Introduction

The excessive discharge of nitrogen and phosphorus is the main reason leading to water eutrophication. Quantities of wastewater treatment plants (WWTPs) had been established to prevent nitrogen and phosphorus from entering aquatic water systems (Zhang et al. 2016). For these two nutrients, reducing the phosphorus input into lakes possibly played a more important role in the eutrophication control (Schindler et al. 2008).

As an economical process for phosphorus removal, an enhanced biological phosphorus removal (EBPR) system had been widely applied in WWTPs. However, the performance of EBPR was vulnerable to the insufficient carbon source (Oehmen et al. 2007), high temperature (Panswad et al. 2003), and other adverse environmental factors (Nielsen et al. 2019). Thus, the simultaneous chemical phosphorus removal (SCPR) process was increasingly adopted in WWTPs. In the SCPR system, phosphorus removal reagent was dosed at the end of the aerobic zone, which could coprecipitate with the residual phosphorus and further decrease the phosphorus concentration in effluent. However, the chemical reagent could flow back into the biological zone along with the sludge reflux, which would reversely affect the operational characteristics and the bacterial community in the system, especially for those related to phosphorus removal (Auger et al. 2013; Zheng et al. 2018).

Polyphosphate-accumulating organisms (PAOs) and glycogen-accumulating organisms (GAOs) were the two common bacterial groups related to the biological phosphorus removal (BPR). For PAOs, Candidatus Accumulibacter (hereafter referred to as Accumulibacter) was considered to be the most significant genus which had been extensively examined (Li et al. 2019; Nielsen et al. 2019; Zhang et al. 2018). Accumulibacter could take up volatile fatty acids (VFAs) to synthesize poly-β-hydroxyalkanoates (PHAs) in the anaerobic zone accompanied by the phosphorus release and excessively absorb phosphorus in the aerobic zone using the energy from the PHA decomposition (Oehmen et al. 2007). In recent years, Tetrasphaera was also found to be an important PAO group, which widely existed in WWTPs around the world (Muszynski and Zaleska-Radziwill 2015; Nielsen et al. 2019; Stokholm-Bjerregaard et al. 2017). Tetrasphaera could obtain energy from fermenting glucose and proteins (Liu et al. 2019). Several other putative PAOs were also identified, such as Comamonadaceae (Ge et al. 2015), Dechloromonas (Gunther et al. 2009; Terashima et al. 2016), Thiothrix (Rey-Martínez et al. 2019), and Microlunatus (Kawakoshi et al. 2012; Zhong et al. 2018). GAOs always coexisted with PAOs in the WWTPs. The common GAOs included Candidatus Competibacter (hereafter referred to as Competibacter) and Defluviicoccus (Coats et al. 2017; Xia et al. 2018). It is generally accepted that PAOs decomposed polyphosphate to obtain energy, while GAOs needed to decompose glycogen. Polyphosphate hydrolysis was much faster than glycogen decomposition; thus, PAOs took up VFAs more quickly and possessed a competitive advantage. But the PAO abundance declined under some adverse conditions, and then, the abundance of GAOs increased (Tu and Schuler 2013). PAOs could survive by changing metabolic pathways from polyphosphate accumulating metabolism (PAM) to glycogen accumulation metabolism (GAM) under adverse conditions according to Acevedo et al. (2012, 2017). In the SCPR system, the addition of phosphorus removal reagent reduced the amount of phosphorus available to PAOs, which was disadvantageous to PAOs. At present, there are few studies on the community structures of PAOs and GAOs in the SCPR system.

In this study, the phosphorus removal performance in a full-scale SCPR system operated for a long term was investigated. And the bacterial community structures, particularly the populations of PAOs and GAOs, were analyzed using quantitative polymerase chain reaction (qPCR) and high-throughput sequencing.

Materials and methods

Description of the SCPR system

The modified anaerobic-anoxic-aerobic (A2O) process was applied in the SCPR system (Fig. 1). In this system, the reflux sludge mixed with 10% of the influent flowed into the preanaerobic zone to deplete the oxygen and subsequently entered into the anaerobic zone, where it mixed with the remaining 90% of the inflow. The internal recycle from the end of the aerobic zone to the anoxic zone was adopted. The ratios of sludge reflux and internal reflux were 100% and 300%, respectively. The municipal wastewater was treated in this system, without industrial wastewater. The influent flow rate was 2.0 × 105 t/day. The influent chemical oxygen demand (COD) and total phosphorus concentrations were 180–400 mg/L and 3–6 mg/L, respectively. The hydraulic retention time (HRT) and the sludge retention time (SRT) were 10 h and 16 days, respectively. Aluminum sulfate (Al2(SO4)3) was added at the end of the aerobic zone (that is the end of the O3 zone) to ensure that the total phosphorus concentrations in effluent stably met the discharge standard of the total phosphorus (0.5 mg/L). The dosage of Al2(SO4)3 was 33 g/m3, which was excessive compared with the value from the theoretical calculation. Samples were collected along the wastewater flow. The COD, ortho-phosphate, mixed liquor suspended solid (MLSS), mixed liquor volatile suspended solids (MLVSS), and some other parameters were detected. Moreover, activated sludge samples were lyophilized using the freeze dryer system (FreeZone®, Labconco Co., USA) and then stored at − 20 °C for subsequent DNA extraction.

Fig. 1
figure 1

Schematic of the full-scale SCPR system. The numbers represent the manual sampling points (1, effluent of the primary sedimentation tank; 2, effluent of the anaerobic zone; 3, effluent of the anaerobic zone after HAc addition; 4, effluent of the anoxic zone; 5, effluent of the aerobic zone 1; 6, effluent of the aerobic zone 2; 7, effluent of the aerobic zone 3; 8, effluent of the secondary sedimentation tank)

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Restorability tests for the BPR performance

Batch tests were carried out to verify the BPR performance of the activated sludge in the SCPR system. The activated sludge was collected at the end of the aerobic zone, which was upstream of the Al2(SO4)3 addition point. Three conical bottles of 1000 mL were used in the tests, and the real domestic sewage was used as the influent. Tests were operated under alternating anaerobic/aerobic mode for three cycles without the addition of Al2(SO4)3. The anaerobic phase was 1.5 h, and the aerobic one was 3 h. The mixed liquid was sampled regularly during the cycles, and the VFAs and phosphorus concentrations were determined.

Analytical methods

MLVSS, MLSS, sludge volume index (SVI), phosphorus, and COD were measured as described by the APHA standard methods (2005). VFAs and PHAs were analyzed by gas chromatography (Agilent 6890A, USA) according to Zeng et al. (2016). A PerkinElmer fluorescence spectrometer (LS55, USA) was adopted to analyze the chemical compositions in the aquatic water from the SCPR system, and the results were presented and analyzed as described by Chen et al. (2003). The Standards of Measurements and Testing (SMT) protocol was used to analyze the phosphorus fractions in the activated sludge (Pardo et al. 2004). According to the SMT protocol, the total phosphorus was fractionated into organic phosphorus (OP) and inorganic phosphorus (IP), and IP was further fractionated into apatite phosphorus (AP) and non-apatite inorganic phosphorus (NAIP).

Quantitative PCR

Genomic DNA of the activated sludge in the SCPR system was extracted using the FastDNA™ SPIN Kit for Soil (MP Biomedicals, USA). The 16S rRNA genes of bacterial and Accumulibacter were quantified by Mx3005P (Agilent Technologies, USA). The ppk1 genes of Accumulibacter (clades IA, IIA, IIB, IIC, IID, and IIF) were quantified to reveal its clade-level structures. The programs and primers referred to the previous research (Fan et al. 2019).

High-throughput sequencing

The high-throughput sequencing of the genomic DNA was conducted by the Illumina MiSeq PE3000 platform in Majorbio Co., Ltd. (Shanghai, China). The 338F and 806R primers were used. The sequences were quality-filtered and analyzed according to Guo et al. (2018). The raw sequences had been submitted to Sequence Read Archive (SRA) at NCBI with PRJNA595869.

Results and discussions

Characteristics of activated sludge

The MLSS and MLVSS maintained at 4500 ± 235 mg/L and 2900 ± 190 mg/L, respectively. The ratio of MLVSS/MLSS was about 64%, which indicated a considerable proportion of inorganic components in the activated sludge. Inorganic components are favorable to improve sludge compactness and settleability, which signified the low SVI of 83 mL/g MLSS. However, there are some disadvantages with the low MLVSS/MLSS, such as the high sludge yield and low sludge fermentation efficiency.

Al2(SO4)3 altered the phosphorus fractions in the activated sludge through coprecipitating with phosphorus. Thus, the phosphorus fractions in the sewage sludge were analyzed using the SMT protocol, and the results are shown in Table 1. Based on the results, the TP content reached up to 37.8 ± 1.8 g/kg MLSS. The IP fraction was the dominant constituent, accounting for 84.7 ± 1.3% of TP. Among IP, NAIP was the major fraction with a percentage of 75.8 ± 4%. AP represents phosphorus combined with calcium ions, and NAIP represents those coprecipitated with aluminum ions, iron ions, and other ions. Results indicated that the activated sludge contained higher TP content than that described by Wang et al. (2018). Meanwhile, the NAIP proportion was also much higher than that found by Wang et al. (2018), which was due to the addition of Al2(SO4)3. Moreover, the sum of OP and NAIP was as high as 86.7% of the TP, indicating the high proportion of releasable and bioavailable phosphorus, which should be considered in the subsequent sludge treatment (Ruban et al. 2001).

Table 1 Phosphorus fractions in the activated sludge
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Carbon and nutrient conversions along the wastewater flow

The SCPR system presented stable removal performance of organic matter and nutrients. The TP removal efficiencies were as high as 93 ± 3% according to long-term monitoring data. Figure 2 shows the conversions of the organic matter and ortho-phosphate concentration along the wastewater flow at 2018-09-26 and 2018-10-24. In the anaerobic zone, most of the VFAs from the influent wastewater were consumed, accompanied by the increase of PHA content in the sludge, while the phosphorus release was not observed. In the anoxic zone, the phosphorus concentration kept stable and the PHA content slightly decreased. In O1 zone, the phosphorus concentration decreased by 36% and 67%, respectively. In O2 zone, the phosphorus concentration rose back, which was probably due to the phosphorus release from sludge. After Al2(SO4)3 addition at the end of the O3 zone, the phosphorus concentration declined. The PHA content in the oxic zones further decreased. In the secondary sedimentation tank, phosphorus concentration continued to descend to 0 mg/L at the lowest.

Fig. 2
figure 2

Variations of the phosphate, VFAs, and PHAs along the wastewater flow sampled at a 2018-09-26 and b 2018-10-24 (PST, effluent of the primary sedimentation; AN1, effluent of the anaerobic zone; AN2, effluent of the anaerobic zone after HAc addition; AX, anoxic zone; O1, aerobic zone 1; O2, aerobic zone 2; O3, aerobic zone 3; EFF, effluent of the secondary sedimentation tank)

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In the EBPR system, PAOs release phosphorus into bulk liquid in the anaerobic stage and excessively absorb phosphorus in the aerobic zone (Oehmen et al. 2007). In this SCPR system, these phenotypes were not observed because Al2(SO4)3 was overdosed. The excess Al2(SO4)3 could recirculate into the anaerobic zone along with the reflux sludge and coprecipitate with the phosphate, which caused a rapid decline of the phosphorus concentration in the anaerobic zone. PAOs could not obtain enough phosphorus, leading to the decline of PAM activity. PAOs obtained energy more quickly through PAM, which ensured the preferential absorption of VFAs. Therefore, the decrease of PAM activity made it difficult for PAOs to obtain enough VFAs and accumulate sufficient polyphosphate, which decreased the proliferation rate of PAOs and finally caused the reduction of their relative abundances. Partial PAOs might survive and proliferate through altering their metabolic mode from PAM to GAM under the low content of polyphosphate (Acevedo et al. 2017; Welles et al. 2015). Besides, some groups of PAOs were possibly alive through PAM, but the ortho-phosphates released by them in the anaerobic zone were immediately precipitated with Al2(SO4)3. All these three situations weakened the phenotype of phosphorus release in the anaerobic zone and absorption in the aerobic zone. The community structures should be investigated to deepen the understanding of this SCPR system.

Organic dynamics along the wastewater flow

The organic compositions in the bulk liquid were analyzed and presented using the excitation-emission matrix (EEM) (Fig. 3). As described by Chen et al. (2003), the EEM could be divided into five regions, i.e., tyrosine-like region (I), tryptophan-like region (II), fulvic acid–like region (III), the soluble microbial by-product-like (SMP) region (IV), and humic acid–like region (V). As shown in Fig. 3a, the effluent of the primary settling tank (PST) contained abundant organics, which presented high fluorescence intensity. The excitation-emission area volume of tryptophan-like organics (ΦII) was 3.1 × 106 Au nm2, lower than those of SMPs (6.6 × 106 Au nm2) and humic acid–like organics (8.7 × 106 Au nm2), but its normalized percent of fluorescence response (PII,n) was the highest of 35% (Table 2).

Fig. 3
figure 3

Three-dimensional fluorescence spectra of soluble organics in the bulk liquid along the wastewater flow (PST, effluent of the primary sedimentation; AN1, effluent of the anaerobic zone; AN2, effluent of the anaerobic zone after HAc addition; AX, anoxic zone; O1, aerobic zone 1; O2, aerobic zone 2; O3, aerobic zone 3; EFF, effluent of the secondary sedimentation tank)

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Table 2 Excitation-emission area volume and the percent of fluorescence response following the wastewater flow
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In the anaerobic zone, the fluorescence intensity of the five regions decreased (Fig. 3b). The excitation-emission area volume of regions I–V, respectively, decreased by 75%, 63%, 43%, 45%, and 20% (Table 2), indicating the higher biodegradability of aromatic proteins (regions I and II). After the acetate addition, the fluorescence intensity was unchanged (Fig. 3c). In the anoxic zone, the fluorescence intensity of these five compositions kept declining. Compared with the influent, the excitation-emission area volume decreased by 95%, 87%, 75%, 54%, and 20%, respectively (Table 2). In the O1 zone, the fluorescence intensity further decreased (Fig. 3e). The excitation-emission area volume of regions I–V decreased by 98%, 94%, 85%, 64%, and 30%, respectively. However, the fluorescence intensity and the normalized percent of fluorescence response kept stable in all the three aerobic zones (Fig. 3e–g), suggesting that aeration had a weak effect on the removal of fluorescent organics. In the effluent, SMPs and humic acid–like compositions became the major fluorescent substances with their normalized fluorescence response of 29% and 38%, respectively (Fig. 3h, Table 2). EEM results showed that there were diverse organic matters in the wastewater, which provided the material basis for microbial diversity.

Bacterial community structure in the SCPR system

The community structure in the SCPR system was investigated using high-throughput sequencing. The bacterial population exhibited high richness and diversity, with the Chao1 and Shannon indexes as high as 1498 and 5.46. In the phylum level, Proteobacteria was the dominant phylum with a relative abundance of 52%, and the Gammaproteobacteria (18.6%), Betaproteobacteria (15.4%), and Alphaproteobacteria (11.1%) were the major classes which affiliated to Proteobacteria. Chloroflexi, Bacteroidetes, Actinobacteria, and Nitrospirae were also the major bacterial phyla in this SCPR system, and the relative abundances were 13.7%, 9.51%, 8.45%, and 4.2%, respectively. The relative abundances of Firmicutes, Acidobacteria, Planctomycetes, and Saccharibacteria were all below 3%.

In this study, the community structures of PAOs and GAOs were the focuses. The putative populations of PAOs are shown in Fig. 4a, including Ca. Accumulibacter, Tetrasphaera, Dechloromonas, Microlunatus, Tessaracoccus, Thiothrix, and Comamonadaceae. In this study, the relative abundance of Accumulibacter was only 0.59%. Five Accumulibacter OTUs were observed, among which OTU1816 (41.8%) and OTU1602 (37.8%) were dominant. The relative abundance of Accumulibacter was also quantified using qPCR (Fig. 5a). The copy numbers of the Accumulibacter 16S rRNA gene and bacterial 16S rRNA gene were 2.40 × 1010 and 3.24 × 1012 copies/g sludge, respectively. According to Harms et al. (2003) and He et al. (2007), the average copy numbers of bacterial and Accumulibacter 16S rRNA gene were 3.6 and 2. Therefore, the relative abundance of Accumulibacter was 1.33% according to qPCR data, slightly higher than the ratio got by high-throughput sequencing. The clade-level population of Accumulibacter was investigated using ppk1 genes (Fig. 5b). Based on the qPCR results, clades IIC and IID were the dominant clades in this SCPR system with ratios of 39.3% and 37.6%, respectively. Different Accumulibacter clades had different metabolic characteristics. Clade IIC was observed as a dominant clade in diverse wastewater treatment systems (He et al. 2007; Mao et al. 2015; Mielczarek et al. 2013; Muszynski et al. 2018; Ong et al. 2014; Qiu et al. 2019). Particularly, the dominant clades in this SCPR system were consistent with those in 18 WWTPs from six countries (Mao et al. 2015). These suggested that the fine-scale population of Accumulibacter in this SCPR system was not affected by the addition of Al2(SO4)3 although the relative abundance of Accumulibacter was low. Results both from high-throughput sequencing and qPCR indicated that Accumulibacter exhibited relatively lower abundance in the SCPR system than that observed in EBPR plants (Stokholm-Bjerregaard et al. 2017).

Fig. 4
figure 4

The populations and relative abundance of a PAOs and b GAOs in the SCPR system

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Fig. 5
figure 5

The abundances of 16S rRNA genes of bacteria and Accumulibacter in the SCPR system

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Tetrasphaera was another significant genus of PAOs (Fernando et al. 2019; Liu et al. 2019). In this study, the relative abundance of Tetrasphaera was only 0.20%, and only two OTUs were observed. Tetrasphaera was a fermentative bacterial group, which obtained energy from the fermentation of proteins and polysaccharides. In a WWTP in Denmark, the relative abundance of Tetrasphaera reached up to 21.2% (Stokholm-Bjerregaard et al. 2017). Similarly, in 4 Polish full-scale WWTPs, Tetrasphaera constituted 11–25% of all bacteria (Muszynski and Zaleska-Radziwill 2015). The diet structure with abundant milk and meat in Denmark and Poland possibly caused the high content of protein compositions in the sewage, which was conducive to Tetrasphaera proliferation. However, the Al2(SO4)3 was added in the SCPR system, which consumed the phosphorus source and limited the proliferation of Tetrasphaera.

Dechloromonas was affiliated to the Rhodocyclaceae family, the same as Accumulibacter (Terashima et al. 2016). In this study, the relative abundance of Dechloromonas was 1.85%, which was higher than Accumulibacter and Tetrasphaera. Dechloromonas was subdivided into four taxa with OTU1181 (43.0%) and OTU2458 (48.3%). Comamonadaceae had a high growth rate in the activated sludge ecosystems (Ge et al. 2015; Saunders et al. 2016). In this SCPR system, Comamonadaceae was divided into seven OTUs with a relative abundance of 1.92%, among which OTU2205 and OTU2208 exhibited higher proportions of 42.6% and 28.6%.

In the SCPR system, Competibacter and Defluviicoccus were the predominant genera of GAOs (Fig. 4b). The relative abundances of Competibacter and Defluviicoccus were 5.77% and 1.28%, respectively. For Competibacter, sixteen OTUs were observed, among which OTU124 was the dominant one with a ratio of 91.2%. For Defluviicoccus, five OTUs were found with OTU725 (83.2%) as the dominant one. Results showed that the relative abundances of Competibacter and Defluviicoccus were both higher than previous findings (Stokholm-Bjerregaard et al. 2017). According to Nielsen et al. (2019), Competibacter and Defluviicoccus were GAOs possessing the canonical GAM mode. In this mode, VFAs were taken up and transformed into glycogen at anaerobic conditions. At aerobic conditions, glycogen was degraded to supply the energy and substrate for their proliferation, and the remaining glycogen was transformed into PHAs for storage. In the EBPR system, PAOs used the energy from the hydrolysis of polyphosphate to preferentially absorb VFAs. In the SCPR system, Al2(SO4)3 coprecipitated with the phosphorus in the bulk liquid, which created an environment with a high ratio of carbon to phosphorus. These conditions limited the metabolism and proliferation of PAOs, whereas GAOs got more VFAs and grew better.

BPR restorability batch tests

To verify the BPR performance of the activated sludge in this SCPR system, the alternating anaerobic/aerobic operation was carried out (Fig. 6). Chemical reagent was not added in the tests. The phosphorus and VFA concentrations were regularly detected, and the ratios of phosphorus release to VFA uptake (Prel/VFAupt) were calculated. In the first cycle, the activated sludge showed slight BPR performance. The phosphorus concentration increased from 2.6 to 6.4 mg/L during the anaerobic stage and rapidly declined to 0.2 mg/L after 1 h of aerobic reaction. The Prel/VFAupt was 0.13 mmol-P/mmol-C. In the second cycle, the phosphorus concentration at the end of the anaerobic stage increased to 10.0 mg/L, and the Prel/VFAupt increased to 0.19 mmol-P/mmol-C. In the third cycle, the maximum phosphorus concentration and Prel/VFAupt increased to 12.2 mg/L and 0.20 mmol-P/mmol-C, respectively. The Prel/VFAupt had been used as an indicator of the PAM activity, which was generally higher than 0.5 mmol-P/mmol-C in EBPR systems (Acevedo et al. 2017; Oehmen et al. 2007; Schuler and Jenkins 2003). Results showed that although no typical characteristics of PAM were observed in the SCPR system, the PAM activity of the activated sludge gradually recovered after the Al2(SO4)3 addition was stopped, which corresponded to the existence of diverse PAOs.

Fig. 6
figure 6

The variations of phosphorus and VFA concentration during the anaerobic/aerobic cycles after stopping dosing Al2(SO4)3

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The addition of chemical reagents in the SCPR system increased the operation cost of sewage treatment. This study demonstrated that PAOs still existed in the SCPR system and could recover the PAM activity after stopping the addition of the chemicals. Therefore, it is possible to enhance the BPR performance by gradually reducing or even stopping the addition of chemical reagents in the SCPR systems, so as to reduce the operating costs.

Conclusion

In the long-term SCPR system with Al2(SO4)3 addition, the PAM phenotype disappeared. PAOs were still present in the system with relatively low abundances, while GAOs of Competibacter and Defluviicoccus possessed relatively high abundances. However, the PAOs in the system showed high diversity and maintained the PAM ability. The activated sludge could recover the BPR performance after the Al2(SO4)3 addition was stopped. These results proved the potential of the SCPR system to resume EBPR performance.