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Article

Enhanced Simultaneous Nitrogen and Phosphorus Removal in a Continuous-Flow Granular Sludge System under Gradient-Controlled Hydraulic Loading

by
Yaguang Zhao
1,
Pengkun An
1,
Junfeng Wan
2,3,* and
Xuehui Zhang
1
1
School of Environmental and Municipal Engineering, North China University of Water Resources and Electric Power, Zhengzhou 450046, China
2
School of Ecology and Environment, Zhengzhou University, Zhengzhou 450001, China
3
Henan International Joint Laboratory of Environment and Resources, Zhengzhou University, Zhengzhou 450001, China
*
Author to whom correspondence should be addressed.
Water 2024, 16(11), 1510; https://doi.org/10.3390/w16111510
Submission received: 20 April 2024 / Revised: 21 May 2024 / Accepted: 22 May 2024 / Published: 24 May 2024
(This article belongs to the Section Wastewater Treatment and Reuse)
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Abstract

:
The feasibility of the aerobic granulation of activated sludge was investigated in a continuous-flow anaerobic–anoxic–oxic system under gradient-controlled hydraulic loading on the surface of a cyclone separator. Concentrated domestic sewage was used. After 80 days of operation, 80% of activated sludge in the system was in the form of granular sludge with an average particle size of 373 μm. High removal efficiency was achieved for chemical oxygen demand (94.40%), NH4+-N (99.93%), total nitrogen (89.44%), and total phosphorus (96.92%). A batch study revealed that Pseudomonas (1.34%) and Dechloromonas (1.05%) as the main denitrifying phosphorus-accumulating organisms could efficiently remove phosphorus using nitrate as an electron acceptor, which improved the utilization efficiency of carbon sources and achieved simultaneous denitrification and phosphorus removal. Overall, the study demonstrates the feasibility of enhanced denitrification and phosphorus removal in a continuous-flow granular sludge system. The sludge system enables simultaneous nitrogen and phosphorus removal under low carbon-to-nitrogen ratios.

1. Introduction

Aerobic granular sludge (AGS) is activated sludge formed through the spontaneous aggregation and proliferation of microorganisms in specific environments. It is characterized by a compact structure, excellent settling performance, and the capacity to withstand shock and high loading rates [1]. The AGS formation process is complex and involves physical, chemical, and biological processes [2]. Various hypotheses of the AGS formation mechanism have been proposed, including the nucleation hypothesis, selective pressure-driven hypothesis, filamentous bacteria hypothesis, and stage formation hypothesis [3,4,5]. However, “selective pressure” is widely recognized as the most effective influencing factor [6]. Wang et al. [7] indicated that in the sedimentation tank of a continuous-flow reactor, the “selective pressure” based on the sludge settling rate was positively correlated with hydraulic retention time (HRT). When the settling velocity of granular sludge exceeds the upward flow velocity of water, the granular sludge will settle and be retained in the reactor. Higher “selective pressure” requires an increase in the hydraulic load in the settling tank through shortening the HRT to achieve the sludge screening effect. However, in the AGS system, a short HRT can significantly reduce the nitrogen and phosphorus removal efficiency of the reaction system. Other studies [8,9] have shown that a constant “selective pressure” cannot adequately adapt to the varying hydraulic load requirements throughout different stages of sludge granulation. Excessive “selective pressure” in the initial stages of granulation can lead to considerable losses of activated sludge, which are challenging to recover independently. Conversely, a lower selection pressure can influence both the sludge granulation process and the particle size of granular sludge. Guo et al. [8] explored the shear and centrifugal forces provided using a hydrocyclone separator and their impact on the formation of granular sludge. The results showed that the use of a hydrocyclone separator significantly reduced the sludge volume (SV) index after 30 min (SVI30) but caused a large amount of sludge loss and an insignificant granulation effect. Therefore, separating the “selective pressure” from the reaction system HRT and providing controlled “selective pressure” according to the granulation process of activated sludge are important research topics for achieving sludge granulation in continuous-flow AGS systems.
Most of the studies on the removal mechanism of pollutants using AGS are based on the microstructural characteristics of AGS. AGS exhibits unique structural features, which allow for the maintenance of various oxygen concentrations and nutrient environments within a single particle. The specific oxygen concentration gradient within the particle provides favorable growth conditions for various microorganisms, leading to diverse metabolic activities [10]. Therefore, in recent years, AGS has become an active research area in the field of wastewater treatment, attracting widespread attention [11,12]. He et al. [13] found that in the aerobic section of the AGS system, accumulated NO3−-N was less than the removed NH4+-N, indicating simultaneous nitrification and denitrification. However, in continuous-flow AGS reactors, the process of activated sludge granulation driven by “selective pressure” is essentially a targeted screening of activated sludge. This process is coupled with the conditions of pollutants (culture medium) within the system, leading to the directional evolution of the microbial community structure within the system and thereby altering the pollutant removal mechanism. Studies on the enhanced effect of sludge granulation on the denitrifying phosphorus removal mechanism in anaerobic–anoxic–oxic (AAO) reactors are limited.
In this study, a hydrocyclone, capable of controlling “selective pressure”, was installed behind the AAO reactor to conduct the directional screening of activated sludge without affecting the HRT of the system. This study investigated the influence of selection pressure on the process of activated sludge granulation and the removal patterns of nitrogen (N) and phosphorus (P) within the reactor under gradient-controlled hydraulic loading on the cyclone separator surface. According to the results of batch experiments and biodiversity analysis, the N and P removal mechanisms were further elucidated, providing relevant foundational data for future research on the cultivation, domestication, and denitrification and phosphorus removal mechanisms of AGS in wastewater treatment.

2. Materials and Methods

2.1. Experimental Setup

An AAO continuous-flow plug-flow reactor system (Figure 1) was used in this experiment. The reaction system consisted of seven (numbered 1–7) equally sized reaction columns with an effective volume of 1.72 L each, connected end-to-end, and a hydraulic vortex at the end of column 7. The ratio of height to diameter of the aerobic column is 16:1. Water is absorbed from the settling tank using a peristaltic pump and transferred into the hydrocyclone to provide a surface hydraulic load q0 (selective pressure). By controlling the flow rate of the peristaltic pump, q0 is precisely controlled. Columns 1–3 were stirred using agitators to maintain a uniform sludge–water mixture, forming the anaerobic zone. Columns 4–7 were equipped with aeration devices, forming the aerobic zone. The nitrification liquid was recirculated from column 7 to column 2 through gas-lift action, forming internal circulation, and the sludge was recirculated to column 1 through gas-lift action, forming external circulation.

2.2. Substrate Composition and Seed Sludge

Artificially simulated wastewater was utilized to simulate the high concentration of dispersed domestic sewage, such as that found in highway service areas [14], rural sewage [15], and residential and public buildings [16]. The composition of the prepared synthetic wastewater is detailed in Table 1; volume load is 1.2 kg/(m3·d). The inoculum sludge was taken from the mixed liquor of the aeration tank in a sewage treatment plant in Zhengzhou City. The values for mixed liquor suspended solids (MLSS), SVI30, and sludge particle average size were 2.6 g/L, 185.4 mL/g, and 62 μm, respectively.

2.3. Analytical Methods

The NH4+-N, NO3−-N, chemical oxygen demand (COD), total nitrogen (TN), total phosphorus (TP), SV, and MLSS of the samples were analyzed through standard methods [17]. At the end of the aerobic period, 100 mL of the mud mixture was collected in a cylinder for static sedimentation. Measurements were taken at 5 min and 30 min after collection to determine the SV percentage, denoted as SV5 and SV30, respectively. SVI30 was calculated as the ratio of SV30 to MLSS. Total nitrogen (TN) was computed as the sum of NH4+-N, NO2−-N, and NO3−-N concentrations. The sludge morphology was observed using an optical microscope (XS-213, JNOEC, Jiangnan, China), and the AGS particle size was measured using a laser particle size analyzer (BT-Online1A, Bettersize, Liaoning, China). The extracellular polymeric substance (EPS) content of the sludge sample was extracted via heating, as described in a previous study [18]. The sludge suspension was centrifuged at 8000 rpm for 5 min in a 50 mL tube and washed three times with deionized water. After the supernatant was removed, the sludge samples were then resuspended into 15 mL of 0.05% NaCl solution. Subsequently, the sludge mixture was placed in a water bath at 80 °C for 30 min, followed by centrifugation at 8000 rpm for 15 min. Finally, the supernatant filtered through a 0.22 mm cellulose membrane was collected to analyze the EPS content. The protein (PN) content of the EPSs was determined through the Lowry method [19] and the polysaccharide (PS) content was determined through the phenol–sulfuric acid method [20].

2.4. Microbial Community Analysis

The seed sludge (day 1) and AGS (day 80) samples were taken from the end of the aerobic section before the hydrocyclone to analyze the microbial community succession during AGS formation. DNA extraction, polymerase chain reaction amplification, and pyrosequencing were conducted sequentially using the general primers 338 F (5′- ACTCCTACGGGAGGCAGCAG-3′) and 806 R (5′- GGACTACHVGGGTWTCTAAT-3′) for bacterial diversity analysis according to previously established procedures [21].

2.5. Calculation of Pollutant Removal Efficiency

2.5.1. Continuous Operation

After the inoculation of the sludge into the reaction apparatus, as previously mentioned, the reaction was initiated under controlled conditions with an HRT of 12 h, a nitrification liquid recirculation ratio of 300%, and a sludge recirculation ratio of 100%. The influent flow rate, nitrification liquid recirculation flow rate, and sludge recirculation flow rate were 1000 mL/h, 3000 mL/h, and 1000 mL/h, respectively. Temperature indicates the ambient temperature. The pH was between 6.5 and 8.5. As shown in Table 2, the experimental period was divided into three phases (Phases I, II, and III) according to the hydraulic load.

2.5.2. Batch Experiments

To further analyze the denitrifying phosphorus removal mechanism, batch experiments were conducted using sludge samples from the reactor on day 1 (seed sludge) and day 80 (AGS). A total of 500 mL of sludge samples are taken from the end of the aerobic section of the system. These samples were washed three times with deionized water to remove N, P, and COD from the mixed sludge. After sufficient COD was added under anaerobic conditions and complete phosphorus release was confirmed through testing, the sludge was washed again three times with deionized water. The two sludge types were then separately introduced into seven reactors, with equal amounts of sludge in each reactor. Subsequently, batch experiments were conducted, and divided into three groups: (1) After N2 stripping, only PO43−-P at a concentration of 30 mg/L was added to the reactors. (2) In the reactors, 30 mg/L of PO43−-P was added along with different concentrations of NO3−-N (10, 20, 30, 40, and 50 mg/L). (3) Under continuous O2 supply, only PO43−-P at a concentration of 30 mg/L was added to the reactors. All batch experiments lasted for 150 min, with water samples collected every 30 min to measure the concentrations of NO3−-N and TP.

3. Results and Discussion

3.1. Effect of Hydraulic Load on Sludge Granulation

After the inoculation of the flocculent sludge with an average particle size of 62.91 μm into the reactor, and following the previously mentioned experimental method of gradient-controlled hydraulic loading (q0) on the cyclone separator’s surface, the system was continuously operated for 80 days. The average particle size of the granular sludge within the system reached 373.3 μm, with particles larger than 0.2 mm accounting for 80% of the total granular sludge (Figure 2a). The activated sludge within the system successfully achieved granulation.
During the 80-day operation of the reactor, as q0 increased, various physical indicators of the activated sludge within the reactor system (MLSS, SVI30, SV5, and SV30) exhibited corresponding patterns of change (Figure 2b). The entire sludge granulation process can be divided into three phases for analysis: Phase I (1–40 days), Phase II (41–65 days), and Phase III (66–80 days). In Phase I, as q0 increased from 2.04 m3/(m2·h) to 3.99 m3/(m2·h), the overall concentration of pollutants in the reaction system increased. Correspondingly, MLSS increased from 4.90 g/L to 5.70 g/L. Meanwhile, SVI30, SV30, and SV5 decreased from initial values of 182.0 mL/g, 85%, and 95% to final values of 109.7 mL/g, 62%, and 91%, respectively. By day 40, the average particle size of the sludge within the system had reached 93.39 μm (Figure 2a), suggesting that while the hydraulic screening effect of the cyclone separator’s surface hydraulic load improved the settling performance of the activated sludge within the system, the granulation effect on the sludge was not significant [22]. In Phase II, q0 increased from 3.99 m3/(m2·h) m/h to 5.00 m3/(m2·h) m/h until day 50, and during this period, MLSS consistently decreased from 5.70 g/L to 2.1 g/L. Nearly 60% of the sludge was washed out of the reaction system. Additionally, SVI30, SV30, and SV5 exhibited a reduction from initial values of 109.7 mL/g, 62%, and 91% to final values of 51.5 mL/g, 10%, and 10%, respectively. By day 60, a large amount of light-yellow AGS appeared in the reaction system, and the average particle size of the sludge reached 291.8 μm (Figure 2a). By day 65, SV5/SV30 decreased from an initial value of 1.49 to nearly 1, further indicating that the activated sludge within the system gradually transformed from flocculent sludge to granular sludge with better settling performance [22,23], achieving sludge granulation. In Phase III, while q0 is maintained at 5.00 m3/(m2·h), MLSS within the system gradually increased, stabilizing at around 3.5 g/L. During this phase, SV5/SV30 slightly increased but remained close to 1. By day 80, the average particle size of the granular sludge within the system reached 373.3 μm, with particles larger than 0.2 mm accounting for 80% (Figure 2a). Granular sludge dominated the activated sludge within the system, successfully achieving granulation and stable operation, thereby promoting the proliferation of granular sludge.
The changes in sludge indicators at various stages with varying q0 demonstrated that the “selective pressure” for sludge settling played a dominant role in sludge granulation within the reaction system. The sludge granulation process was characterized as “screening–growth–rescreening–regrowth”. In the initial stages of granulation, lower “selective pressure” resulted in a lower screening rate for the sludge. It improved the settling performance of the system’s sludge without causing significant sludge loss, providing the material basis for further increasing the hydraulic screening load. However, to achieve sludge granulation within the system, sufficient “selective pressure” must be provided. Therefore, controlled “selective pressure” is a key factor for this reactor to achieve sludge granulation under continuous flow conditions. Guo et al. [8] utilized a hydraulic cyclone separator to separate activated sludge in a reactor through a constant-flow-rate (constant “selective pressure”) separation method. They found that although the cyclone separator significantly improved the settling performance of activated sludge, its effect on sludge granulation was not significant, and higher flow rates led to significant sludge loss that was difficult to recover. Liu et al. [9] utilized a “dual-zone settling tank” to screen the sludge settling process, adjusting the hydraulic screening load by varying the height of the baffle set above the first settling tank. They cultured granular sludge with an average particle size of 210 μm over 115 days. However, they found it challenging to increase the hydraulic screening load in the later stages of device operation solely by adjusting the baffle height. This difficulty affected the granulation process of activated sludge and the particle size of granular sludge. Furthermore, many researchers have used the sequential batch reactors (SBRs) for AGS cultivation and achieved satisfactory granulation results [9,24,25]. They mostly achieved sludge granulation through adjusting the sludge settling time, which is a form of “selective pressure” control. The experimental setup used in the present study allowed for the gradient-controlled “selective pressure” during sludge settling based on the characteristics of MLSS and its settling performance at different sludge granulation stages. The activated sludge was subjected to hydraulic screening under a relatively sufficient microbial population within the system. This approach resulted in the granulation of activated sludge under continuous flow conditions. However, the average particle size of the AGS in the system was smaller than that of the other AGS cultured in the SBRs. Studies have shown that in addition to selection pressure, feast/famine conditions are key factors in achieving granulation in a continuous flow [26]. The absence of feast/famine conditions in the reactor might be the limiting factor preventing further increase in particle size.
The results of the analysis of the changes in the EPS content of the sludge within the reactor on day 1, day 40, day 65, and day 80, as well as the excess sludge on day 80, are presented in Figure 2c. The concentrations of PN and PS in sludge EPS gradually increased from 54.34 and 45.05 mg/g·SS on day 1 to 147.02 and 73.05 mg/g·SS on day 80. An increase in the EPS content is advantageous for the mutual aggregation of microorganisms [27]. Additionally, calculations indicated that the PN-to-PS ratio increased from 1.2 on day 1 to 2.0 on day 80. The increase in the relative content of PN enhances the hydrophobicity and settling performance of activated sludge, making it more conducive to the formation and stability of AGS [28]. Furthermore, the PN content, PS content, and PN/PS ratio in the residual sludge washed out by the hydraulic cyclone separator on day 80 were 85.89 mg/g·SS, 61.89 mg/g·SS, and 1.4, respectively, significantly lower than the corresponding parameters in the activated sludge within the system during that period. This indicates that hydraulic screening played a role in the targeted selection of activated sludge within the system. This experimental phenomenon further illustrates that the “selective pressure” provided using the hydraulic cyclone separator was conducive to retaining sludge with a higher EPS content and better settling performance within the system. The selection pressure also had a directional screening effect on the granulation of activated sludge within the system, consistent with the observed patterns in the sludge granulation process. The changes in the apparent morphology of activated sludge are illustrated in Figure 3.

3.2. Removal Patterns of N and P in the Reaction System

In the AAO reactor, the nitrification of NH4+-N serves as the foundation for denitrification and has a significant impact on the denitrification efficiency of the system [29]. During the reactor operation from day 1 to day 65, the average removal efficiency of NH4+-N remained around 97%. However, in the later stages of reactor operation (after 65 days), the removal efficiency notably increased, reaching 99.80% (Figure 4a), possibly because as the sludge granulated, nitrifying bacteria with longer generation times could attach to well-settling AGS for growth, reproduction, and gradual enrichment within the reactor. This effectively increased the biomass of nitrifying bacteria in the activated sludge, thereby enhancing the removal efficiency of NH4+-N [9]. The removal pattern of NH4+-N along the reactor (Figure 4b) showed that NH4+-N removal primarily occurred in the aerobic phase, with removal efficiencies ranging from 60% to 85%. However, NH4+-N removal also occurred in the anaerobic and anoxic phases, with removal efficiencies of 10.3% and 16.4%, respectively. Considering that a part of the dissolved oxygen was carried via the nitrification liquid and sludge backflow, a part of the NH4+-N was also removed in the anaerobic and anoxic sections. Li et al. [30] investigated large urban wastewater treatment plants and discovered the presence of anaerobic ammonia oxidation as a denitrification method in the system. They found that anaerobic ammonia oxidation accounted for 32.4% of the nitrogen loss, with anaerobic ammonia-oxidizing bacteria exhibiting an abundance of 0.11%. This indicates the presence of anaerobic ammonia-oxidizing bacteria in urban wastewater treatment plants. Accordingly, the loss of NH4+-N in the anaerobic and anoxic phases of the reactor is attributable to anammox processes.
The denitrification performance of the reaction system gradually strengthened with the operation of the reactor and the sludge granulation process. The average TN removal efficiency increased from 66.99% in the first stage to 75.55% in the second stage and 89.41% in the third stage (Figure 5a). According to the removal pattern of TN along the reactor (Figure 5b), the improvement in denitrification performance was primarily due to the increase in TN removal in the anoxic zone of the reaction system. TN removal in the anoxic zone gradually increased from 18.12% at the beginning to 59.58% at the end of the experiment. Alongside the removal of TP in the anoxic zone, a portion of NO3−-N in the anoxic zone was removed through denitrifying phosphorus removal, which conserved the carbon source that would otherwise be utilized via traditional denitrification for this portion of NO3−-N. This coordination of denitrification and phosphorus removal actions, competing for carbon sources, rendered the carbon source in the anoxic zone more abundant, further enhancing denitrification in the anoxic zone. Additionally, in the later stages of reactor operation (days 60–80), ~10% TN loss occurred in the aerobic zone. This further demonstrates that with the granulation of sludge, short-cut nitrification and denitrification and simultaneous nitrification–denitrification processes within the microenvironment of granular sludge contributed to the denitrification performance of the reactor, according to the microstructure of the granular sludge [31]. The decrease in the TN removal rate in the anaerobic zone was mainly due to the reduction in NO3−-N carried to the anaerobic zone via the sludge’s return, as the overall denitrification efficiency of the reaction system improved.
The removal efficiency of TP in the reactor remained relatively stable over 80 days, with a total removal efficiency of >97% (Figure 6a). However, the removal characteristics of TP along the different functional units within the reactor (Figure 6b) provide insights into the P removal mechanism of the reaction system. As depicted in Figure 6b, the significant removal of TP in the anoxic zone deviated from traditional biological phosphorus removal theory [32]. Traditional biological nitrogen and phosphorus removal theories suggest the existence of a tradeoff relationship between denitrification and phosphorus removal owing to competition for carbon sources. This often entails sacrificing the removal efficiency of one to enhance the removal efficiency of the other. Carvalhoa et al. [33] improved denitrifying phosphorus removal through gradually shifting the anaerobic-aerobic environment to an anaerobic–anoxic setting in an SBR. They also supplemented NaNO3 solution during anoxic conditions to increase nitrate concentration and retention time in the anoxic zone. After 87 days of operation, the system exhibited enhanced denitrifying phosphorus removal, with the removal rate increasing from 0.10 to 0.63 mmol/g VSS/h. Chen et al. [34] improved TP removal in the anoxic zone (from 1.29 to 1.69 g TP/g VSS/d) by increasing the recycle ratio of nitrified liquid in an AAO system (from 100% to 400%). The denitrifying phosphorus removal mechanism allows for the synergistic treatment of biological denitrification and phosphorus removal, which can coordinate the competition for carbon sources and reduce the dependence of biological denitrification and phosphorus removal on carbon sources, facilitating simultaneous nitrogen and phosphorus removal under low carbon-to-nitrogen ratios. The transition from traditional biological phosphorus removal to denitrification–phosphorus removal mechanisms, along with the enrichment of denitrifying phosphorus-accumulating organisms (DPAOs), can be achieved through the introduction of nitrate nitrogen stress within the anoxic zone. The substantial removal of TP in the anoxic zone in this experiment was due to the provision of an electron acceptor for DPAOs by nitrate nitrogen, enabling the proliferation of DPAOs. Conversely, in a previous study, the activity of traditional phosphorus-accumulating organisms (PAOs) was inhibited owing to insufficient phosphorus sources for their energy requirements in the aerobic zone [35]. Additionally, according to the TP removal pattern along the reactor, a negative correlation existed between the TP removal efficiency in the anoxic and aerobic zones. The TP removal efficiency in the anoxic zone from day 1 to day 40 was above 75%, but after day 40, the removal efficiency gradually decreased. Considering the characteristics of TN changes along the reactor (Figure 5b), as the denitrification efficiency of the system increased, the decrease in NO3−-N in the return nitrified liquid, which acted as an electron acceptor for DPAOs (decreasing from 24.85 mg/L on day 1 to 15.60 mg/L on day 40), resulted in a decline in TP removal efficiency in the anoxic zone. Unremoved phosphate in the anoxic zone was carried into the aerobic zone, where it underwent further removal under aerobic conditions. This process enhanced the TP removal efficiency in the aerobic zone, marking a “transition” of TP removal from the anoxic zone to the aerobic zone. One study found that DPAOs in wastewater treatment systems can use NOx−-N as an electron acceptor for denitrification-polyphosphate accumulation under anaerobic conditions, while still employing O2 as an electron acceptor for phosphorus removal under aerobic conditions [21]. According to the study’s findings, the observed aerobic polyphosphate accumulation might be facilitated by DPAOs utilizing molecular oxygen as the electron acceptor. Nevertheless, there remains the possibility of aerobic polyphosphate accumulation by PAOs in this context.

3.3. Mechanism of Nitrogen and Phosphorus Removal

To further analyze the denitrifying phosphorus removal mechanism, batch tests were conducted on the sludge collected on day 1 (seed sludge) and day 80 (AGS) within the reactor, following the experimental methods mentioned earlier. As depicted in Figure 7, after nitrogen gas stripping (anaerobic conditions), the simultaneous removal of TP and NO3−-N occurred in the test groups where both PO43− and NO3− were added. In contrast, the control group, which received only PO43−, did not exhibit significant TP removal throughout the entire reaction process. In the seed sludge, under the two abovementioned addition scenarios, no significant removal of TP and NO3−-N occurred in any of the test groups. Under continuous aeration (aerobic conditions), when only PO43−-P was added, both the granular sludge and seeded sludge test groups exhibited varying degrees of TP removal. The test results demonstrate that granular sludge could utilize NO3− as an electron acceptor for TP removal in the absence of molecular oxygen, indicating that TP removal was achieved through the denitrifying phosphorus removal mechanism. In contrast, the seeded sludge utilized molecular oxygen as an electron acceptor for polyphosphate formation under aerobic conditions but could not utilize NO3− as an electron acceptor for polyphosphate formation in the absence of molecular oxygen, indicating that the TP removal mechanism in the seeded sludge followed the traditional biological phosphorus removal mechanism. This experimental phenomenon was consistent with the significant removal of TP in the anaerobic section of the reaction system and the synergistic treatment of biological denitrification and phosphorus removal through the denitrifying phosphorus removal mechanism. This coordination between denitrification and phosphorus removal processes in the competition for carbon sources further enhanced the denitrification effect in the anaerobic section.
Moreover, the test results illustrated in Figure 7a,b demonstrate a correlation between the removal efficiencies of TP and NO3− across the test groups in the granular sludge, where PO43− and varying concentrations of NO3− were simultaneously introduced. The removal efficiency of TP was greatly influenced by the concentration of NO3−-N (electron acceptor). Zhang et al. [36] investigated the influence of the nitritation liquid recycling ratio (100–500%) on the system’s pollutant removal efficiency. They discovered that lower nitritation liquid recycling ratios (100–200%) reduced the availability of electron acceptors for nitrite, thereby restricting the performance of denitrifying phosphorus removal. Their findings are consistent with the experimental results of the present study. The molar ratio of TP to NO3−-N removal was calculated as 1:1.614. This further explains that in the aforementioned continuous-flow reaction system, as the sludge granulation and denitrification performance improved, the TN removal efficiency in the reactor reached 73.24% after the reactor ran for 40 days, leading to a decrease in NO3− recycled to the anoxic zone (15.6 mg/L), which subsequently resulted in insufficient electron acceptors for DPAOs in the anoxic zone and the shift in TP removal from the anoxic zone to the aerobic zone.
As depicted in Figure 7a, the specific phosphorus uptake rate for granular sludge in the presence of NO3− as the electron acceptor was higher than the corresponding value in the presence of molecular oxygen as the electron acceptor. The specific phosphorus uptake rates under these two conditions were calculated as 14.075 mg/(g VSS·h) and 7.025 mg/(g VSS·h), respectively. Li and Liu [37] established a one-dimensional model and simulated the diffusion of O2 in AGS. They found that the particle size of granular sludge affected the diffusion of oxygen within the sludge, and the diffusion of dissolved oxygen was the bottleneck limiting the removal of pollutants through AGS. Therefore, DPAOs were primarily distributed within the granular sludge, and the higher diffusion rate of the electron acceptor (NO3−) into the granular sludge under anaerobic conditions compared with the diffusion rate of oxygen molecules under aerobic conditions was the main reason for the higher specific phosphorus uptake rate of DPAOs under anaerobic conditions than under aerobic conditions.

3.4. Microbial Structure Analysis

To further understand the granulation process in activated sludge and the mechanisms behind its denitrification and phosphorus removal, microbial diversity analysis was performed on sludge samples collected on day 1 (seed sludge) and day 80 (AGS). The analysis of the microbial community structure for both sludge samples revealed coverage rates greater than 99.9%, indicating that the sequencing depth was sufficient to cover the entire microbial community. The results of the analysis at the phylum and genus levels with relative abundances greater than 1% in at least one sample are shown in Figure 8.
At the phylum level (Figure 8a), the relative abundances of three phyla Firmicutes, Actinobacteriota, and Proteobacteria showed significant changes between the seed sludge and granular sludge samples. Firmicutes and Actinobacteriota exhibited a substantial decrease in their relative abundance in granular sludge compared with seed sludge, decreasing from 25.14% and 21.15% to 0.76% and 0.45%, respectively. These two phyla include many filamentous bacteria [38]. Proteobacteria showed a remarkable increase in relative abundance, from 22.52% to 76.7%, in both samples. This phylum contained numerous genera closely related to denitrification and phosphorus removal, including ammonia-oxidizing bacteria such as Nitrosomonas and polyphosphate-accumulating bacteria such as Candidatus Accumulibacter [13]. Additionally, Proteobacteria are known to secrete a substantial number of EPSs [39], which are crucial for granule formation. Chloroflexi and Bacteroidota are also common in AGS reactors [40], and they play a structural role in granule formation [41]. Their relative abundances showed some variation between the seed sludge and granular sludge samples. Chloroflexi had a slightly higher relative abundance, at 10.27% and 16.06% for the seed sludge and granular sludge samples, respectively, while Bacteroidota exhibited a slight decrease, at 3.77% and 8.19%, respectively, although the overall abundance of these two phyla increased slightly. The bacterial phylum-level analysis of these two samples indicates that the “selective pressure” in the reactor had a directional selection effect on the sludge population. Filamentous bacteria with poor settling characteristics were washed out of the reactor under higher hydraulic loads, while bacteria with good settling properties, which were more likely to form granules, proliferated and accumulated in the system. This analysis aligns with the earlier conclusions regarding the granulation process of the activated sludge within the system.
At the genus level (Figure 8b), the dominant genera in the seed sludge sample were Trichococcus (18.2%) and Candidatus Microthrix (6.95%), both belonging to Firmicutes and Actinobacteriota, respectively. These genera are filamentous bacteria, and their excessive proliferation can lead to a poor settling performance of activated sludge [38,42]. Under greater “selective pressure”, a significant proportion of filamentous bacteria free from granular sludge were washed out of the system. This might be a major reason for the gradual decrease in the relative abundance of such bacteria in mature granular sludge systems. Furthermore, compared with the seed sludge, the granular sludge sample exhibited significant changes in the types and relative abundances of bacteria with denitrification and phosphorus removal functions with increasing pressure. The granular sludge sample contained typical DPAOs, including Pseudomonas (1.34%), Dechloromonas (1.05%), and a small number of PAOs such as Candidatus Accumulibacter (0.017%) [43]. Li et al. [44] found that in the A2O/A+MBR process, denitrifying phosphorus removal resulted in a 70% reduction in TP in the anoxic section with 0.92% DPAOs (Dechloromonas). In contrast, the abundance of phosphorus-removing bacteria in the seed sludge was relatively low. In the granular sludge sample, the relative abundance of typical DPAOs (Pseudomonas and Dechloromonas) was significantly higher than that of PAOs (Candidatus Accumulibacter). This disparity is attributable to the inhibitory conditions within the denitrifying phosphorus removal system, which might have hindered the growth and reproduction of PAOs, leading to their lower relative abundance in the system. This conclusion is consistent with those of previous research [35]. Additionally, the relatively abundant Defluviicoccus (28.38%) can absorb volatile fatty acids (VFAs) under anaerobic conditions and convert them into polyhydroxyalkanoates (PHA), competing for VFAs with PAOs [45]. This further inhibits the proliferation of PAOs. These results further illustrate that TP removal from the system was mainly based on denitrifying phosphorus removal. Regarding denitrifying bacteria, Rhodobacter was the dominant genus (2.71%) in the seed sludge sample, while the granular sludge sample exhibited an increase in genera with denitrification capabilities, including Azospira (3.68%), Candidatus Competibacter (2.99%), and unclassified_f__Rhodobacteraceae (1.11%) compared with the seed sludge sample [13,46]. Both the genus and relative abundance of denitrifying bacteria increased significantly in the granular sludge sample. These denitrifying bacteria, together with DPAOs, contributed to the enhancement of the system’s denitrification efficiency. Additionally, Candidatus Brocadia, a typical anaerobic ammonia-oxidizing bacterium [47], was found in the system, accounting for 0.02% of the population. These findings further suggest that the reaction system exhibited multiple denitrification mechanisms, including traditional denitrification and anaerobic ammonia oxidation, functioning in both the anaerobic and anoxic stages.
The Chao index values for the seed sludge and AGS samples were calculated as 1127 and 374, respectively, while the Shannon index values were 5.01 and 3.07, respectively. The calculation results indicate that after AGS formation, the overall diversity of the activated sludge, the abundance of microbial species, and evenness decreased to some extent [48]. However, the types and abundance of bacteria with denitrification and phosphorus removal functions significantly increased. This further elucidates the enhancement mechanism of the continuous-flow AGS process on denitrification and polyphosphate accumulation functions. According to the microbial community structure of AGS, filamentous bacteria play a structural role in AGS formation. EPSs facilitate microorganism aggregation, promoting AGS formation. However, the decrease in microbial diversity may lead to the decrease in the stability of the system, especially when the water quality changes, which may lead to the disintegration of AGS.

4. Conclusions

Under controlled gradient system conditions, after 80 days of continuous operation, activated sludge granulation was achieved in a continuous-flow AAO reactor. The activated sludge granulation process was the result of the directed screening and enrichment of microorganisms in the activated sludge system, and the process led to significant improvements in the stability and denitrification and phosphorus removal performance of the reaction system. The analysis of AGS biodiversity revealed that the bacterial genera closely associated with phosphorus removal in the reaction system were mainly typical DPAOs such as Pseudomonas and Dechloromonas. In contrast, the relative abundance of typical PAOs such as Candidatus Accumulibacter was relatively low in the granular sludge system. The AGS system exhibited a significantly higher diversity and abundance of denitrifying bacteria, including genera such as Azospira, Candidatus Competibacter, and unclassified_f_Rhodobacteraceae, than the seed sludge system. These bacteria collaborated with DPAOs to enhance the system’s denitrification efficiency. Additionally, the AGS system contained a certain number of typical anaerobic ammonia-oxidizing bacteria, Candidatus Brocadia, indicating the existence of anaerobic ammonia oxidation pathways in the reaction system.

Author Contributions

All authors contributed to the study conception and design. Material preparation, data collection, and analysis were performed by Y.Z., P.A. and X.Z.; the first draft of the manuscript was written by P.A. and revised by Y.Z. and J.W. All authors commented on previous versions of the manuscript. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by Henan Provincial Science and Technology Development Program grant number 212102110030.

Data Availability Statement

All data generated or analyzed during this study are included in this published article.

Conflicts of Interest

The authors declare no competing interests.

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Water 16 01510 g001
Figure 1. (a) Experimental setup and (b) schematic diagram.
Figure 1. (a) Experimental setup and (b) schematic diagram.
Water 16 01510 g001
Water 16 01510 g002
Figure 2. Sludge characteristics variations during the whole experiment; (a) particle size distribution, (b) MLSS, SVI30, SV5, and SV30, (c) PN, PS and PN/PS.
Figure 2. Sludge characteristics variations during the whole experiment; (a) particle size distribution, (b) MLSS, SVI30, SV5, and SV30, (c) PN, PS and PN/PS.
Water 16 01510 g002
Water 16 01510 g003
Figure 3. Changes in the morphology of activated sludge; (a) day 1, (b) day 40, (c) day 65, (d) day 80, (e) day 80 excess sludge.
Figure 3. Changes in the morphology of activated sludge; (a) day 1, (b) day 40, (c) day 65, (d) day 80, (e) day 80 excess sludge.
Water 16 01510 g003
Water 16 01510 g004
Figure 4. Changes in NH4+-N (a) influent and effluent concentrations and (b) removal efficiency in various units during the whole experiment.
Figure 4. Changes in NH4+-N (a) influent and effluent concentrations and (b) removal efficiency in various units during the whole experiment.
Water 16 01510 g004
Water 16 01510 g005
Figure 5. Changes in TN (a) influent and effluent concentrations and (b) removal efficiency in various units during the whole experiment.
Figure 5. Changes in TN (a) influent and effluent concentrations and (b) removal efficiency in various units during the whole experiment.
Water 16 01510 g005
Water 16 01510 g006
Figure 6. Changes in TP (a) influent and effluent concentrations and (b) removal efficiency in various units during the whole experiment.
Figure 6. Changes in TP (a) influent and effluent concentrations and (b) removal efficiency in various units during the whole experiment.
Water 16 01510 g006
Water 16 01510 g007
Figure 7. Changes in TP and NO3−-N in batch on day 1 (a,b) and day 80 (c,d).
Figure 7. Changes in TP and NO3−-N in batch on day 1 (a,b) and day 80 (c,d).
Water 16 01510 g007
Water 16 01510 g008
Figure 8. Microbial community analysis of seed sludge and AGS; (a) phylum level, (b) genus level.
Figure 8. Microbial community analysis of seed sludge and AGS; (a) phylum level, (b) genus level.
Water 16 01510 g008
Table 1. Composition of the synthetic wastewater.
Table 1. Composition of the synthetic wastewater.
ComponentsConcentration
(mg/L)		Trace ElementsConcentration
(μg/L)
glucose280H3BO445
sodium acetate385ZnSO4·7H2O36
NH4Cl230FeCl345
KH2PO426CuSO4·5H2O45
CaCl25MnSO436
MgSO4·7H2O55KI54
KCl8EDTA300
NaHCO3250
Table 2. Hydraulic loading during the whole experiment.
Table 2. Hydraulic loading during the whole experiment.
PhaseTime (day)Hydraulic Loading (m3/(m2·h))
Phase I1–102.04
11–202.69
21–303.34
31–403.99
Phase II41–514.50
51–655.00
Phase III65–805.00
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Zhao, Y.; An, P.; Wan, J.; Zhang, X. Enhanced Simultaneous Nitrogen and Phosphorus Removal in a Continuous-Flow Granular Sludge System under Gradient-Controlled Hydraulic Loading. Water 2024, 16, 1510. https://doi.org/10.3390/w16111510

AMA Style

Zhao Y, An P, Wan J, Zhang X. Enhanced Simultaneous Nitrogen and Phosphorus Removal in a Continuous-Flow Granular Sludge System under Gradient-Controlled Hydraulic Loading. Water. 2024; 16(11):1510. https://doi.org/10.3390/w16111510

Chicago/Turabian Style

Zhao, Yaguang, Pengkun An, Junfeng Wan, and Xuehui Zhang. 2024. "Enhanced Simultaneous Nitrogen and Phosphorus Removal in a Continuous-Flow Granular Sludge System under Gradient-Controlled Hydraulic Loading" Water 16, no. 11: 1510. https://doi.org/10.3390/w16111510

APA Style

Zhao, Y., An, P., Wan, J., & Zhang, X. (2024). Enhanced Simultaneous Nitrogen and Phosphorus Removal in a Continuous-Flow Granular Sludge System under Gradient-Controlled Hydraulic Loading. Water, 16(11), 1510. https://doi.org/10.3390/w16111510

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