB-EPS were increased, but that of LB-EPS was reduced, which could be ascribed to that TB-EPS mainly came from microbial lysis and endogenous metabolism secretion, whilst LB-EPS mainly came from matrix utilization secretion. Compared with the SRTs of 12 d and 18 d, the TB-EPS yield of sludge with the SRT of 25 d was significantly increased. Furthermore, the compositions of LB-EPS and TB-EPS were characterized by the polysaccharide content in 1 mg TOC and 3DEEM spectrum of 1 mg TOC·mL−1. The polysaccharide contents in LB-EPS from the SBR-1, SBR-2 and SBR-3 sludges were 1.06±0.19, 0.93±0.18 and 0.76±0.19 mg·mg−1 TOC, respectively, while those in TB-EPS were 1.09±0.10, 0.73±0.04 and 0.56±0.03 mg·mg−1 TOC. As shown in 3DEEM spectra, the fluorescence signals of the III (i.e., fulvic acids) and V (i.e., humic acids) regions of LB-EPS from the sludges were weak, while those of TB-EPS were markedly enhanced by increasing SRT. The fluorescence intensities of the I (i.e., aromatic protein I) and II (i.e., aromatic protein II) regions of LB-EPS and TB-EPS were increased with the increase of SRT, in which the increase of TB-EPS was much greater than that of LB-EPS. Thus, the influences of SRT on the yield and compositions of TB-EPS were much greater than those of LB-EPS. In a word, the aromatic protein (i.e., I and II regions) and humic substances (i.e., III and V regions) contents of TB-EPS were much increased and the polysaccharide content was significantly reduced by increasing SRT. Compared with the SRTs of 12 d and 18 d, the aromatic protein content of TB-EPS with the SRT of 25 d was much higher.
There are much hydrophobic groups (e.g., phenyl, aromatic heterocyclic, alkyl, etc.) in aromatic proteins and humic substances [34], which are related with strong hydrophobic interactions, corresponding to weaker ionic and hydrogen bonds. However, there are much carboxyl, hydroxyl and other polar groups in polysaccharides, which are related with ionic and hydrogen bonds, corresponding to the weaker hydrophobic interactions [35]. According to the composition variations of LB-EPS and TB-EPS (Fig. 2), it could be inferred that: i) The relative magnitudes of ionic and hydrogen bonds in EPS (i.e., LB-EPS and TB-EPS) were decreased by increasing SRT, while that of the hydrophobic interaction was increased; ii) For a kind of sludge, the ionic and hydrogen bonds in TB-EPS were weaker than those in LB-EPS, while that of hydrophobic interactions in TB-EPS was stronger than that in LB-EPS. Furthermore, the relative magnitudes of ionic and hydrogen bonds in EPS were characterized through the relative deflocculation rate of sludge flocs under the different liquid environments, respectively [32]. As shown in Fig. S1, the relative deflocculation rates of the raw sludge flocs and dispersed sludge flocs under the 0.5 mmol/L EDTA or 6 mmol/L urea were decreased by increasing SRT. Compared with the SRTs of 12 d and 18 d, the relative deflocculation rates of the raw sludge flocs and dispersed sludge flocs at the SRT of 25 d were obviously lower, indicating that the relative magnitudes of ionic and hydrogen bonds in EPS (i.e., LB-EPS and TB-EPS) were much lower. Compared with the raw sludge flocs, the relative deflocculation rates of the dispersed sludge flocs were greater under the 0.5 mmol/L EDTA, because the EDTA mass transfer was improved after the ultrasonic dispersion, effectively weaking or destroying the ionic bonds in TB-EPS. The relative flocculation rates of the dispersed sludge floc were lower than those of the raw sludge floc under the 6 mmol/L urea for 15 and 30 min, suggesting that the relative magnitude of hydrogen bonds in TB-EPS should be less than that of LB-EPS.
3.3. K, Mg, Ca and Fe Content and Distribution in Sludge
From Fig. 3, the metal elements in these sludges were predominantly distributed in TB-EPS and bacterial cells, which was similar to the distributions of the P and polyP. The sludges from the 3 reactors, as well as TB-EPS and bacterial therein, had the highest K content, followed by the Mg and Ca contents, while the Fe contents were lowest. Therefore, the P in the TB-EPS and bacterial cells should be mainly active K-polyP and Mg-polyP. Additionally, the Ca content in the LB-EPS was more than the K and Mg contents, implying that much Ca-orthoP precipitate should be existed in the LB-EPS. The K, Mg, Ca, and Fe contents in the sludges, TB-EPS and bacterial cells were substantially reduced with the increase of SRT. For example, the K, Mg, Ca and Fe contents in TB-EPS at the SRT of 25 d were 55.7 ± 4.4%, 60.6 ± 8.1%, 55.5 ± 4.4% and 33.9 ± 11.0% of those at the SRT of 12 d, respectively. As shown in Fig. 2 and Fig. S1, the TB-EPS composition and the actions or forces in TB-EPS markedly changed by increasing SRT, such as the increase of the aromatic protein and humic substances contents, as well as the weakness of the ionic bonds and electrostatic adsorption, etc., which should be the key reason that the K, Mg, Ca and Fe contents in TB-EPS were reduced by increasing SRT.
3.4. Analysis and Evaluation of the P Adsorption Effect of EPS
In the study, the TP contents of EPS were 59.6±6.0 – 87.4±4.4 mg·g−1 VSS and accounted for 63.7% – 68.1% of those of sludge (Fig. 1), which was in the upper-range of the reported values [5–12]. After extracting LB-EPS and TB-EPS, the residues of TB-EPS and extracellular P were observed with SEM-EDS to evaluate the P adsorption effect of EPS. From Fig. 4, the outer cloud-like LB-EPS layer was extracted by ultrasonic, exposing the inner smooth and dense TB-EPS layer. The TB-EPS layer was significantly thinned by CER extraction, indicating that the TB-EPS had been effectively extracted. However, the surface of cell membrane was still surrounded by TB-EPS residues, protecting the cell-membrane integrity. There was a certain amount of P, K, Mg and other elements in the residual TB-EPS layer, indicating that the P adsorption effect of EPS was still underestimated. Compared with the SRTs of 12 d and 18 d, the residual TB-EPS layer at the SRT of 25 d was obviously thicker, suggesting that the underestimation of the extracellular P content at SRT of 25 d should be more than those at SRTs of 12 d and 18 d.
According to the prioritization of Ca2+>Mg2+>K+ ≈NH4+>Na+, the Ca2+ than the Mg2+ and K+ in TB-EPS was more easily exchanged by the Na+ in CER. As shown in Fig. 4c, 4f and 4i, the energy dispersive signal of Ca in the residual TB-EPS layer with the SRTs of 12 d and 18 d was very weak, while that with the SRT of 25 d was distinct, suggesting that the resistance of cation exchange between Ca2+ and Na+ in the TB-EPS with the SRT of 25 d should be greater than those with the SRTs of 12 d and 18 d. The Ca2+ was not only an important high-valent cation maintaining the structure stability of EPS three-dimensional hydrogel layer [13], but also bound with polyP or orthoP to form inert Ca-polyP or Ca-orthoP precipitate [10]. Compared with the SRTs of 12 d and 18 d, the hydrophobic interaction in the TB-EPS layer at the SRT of 25 d was stronger and the structure of the TB-EPS layer was denser (Fig. 2 and Fig. S1), with the significantly more Ca and TP contents in the residual TB-EPS layer (Fig. 4), which could be ascribed to that the extraction and its cation exchange process of the TB-EPS and extracellular P with the SRT of 25 d were more suppressed because of the greater resistance of cation exchange. In brief, the extracellular P adsorption effect of the TB-EPS with the SRT of 25 d should be significantly greater than those with the SRTs of 12 d and 18 d.
3.5. Influence of SRT on the P Adsorption of EPS
3.5.1. Influence of SRT on the P and metal element adsorption properties by EPS
The P and metal element adsorption properties by LB-EPS and TB-EPS were characterized with the contents in 1 mg TOC extract (in mg·mg-1 TOC), respectively. From Fig. 5, the P adsorption properties of LB-EPS for the SBR-1, SBR-2 and SBR-3 reactors were 0.699±0.107, 1.067±0.077 and 0.658±0.136 mg·mg−1 TOC, respectively, while those of TB-EPS were 0.593±0.054, 0.619±0.079 and 0.679±0.038 mg·mg−1 TOC, exhibiting that the former were fluctuated with the increase of SRT, while the latter were slightly increased. The K, Mg and Ca adsorption properties of LB-EPS were also fluctuated with the increase of SRT, while those of TB-EPS were reduced. In addition, the K/P, Mg/P and Ca/P molar ratios in TB-EPS were markedly reduced by increasing SRT. Specifically, the above ratios in TB-EPS for the SBR-1 sludge were 0.222±0.033, 0.239±0.015 and 0.127±0.010, respectively, while those for the SBR-3 sludge were 0.079±0.003, 0.093±0.017 and 0.045±0.004.
Only a small portion (e.g., carboxyl) of electronegative groups in EPS could be bound with metal cations (e.g., K+, Mg2+, and Ca2+) through ionic bonds, whilst the most (e.g., hydroxyl, aldehyde, ester, amino, sulfhydryl, etc.) of them were bounded with those through electrostatic adsorption and its bridging effect [13,32]. However, the phosphates (e.g., orthoP and polyP) were more preferentially bounded with metal cations than the electronegative groups of EPS, forming the extracellular active polyP, inert polyP and phosphate precipitate by ionic bonds, because the bonding actions or forces between the phosphates and the metal cations were significantly stronger than those between the electronegative groups of EPS and the metal cations. Furthermore, the groups of EPS were rarely directly bound with phosphates because of electrostatic repulsion, suggesting that the influence of TB-EPS compositions and properties on the P adsorption property was weak. But, the influence of them on the metal element adsorption properties was very strong. Compared with the SRTs of 12 d and 18 d, the metal element adsorption properties of TB-EPS at the SRT of 25 d was much lower, because of the less electronegative groups and the much stronger hydrophobic interactions in the latter TB-EPS (Fig. 2 and Fig. S1).
3.5.2. Influence of SRT on the bacterial community and the extracellular polyP
From Fig. 6, the bacterial community structure at genus level with the SRT of 25 d was obviously different from those with the SRTs of 12 d and 18 d. OLB8 in Saprospiraceae of Bacteroidetes is a typical saprophytic bacterium, which can degrade EPS [36]. The relative abundances of OLB8 in the SBR-1, SBR-2 and SBR-3 reactors at day 135 were 0.16 ± 0.01%, 0% and 14.16 ± 1.73%, respectively. Hydrogenophaga in Comamonadaceae of Proteobacteria tends to utilize amino acids or peptone as substrates [37], of which the relative abundances in the 3 reactors at day 135 were 0.12 ± 0.01%, 0.06 ± 0.01%, and 7.54 ± 0.04% in turns. The relative abundances of OLB8 and Hydrogenophaga with the SRT of 25 d were significantly higher than those with the SRTs of 12 d and 18 d, ascribed to the significantly higher endogenous metabolism level and the much higher TB-EPS aromatic protein content for the former (Fig. 2). Ca. Accumulibacter and Ca. Competibacter were the dominant PAOs and GAOs in the 3 reactors, respectively. The relative abundances of Ca. Accumulibacter in the SBR-1, SBR-2 and SBR-3 reactors at day 75 were 26.09 ± 1.05%, 39.78 ± 3.52% and 18.63 ± 1.31%, respectively, and at day 135 were 55.26 ± 3.55%, 51.46 ± 2.38% and 20.20 ± 4.64%. In addition, the relative abundance of Ca. Competibacter in the SBR-3 reactors at day 135 was 7.77 ± 0.64%. The Ca. Accumulibacter in the 3 reactors belonged to one OTU, while the Ca. Competibacter included 8 OTUs, in which the relative abundances of the 4 Ca. Competibacter OTUs with the SRT of 25 d were higher, such as OTU169, OTU185, OTU259 and OTU260. Although the Ca. Accumulibacter had always taken advantages over Ca. Competibacter, the loss risk of Ca. Accumulibacter dominance was increased when increasing SRT from 18 d to 25 d.
In this study, the results on the polyP contents and average chain lengths in LB-EPS and TB-EPS were reliable. The average polyP chain lengths in the TB-EPS with the SRTs of 12 d, 18 d and 25 d were 199.1 ± 2.4, 207.8 ± 8.1 and 227.5 ± 21.6 (Fig. 1c), respectively, which were close to those in the bacterial cells. It had been reported that the polyP in TB-EPS presented the transformation characteristic of anaerobic-decomposition/ aerobic-synthesis [11,12]. Therefore, the polyP in TB-EPS should be in situ synthesized by extracellular polyP kinases that came from the secretion or lysis of PAOs [10]. On the other hand, the average polyP chain lengths in the LB-EPS of the above sludges were 3.1±1.2, 4.0±1.6 and 2.2±0.7 in turns. It was inferred that the extracellular polyphosphate kinases should not be evenly dispersed throughout the EPS layer but located around the cell-membrane surface of PAOs. The long chain-length extracellular polyP could not transfer through inner TB-EPS into outer LB-EPS, whilst the short chain-length polyP could transfer into the outer LB-EPS layer. Compared with the SRTs of 12 d and 18 d, the more polyphosphate kinase might be secreted or lysis leached into the TB-EPS at the SRT of 25 d due to the obvious enhancement of endogenous metabolism, leading to the in-situ synthesis promotion of long chain-length extracellular polyP.
3.5.3. Influence of SRT on the extracellular P adsorption and its implications
As shown in Fig. 1 and Fig. 2, the TB-EPS was the main component of EPS, and the extracellular P and polyP were mainly distributed in TB-EPS. The TB-EPS yield, and the polyP content and average chain length in TB-EPS were significantly increased at 20±1°C by increasing the SRT from 18 d to 25 d, because of the endogenous-metabolism enhancement. The growth state of microorganisms influences the EPS yield, in which the microorganisms under endogenous metabolic conditions can produce more EPS [13]. Thus, increasing SRT can promote the production of EPS. In the study, the extracted TB-EPS content by CER was increased by increasing the SRT from 12 d to 25 d (Fig. 2), while both the residual TB-EPS and metal cations (e.g., Ca2+) amounts were also increased (Fig. 4). Moreover, the much stronger hydrophobic interaction and denser structure in TB-EPS at the longer SRT (25 d) could increase the resistance of cation exchange and then reduce the TB-EPS extraction. In brief, both the extracted and residual TB-EPS amounts were increased by increasing SRT, indicating that the TB-EPS yield were promoted by the enhancement of endogenous metabolism that was related with the decrease of metabolism activity and sludge loading. In this study, LB-EPS production was decreased by increasing SRT, suggesting that LB-EPS might be mainly derived from the secretion linked with the substrate utilization. In the previously reports, the conclusions about the effect of SRT on EPS yield were inconsistent and even contradictory [38–40], which was not only related to the differences in environmental factors and operating parameters, such as carbon source type, influent water quality, temperature, pH and hydraulic residence time, etc., but also directly connected with the differences in EPS extraction methods and procedures. Comparatively, the ultrasonic-CER method is a more efficient and reliable method for the EPS extraction as well as the LB-EPS and TB-EPS fractional extraction.
Zhang et al. reported that the EPS TP content in BPR-MBRs at 25±2°C was increased through increasing the SRT from 20 d to 40 d [5], but it was relatively stable when continuously increasing the SRT from 40 d to 50 d, which was corresponded to obviously reduced BPR efficiency. Thereby, increasing extracellular P content should be a feasible way to kept good BPR efficiency at longer SRTs. On the other hand, the proliferation of GAOs (e.g., Ca. Competibacter) was significantly promoted and the loss risk of PAOs (e.g., Ca. Accumulibacter) dominance was also increased through the enhancement of endogenous metabolism, which would lead to the loss of microbial population base for synthesizing and storing intracellular and extracellular polyP, resulting in the rapid decrease of sludge TP content and BRP efficiency. Therefore, maintaining the dominance of PAOs over GAOs at longer SRTs is a focus and difficulty in the design and operation of simultaneous nitrogen and phosphorus removal system and BRP-MBR. It should be an important strategy of adopting lower dissolved oxygen (DO) (e.g., 0.5–1.5 mg/L) and higher pH (e.g., 7.5–8.5) based on the optimized process configuration, to maintain the dominance of PAOs over GAOs at longer SRTs [1,20,21], in which the lower DO could effectively reduce the endogenous metabolism level at longer SRTs and alleviate the rapid proliferation of GAOs, and the higher pH was conducive to increase PAOs/GAOs ratio.
Synchronous chemical P removal (SCPR) is a simply and economical technology to improve the P removal steady. The practical SCPR dosage (i.e., 1.5–4.0 times) was usually much greater than the theoretical dosage [41,42], which could be ascribed to that the electronegative groups in EPS was bound with metal cations and their hydrolyzed products. Based on the results in this study, the implications could be acquired as follows: i) the metal elements adsorption properties of the inner layer TB-EPS were significantly reduced at longer SRTs (e.g., 25 d), implying that the SCPR dosage could be significantly reduced in a BPR system, and ii) the phosphates were more preferentially prone to be bound with metal cations than the electronegative groups of EPS, and the main species of phosphate in the TB-EPS of BPR sludge at the end of aerobic stage was polyP, thus the addition of Mg salt should be conducive to the extracellular active polyP formation, and iii) the main species of phosphate was orthoP in LB-EPS, in which the content of Ca was more than those of K and Mg, thereby the appropriate addition of Ca salt could also improve the P removal, besides the improvement of sludge floc stability [13]. It had been reported that aquatic othroP was completely removed in an anaerobic-anoxic BPR system with the Mg2+/Ca2+ molar ratio of 3.8, while the P removal performance was deteriorated with the molar ratio of 5.0 [43]. Therefore, the addition of cheap Mg salt (e.g., MgCl2) and Ca salt (e.g., CaCl2) as well as the optimization of Mg/Ca ratio would be a potentially effective and economical way to improve the P removal efficiency of BPR system at longer SRTs, which needed to be verified in future research.
Conclusions
Although the EPS TP contents accounted for 63.7%–68.1% of sludge TP contents, extracellular P content was still underestimated due to the residual TB-EPS layer protecting cell membrane integrity. The influences of SRT on the yield and compositions of TB-EPS were obviously greater than those of LB-EPS. The P and polyphosphate (polyP) contents of TB-EPS were significantly increased through increasing the SRT from 18 d to 25 d, in which the P adsorption of TB-EPS was enhanced mainly with the TB-EPS yield increase. Compared with the SRTs of 12 d and 18 d, the significantly enhanced P storage of BPR sludge with the SRT of 25 d was related with the increases of TB-EPS yield, polyP content and its average chain length in TB-EPS. The P adsorption property of TB-EPS was slightly increased through increasing SRT, while the K, Mg and Ca adsorption properties were markedly reduced. Increasing extracellular P content could kept good BPR efficiency at longer SRTs, but the loss risk of phosphorus accumulating organisms (e.g., Ca. Accumulibacter) dominance was increased. Additionally, the addition of Mg and Ca salts might improve extracellular P adsorption.
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