The Nitrogen Removal Performance and Functional Bacteria in Heterotrophic Denitrification and Mixotrophic Denitrification Process

Abstract

The heterotrophic and autotrophic synergistic denitrification (HAD) system can effectively remove sulfide, nitrate, and organic carbon pollutants from municipal wastewater. However, the effect of sulfide on the functional bacteria in the denitrification system is still unclear. To better understand the mechanism of sulfide affected on bacteria in the system, the up-flow anaerobic sludge blanket (UASB) reactor was operated continuously under heterotrophic (no sulfide added) and mixotrophic conditions (with increased sulfide contents) for 120 days. The contents of protein (PN) in extracellular polymeric substances (EPS) were significantly increased with the addition of sulfide, which enhanced the flocculation of sludge and was beneficial to the formation of high-density microorganism communities. The dominant bacteria showed large differences under different nutrient conditions. The abundances of Thauera increased from 4.13% to over 12.94%, and that of Dechloromonas and Thiobacillus were 2.61–3.01% and 1.04–2.66% respectively after added sulfide. And the efficient performance of the system in mixotrophic conditions was accomplished with the interaction of heterotrophic sulfide-oxidizing, nitrate-reducing bacteria (Thauera, Dechloromonas), autotrophic sulfide-oxidizing, nitrate-reducing bacteria (Thiobacillus) and heterotrophic nitrate-reducing bacteria (Rubrivivax, Acidovorax, Simplicispira, Alicycliphilus). Moreover, the abundances of Nar G, Nap A, Nir S, Nor B, and Nos Z were significantly enhanced in mixotrophic conditions, indicating that the nitrogen metabolism potential of the system was also improved after added sulfide. These results elucidated the reasons for the enhanced denitrifying capacity of the system by adding S²⁻ from the microbiological point of view and provided a theoretical basis for the establishment of an efficient denitrification system.

1. Introduction

Compared with various wastewater treatment methods for eliminating nitrogen-containing pollutants, biological denitrification has many benefits and is recognized as an economical and effective method for treating nitrate pollutants in wastewater [1,2,3]. In recent decades, the traditional biological denitrification process based on the principle of heterotrophic denitrification (HD) has been widely utilized around the world [4,5]. As shown in Equation (1), HD is the process of heterotrophic microorganisms using organic carbon sources (such as CH₃COO⁻) as electron donors to convert nitrate to nitrogen under anoxic and anaerobic conditions [6]. In the process of HD, the contents of organic carbon sources as a crucial role and have a significant impact on denitrification efficiency [7]. In reality, the C/N ratio of municipal wastewater is usually low, and the lack of organic carbon sources limits the denitrifying efficiency of HD [8]. To satisfy the effluent requirements, organic carbon sources are usually needed to add to the treatment process. Although the effectiveness of denitrification is improved by the addition of organic carbon sources, there are many problems such as expensive organic carbon sources, increased sludge production, high cost of subsequent treatments, and increased CO₂ emissions (not in line with the development demands of the era of carbon neutrality and carbon peaking) [9]. Therefore, there is an urgent need to find new methods or strategies to get rid of the dependence on carbon sources in low-C/N wastewater treatment processes.

$$2N{O}_3^{-}+1.25C{H}_{3}CO{O}^{-}\to 2.5C{O}_2+N_2+0.25H_{2}O+3.25O{H}^{-}$$

(1)

In response, many researchers have proposed using autotrophic denitrification (AD) and heterotrophic denitrification (HD) to construct the synergistic denitrification (HAD) system to solve the problem of insufficient organic carbon sources in the denitrification process [1,10,11]. AD is one of the biological denitrification processes without the requirement of organic carbon sources, but utilizes inorganic substances such as H₂ and reduced sulfur compounds (such as S²⁻) to provide the electron donors required for nitrate reduction (as shown in Equations (2) and (3)) [12]. Therefore, the introduction of inorganic electron donors can effectively solve the problem of insufficient organic carbon sources in the wastewater treatment process. Furthermore, the inorganic electron donors required for AD are widely available, especially for sulfide, which often coexists with nitrogen-containing and organic pollutants in municipal wastewater [13,14]. Sulfide is mainly produced by sulfate reduction in municipal wastewater. Sulfate pollution is often present in municipal wastewater because of human production activities. During the pipeline transportation of wastewater, sulfate-reducing bacteria (SRB) can reduce sulfate to sulfide which utilizes organic matters as electron donors (as shown in Equation (4)) [15]. Previous studies had shown that the sulfide contents of municipal wastewater after pipeline transport could reach the range of 1 mg/L to 20 mg/L [16,17,18,19,20]. Although the sulfide contents of such municipal wastewater are relatively low, they still have serious hazards for the environment and humans. In this regard, the traditional methods usually used to inhibit sulfide generation or degrade sulfide pollution by adding oxidants, aeration, oxygenation, and adding iron salts, which greatly increased the cost of wastewater treatment and the difficulty of subsequent treatment [20,21,22]. However, this part of sulfide can be used as an electron donor for denitrification treatment of HAD system, which not only reduced the supplement of organic carbon sources but also simultaneously solved the problem of sulfide pollution and effectively reduced the cost of wastewater treatment.

$$S^{2-}+0.4N{O}_3^{-}+2.4H^{+}\to S^0+0.2N_2+1.2H_{2}O$$

(2)

$$S^{2-}+1.6N{O}_3^{-}+1.6H^{+}\to S{O}_4^{2-}+0.8N_2+0.8H_{2}O$$

(3)

$$S{O}_4^{2-}+C{H}_{3}CO{O}^{-}+3H^{+}\to 2C{O}_2+H_{2}S+2H_{2}O$$

(4)

Although many studies have demonstrated the efficient denitrification performance of HAD system, few studies focused on the changes of functional bacteria in the system during the conversion from HD to HAD system after the introduction of sulfide. On the one hand, the growth and activity of some heterotrophic denitrifying bacteria may be inhibited due to the toxicity of sulfide, which leads to a decrease in the denitrification capacity of HD [23]. On the other hand, the growth rate of heterotrophic denitrifying bacteria is much faster than that of autotrophic denitrifying bacteria [24]. It is still to be investigated whether the autotrophic bacteria can grow stably in the system and perform efficient denitrification capacity when sulfide is introduced. In addition, it has recently been reported that heterotrophic sulfide-oxidizing, nitrate-reducing bacteria (h-soNRB) [25], and these bacteria can oxidize sulfide and use organic matters for denitrification metabolism, which breaks the traditional theoretical notion that the removal of nitrate, sulfide and organic carbon pollutants in HAD system is only accomplished by heterotrophic denitrifying bacteria and autotrophic denitrifying bacteria. Therefore, it is necessary to master the changes of functional bacteria during the conversion of HD to HAD, which is the most fundamental means of detecting and regulating the operational efficiency of the system.

Based on these, in this study, the heterotrophic denitrification (HD) system was first constructed in the up-flow anaerobic sludge blanket (UASB) reactor and the quality of simulated influent was based on municipal wastewater quality. After the performance of the HD system was stable, sulfide was added to the influent, thus the HD system was successfully transformed into the HAD system. During the operation of the system at different stages, the other conditions were kept the same and the concentrations of sulfide in the influent were gradually increased until the best operating efficiency of the system was found. In addition, the effect of sulfide on the structure, community, and nitrogen metabolism functions of microorganisms in the system was investigated by testing the contents of extracellular polymeric substances (EPS), the abundance of functional bacteria, and the functional genes related to nitrogen metabolic activities. The purpose of this study is to examine the mechanism by which sulfide affects the denitrification capacity of the system from the microbiological point of view and to provide a theoretical basis for the establishment of efficient denitrification systems.

2. Materials and Methods

2.1. System Setup and Operation Conditions

As shown in Figure 1, an up-flow anaerobic sludge blanket (UASB) reactor with a working volume of 18 L (diameter = 21 cm, working height = 52 cm) was used to construct the denitrification system. The reactor was made of plexiglass and worked at room temperature (25 ± 3 °C). Simulation of wastewater was pumped into the bottom of the reactor through a peristaltic pump drive and the specific compositions of the simulated wastewater were shown in

Table 1. Influent conditions of the UASB reactor during different stages.

Stage	Nitrate-N (mg/L)	COD (mg/L)	S2− (mg/L)	TP (mg/L)	HRT (h)	C/N	S/N
I (1–30 d)	30	100	0	5	4	3.33	-
II (31–60 d)	30	100	5	5	4	3.33	0.17
III (61–90 d)	30	100	10	5	4	3.33	0.33
IV (91–120 d)	30	100	20	5	4	3.33	0.67

. The hydraulic retention time (HRT) in the system was kept at 4 h. The inoculated sludge for the reactor came from the anoxic section of the Lijiao wastewater treatment plant (Guangzhou, China). The ratio of mixed liquid volatile suspended solids (MLVSS) to mixed liquid suspended solids (MLSS) in inoculated sludge was about 0.72. The reactor was wrapped with aluminum foil to prevent the effects of light and the growth of photosynthetic bacteria. To prevent the oxidation of sulfide, dissolved oxygen was removed from the wastewater using high-purity nitrogen before each water distribution. Na₂S·9H₂O (AR, ≥98.0%, S817428), KNO₃ (ACS, >99.0%, P111637), Glucose (≥98.0%, D810558) and KH₂PO₄ (AR, ≥99.5%, P815662) were purchased from Macklin (Shanghai, China) and used as the sources of sulfur, nitrogen, carbon, and phosphorus, respectively. Micronutrients were also essential for the normal growth of microorganisms and the specific compositions were shown in Table S1 [26].

To analyze the performance and functional bacteria under heterotrophic and mixotrophic conditions in the denitrification system, two nutrition phases (four stages) were set up (Table 1). The conditions were kept consistent throughout the operation, except for the influent sulfide contents. The concentrations of COD, NO₃⁻-N, and TP in the influent were 100 mg/L, 30 mg/L, and 5 mg/L, respectively. Stage I operated under pure heterotrophic conditions (no sulfide added) for 30 days. Then, in stages II–IV (31–120 days), the heterotrophic condition was changed to mixotrophic conditions by adding sulfide. The influent concentrations of sulfide in stages II–IV were 5 mg/L, 10 mg/L, and 20 mg/L, respectively. The influent and effluent samples of the reactor were collected daily. Then the COD, NO₃⁻-N, NO₂⁻-N, NH₄⁺-N, S^2−, and SO₄²⁻ were measured to analyze the effectiveness of the reactor in degrading pollutants.

2.2. Analytical Methods

The daily influent and effluent samples were filtered with 0.45 μm membranes and then tested for the concentrations of pollutants. The concentrations of NO₃⁻-N, NO₂⁻-N, NH₄⁺-N, MLSS, and MLVSS were measured according to standard methods [27]. COD was determined according to the fast digestion-spectrophotometric method. The concentrations of S²⁻ and SO₄²⁻ were determined according to the methylene blue spectrophotometric method and barium chromate spectrophotometry method, respectively. And the specific analysis methods of COD, S²⁻ and SO₄²⁻ were also referred to as the standard methods [27].

2.3. EPS Extraction and Analysis Methods

In this study, the formaldehyde-NaOH method was used to extract extracellular polymeric substances (EPS) from four stages of sludge samples [28]. Subsequently, the protein (PN) and polysaccharide (PS) contents in the extracted EPS were determined. The PN and PS contents were determined according to the folin-phenol method and phenol-sulfuric acid method, respectively [29,30].

2.4. Microbial Communities and Denitrifying Genes

After the removal efficiency of pollutants was stable at each stage, the sludge samples (SI–IV) were collected from the reactor at 30, 60, 90, and 120 days. The residual sludge in the sampling port and 15 mL of sludge before sampling were discarded to ensure that the collected sludge samples were representative. Then, the sludge samples were collected into sample tubes and immediately stored in the ultra-low temperature refrigerator (−80 °C). Microbial communities and denitrifying genes were completed by the high-throughput sequencing method and macro-genome sequencing method, which were done by GENEWIZ (Suzhou, China). The specific methods of the experiment were shown in the Tests S1 and S2.

3. Results and Discussion

3.1. The Performance of the Biological Denitrification System

In this study, different doses of S²⁻ addition were used to investigate the changes in the performance of the denitrification system and functional bacteria during the conversion of the system from HD to HAD. The long-term performance of the system, including the concentrations of COD, NO₃⁻-N, NO₂⁻-N, NH₄⁺-N, S²⁻ and SO₄²⁻, and the removal rates of TN, COD, and S²⁻ are shown in Figure 2.

In stage I (day1–30), there was no addition of S²⁻ and the system operated under heterotrophic conditions. At the beginning of stage I (day1–10), the removal rates of pollutants in the system were unstable. The removal rates of TN and COD were 64.3% ± 5.3% and 78.7% ± 8.6%. Subsequently, the effectiveness of the system gradually stabilized. In the subsequent operation of the system (day11–30), the concentrations of NO₃⁻-N and NO₂⁻-N in the effluent were 3.37 ± 0.76 mg/L and 6.26± 1.38 mg/L, respectively. And the removal rates of TN and COD were increased to 68.0% ± 3.7% and 86.7% ± 5.0% respectively compared with day1 to day10. In stage I, a large amount of NO₃⁻-N and NO₂⁻-N accumulated in the effluent due to the absence of organic carbon sources, which caused inefficient performance of the system.

In stage II (day31–60), after adding 5 mg/L of S²⁻ in the influent, S²⁻ was rapidly utilized in the system. Throughout stage II, no S²⁻ was detected in the effluent, and the average removal rate of S²⁻ reached 100%. The main transformation product of S²⁻ was SO₄²⁻, and the concentration of SO₄²⁻ was 13.67 ± 3.09 mg/L. Compared with stage I, the removal rate of TN increased to 73.9% ± 2.9%. And the removal rate of COD was 86.9% ± 4.5%, which was similar to stage I, indicating that 5 mg/L of S²⁻ did not inhibit heterotrophic denitrification.

In stage III (day61–90), the concentration of S²⁻ was 10 mg/L in the influent. Same as stage II, the removal rate of S²⁻ was 100%. At the initial stage III (day61–69), the overall trend of SO₄²⁻ in the effluent was gradually increased, suggesting that the autotrophic denitrification bacteria might be adapting to the increase of S²⁻ concentrations. Subsequently, the concentration of SO₄²⁻ in the effluent gradually stabilized, which was about 29.60 ± 2.63 mg/L from day 70 to day 90. At this time, the removal rate of TN also increased to 85.4% ± 2.0%. The removal rate of COD was 87.0% ± 3.8% and remained stable.

In stage IV (day91–120), further increased the S²⁻ concentration to 20 mg/L. As a whole, NO₃⁻-N, NO₂⁻-N, and S²⁻ were not detected in the effluent. The removal rate of TN was 99.6% ± 0.9%. However, the utilization of S²⁻ and COD was different. From day 91 to day 100, the concentration of SO₄²⁻ was 57.26 ± 3.93 mg/L and the removal rate of COD was 82.1% ± 4.4%. Then, the concentration of SO₄²⁻ was decreased to 52.94 ± 3.62 mg/L along with the removal rate of COD increased to 88.7% ± 4.1% from day 101 to day 120. This result indicated that the capacity of heterotrophic denitrification was inhibited by the increase of S²⁻ concentration at the beginning of stage IV. With the increased operating time, the heterotrophic bacteria gradually adapted to the new sulfide concentration and the heterotrophic denitrification activity recovered.

From the overall operation of the system, after the supplementation of S²⁻, the denitrification capacity of the system was significantly improved. Compared with stage I and stage IV, the addition of 1 mg/mL of S²⁻ allowed the effective degradation of 0.17 mg/mL NO₃⁻-N and 0.31 mg/mL NO₂⁻-N in the wastewater. This effectively reduced the amount of additional organic carbon sources and CO₂ emissions, which had good economic and environmental benefits. And during the operation, the concentration of SO₄²⁻ in the effluent was less than 63.75 mg/L, which was in line with the standard discharge limit of 250 mg/L set by the US environmental protection agency [31]. Moreover, the removal rates of COD in stages I–IV were respectively 86.7% ± 5.0%, 86.9% ± 4.5%, 87.0% ± 3.8%, and 88.7% ± 4.1% after the operation of the system was stabilized, which was not presented with great volatility. And this result indicated the capacity of heterotrophic denitrification was not inhibited with the concentration of S²⁻ increased from 5 mg/L to 20 mg/L. Overall, these results confirmed that HAD system could efficiently and stably complete the degradation of sulfide, nitrate, and organic carbon pollutants under mixotrophic conditions.

3.2. The Contents and Characteristics of EPS

Extracellular polymeric substances (EPS) are polymeric natural polymers secreted by microorganisms (mainly by bacteria) and consist mainly of protein (PN) and polysaccharide (PS) [32]. EPS is usually attached to the surface of bacteria and used to resist toxic substances [33]. The contents of EPS (SI–SIV) during the operation are shown in Figure 3.

PN and PS contents in EPS were positively correlated with influent S²⁻ concentrations. In SI, the contents of PN and PS were 32.11 ± 1.07 mg/g VSS and 14.24 ± 0.98 mg/g VSS respectively, which was the lowest in all samples. This might be due to the fact that nutrients were most scarce in SI and the microbial metabolic activity was inhibited, which goes against for the microorganisms to secrete EPS. With the increased S²⁻ in the influent, the contents of PN in SII–SIV raised significantly, and the specific contents of PN were 35.22 ± 4.30, 50.25 ± 6.75, and 80.22 ± 4.32 mg/g VSS, separately. Meanwhile, PS contents increased from 15.24 ± 0.88 (SII) to 18.30 ± 1.83 (SIII), 22.25 ± 1.83 (SIV) mg/g VSS, separately. These results indicated that the addition of S²⁻ supplemented the nutrients required for denitrification by microorganisms and accelerated the rate of microbial metabolism, leading to an increase in EPS secretion. In this study, the PN contents were much higher than the PS contents, and were the main substance in EPS. Furthermore, the ratio of PN to PS also showed an upward trend. The PN/PS were 2.25,2.31,2.79 and 3.61, separately, meaning that the percentage of PN in EPS increased. Previous studies had shown that PN contained more hydrophobic amino acids, while the hydroxyl and carboxyl groups on PS were hydrophilic [34]. Therefore, the hydrophobicity of sludge was enhanced with the increased PN contents. In thermodynamics, the hydrophobicity of the sludge surface will lead to a reduction of free energy required by microorganisms in the aggregation process [35]. This indicated that microorganisms were much easier to aggregate together when the hydrophobicity of sludge was stronger. With increased microbial aggregation effects, a highly dense and stable microbial community structure was formed in the system, which enhanced the information communication and synergy between microorganisms. Moreover, PN contained a large number of extracellular enzymes secreted by microorganisms and these enzymes could catalyze the oxidation-reduction reaction of pollutants [36]. The addition of S²⁻ increased the contents of PN and also increased the contents of biological enzymes, therefore promoting the oxidation-reduction reaction process as well as microbial activity. These results better explained that the system had the highest TN removal efficiency in stage IV.

Analyzing the above experimental results, EPS was essential in the denitrification system. EPS acted as a physical barrier between pollutants and the cell membrane, reducing the stress shock on microorganisms due to changes in the external environment. When nutrients were scarce in the water environment, microorganisms had a poor ability to secrete EPS. When nutrients were sufficient in the environment, it was beneficial for microorganisms to secrete EPS, and improved microbial metabolic activity and synergy between microorganisms, thus enhancing the denitrification capacity of the system.

3.3. Microbial Diversity and Communities

3.3.1. Alpha Diversity Index

The Alpha index reflects the diversity and richness of species in the microbial community. The variation of the Alpha index for the four stages of samples (SI–SIV) in the system is shown in

Table 2. Alpha diversity index at the different stages.

Sample	Ace	Chao1	Shannon	Simpson	Coverage
SI	276	276	4.847	0.938	1.000
SII	260	260	4.291	0.950	1.000
SIII	230	229	3.558	0.948	0.999
SIV	198	198	3.218	0.964	1.000

. From an overall perspective, the Ace, Chao1, and Shannon indexes showed a downward trend during the operation. The Chao1 index reduced from 276 to 198 and the Shannon index reduced from 4.847 to 3.218 (SI–SIV). These might be explained by the fact that the addition of S²⁻ inhibited the growth of some microorganisms, which led to the elimination of these microorganisms. In addition, the index reduction was related to the amount of S²⁻ supplementation. Compared with stage I, the Chao1 index decreased by 5.8%, 17.0%, and 28.3%, and the Shannon index decreased by 11.5%, 26.6%, and 33.6% after the addition of 5, 10, and 20 mg/L S²⁻ in stages II–IV, respectively. This indicated that the higher the amount of S²⁻ supplementation, the stronger the inhibitory effect on microbial diversity. The Simpson index reflected the evenness of individuals in the community. When the index was larger, the enrichment of genera in the system was higher. The Simpson index was 0.938 in SI. And in SII–IV, Simpson was raised to 0.948–0.964, which indicated the enrichment of the genera was increased after adding S²⁻. Therefore, we speculated that S²⁻ reduced the diversity of the community in the system, but increased the enrichment of the main denitrifying bacteria, thus promoting the denitrification capacity of the system.

3.3.2. Microbial Communities

As shown in Figure 4a, Proteobacteria, Bacteroidetes, and Chloroflexi were the dominant phyla in SI. When S²⁻ was added, the dominant phylum species in SII–SIV were similar to SI, but with different relative abundances. Proteobacteria occupied an absolute position in all four stages, and this phylum combined a wide range of respiratory and fermentative bacteria, and also included many heterotrophic denitrifying bacteria as well as sulfur-oxidizing bacteria [37,38]. Compared with SI, the relative abundances of this phylum increased from 72.15% to 73.46% after the addition of 5 mg/L S²⁻ to stage II. This might be due to the fact that the addition of S²⁻ promoted the growth of some autotrophic bacteria in this phylum which used S²⁻ as electron donors for denitrification. Subsequently, stages III–IV were added with 10 and 20 mg/L of S²⁻, and the relative abundances of Proteobacteria were reduced to 72.49% and 70.7% at this time, respectively. This suggested that the increased S²⁻ addition might have a suppressive effect on Proteobacteria because S²⁻ was toxic. Bacteroidetes were chemoheterotrophic bacteria and had the ability to degrade a wide range of complex organic substances. The relative abundances of Bacteroidetes gradually decreased from 5.58% (SI) to 4.01% (SIV), which indicated that the addition of sulfide inhibited the growth of this phylum. In contrast to the trend of Bacteroidetes, the relative abundances of Chloroflexi with an upward tendency in SI–SIV were 1.89%, 2.35%, 3.57%, and 5.67%, respectively. Chloroflexi were parthenogenic anaerobic microorganisms and contained a variety of filamentous bacteria that could act as a skeleton in activated sludge [39]. Previous studies showed that Chloroflexi was mainly distributed in the outer surface layer of flocculated sludge and was closely related to the sludge flocculating capacity [40]. Therefore, the increased abundances of Chloroflexi might have enhanced the flocculating ability of the sludge, causing a denser microbial community structure and thus enhancing the stability of the system.

The microbial community structure at the genus level is shown in Figure 4b. Under heterotrophic condition (SI), the microorganisms with relative abundances greater than 1% in the system were Simplicispira (6.16%), Thauera (4.13%), Trichococcus (2.25%), Comamonas (2.18%), Rubrivivax (2.15%), Acidovorax (1.35%), Candidatus_Accumulibacter (1.28%) and Pseudomonas (1.12%), respectively. Among them, Simplicispira, Thauera, Comamonas, Rubrivivax, Acidovorax, and Pseudomonasare were defined as denitrifying bacteria, which could use organic carbon sources as electron donors to degrade nitrate pollution [41,42,43]. It could be seen that, although there were many species of denitrifying bacteria in the system when nutrients were insufficient, their abundances were at low levels.

After adding S²⁻ (SII–SIV), the microbial community in the system changed significantly. The relative abundances of Simplicispira extremely reduced from 6.16% to 1.38%~1.7% along with Comamonas, Pseudomonas, Trichococcus and Candidatus_ Accumulibacter decreased to less than 1%. This phenomenon might be due to the toxicity of S²⁻ which inhibited the growth of these bacteria. Notably, the relative abundances of Thauera rapidly increased from 4.13% to 12.94%~15.6%. Recent studies reported that Thauera not only had the ability to use organic carbon sources for denitrification, but also had the ability to oxidize S²⁻. Therefore, there might be two reasons for the growth of Thauera. Firstly, Thauera might be able to use organic carbon sources and S²⁻ as electron donors for denitrification, thus promoting its growth. Secondly, the toxicity of S²⁻ might inhibit the growth of some other denitrifying bacteria and reduced the competitive effect of these bacteria with Thauera. The relative abundances of Rubrivivax and Acidovorax also increased to approximately 2.87% and 2.0%, separately, indicating that mixotrophic conditions were more suitable for their growth. Moreover, Dechloromonas, Thiobacillus, and Alicycliphilus emerged as the new dominant bacteria after adding S²⁻. Dechloromonas was usually observed in the denitrification systems, but there was no uniform definition of its specific role. Several studies showed that Dechloromonas was able to use carbon sources for the denitrification process [44]. And other studies defined it as the autotrophic denitrifying bacteria and observed its predominance in autotrophic denitrifying biofilters with S²⁻ as an electron donor [45,46]. Therefore, we speculated that Dechloromonas might be able to use carbon sources and S²⁻ as electron donors for denitrification, which was similar to Thauera. The relative abundances of Dechloromonas increased from 0.86% to 2.61–3.01%, which indicated that the coexistence of organic carbon source and S²⁻ was beneficial for its growth. Thiobacillus was able to use inorganic carbon sources such as carbon dioxide, carbonate, or bicarbonate, and gained energy in the process of oxidizing sulfur compounds and transforming nitrate or nitrite into nitrogen gas for release [47]. The relative abundances of Thiobacillus were 1.04%, 2.28%, and 2.66% in S2-S4, respectively, which were positively correlated with the influent S²⁻ contents. This phenomenon proved that the addition of S²⁻ increased the autotrophic denitrification capacity, which was beneficial for the system to remove TN. The relative abundances of Alicycliphilus were about 1.82%. Alicycliphilus was also usually found in wastewater treatment systems, and used organic carbon sources for denitrification [41].

Summarizing the above experimental results, it was found that the operational efficiency exhibited by the system was the result of the interaction of various microorganisms. When nutrients were insufficient, although some denitrifying bacteria were observed in the heterotrophic stage, their abundances were relatively low and the system operated with poor efficiency. After adding S²⁻ to supplement the electron donors required for denitrification, the abundances of denitrifying bacteria in the system were significantly increased, therefore improving the denitrification capacity of the system. Based on the types of electron donors used by denitrifying bacteria, these bacteria could be classified into three groups: heterotrophic nitrate-reducing bacteria (hNRB), autotrophic sulfide-oxidizing, nitrate-reducing bacteria (a-soNRB) and heterotrophic sulfide-oxidizing, nitrate-reducing bacteria (h-soNRB) [48]. Due to the properties of Thauera and Dechloromonas and their variations in the system, we could classify them as heterotrophic sulfide-oxidizing, nitrate-reducing bacteria (h-soNRB). Therefore, the efficient denitrification capacity of the system under mixotrophic conditions was caused by the interaction of h-soNRB (Thauera, Dechloromonas), a-soNRB (Thiobacillus), and hNRB (Rubrivivax, Acidovorax, Simplicispira, Alicycliphilus).

3.4. Abundances of Denitrifying Genes in Different Stages

To better understand the effect of S²⁻ on the nitrogen metabolism pathways of the system, the abundances of denitrifying genes in the system were examined by macro-genome sequencing methods. As shown in Figure 5a, the process of nitrate reduced to N₂ could be divided into four steps: (1) firstly, nitrate was reduced to nitrite by the action of nitrate reductases (Nar GHI, Nap AB); (2) and nitrite was reduced to nitric oxide by the action of nitrite reductases (Nir K, Nir S); (3) subsequently, nitric oxide was reduced to nitrous oxide by the action of nitric oxide reductases (Nor B, Nor C); (4) Eventually, nitrous oxide was reduced to N₂ by the action of nitrous oxide reductases (Nos Z).

The changes in denitrifying gene abundances at different stages are shown in Figure 5b. From an overall perspective, the denitrifying gene abundances showed an upward trend with the increased S²⁻ contents. In the nitrate reductases, Nar G and Nap A were the key denitrifying genes and their abundance was much higher than Nar H, Nar I, and Nap B. With the addition of S²⁻, the abundances of Nar G were increased from 460.25PRM (SI) to 547.23PRM (SIV) along with Nap A raised from 351.18 PRM (SI) to 398.27 PRM (SIV), which directly indicated that the S²⁻ promoted the degradation of NO₃⁻-N. The abundances of Nir K were 270.14, 269.20, 277.92, and 319.47PRM and that of Nir S were 530.12, 589.46, 545.94, and 561.64 PRM in SI–SIV, respectively. It could be seen that the abundances of Nir K and Nir S were improved with the addition of S²⁻, suggesting that S²⁻ promoted the reduction of nitrite to nitric oxide. Moreover, the abundances of Nor B and Nos Z were positively correlated with influent S²⁻ contents, which indicated that the S²⁻ also promoted the reduction of nitric oxide to nitrous oxide and nitrous oxide to nitrogen. In this study, the increased abundance of denitrifying genes could be caused by the following reasons. On the one hand, the addition of S²⁻ supplemented the electron donors required for denitrification reactions and promoted the growth of denitrifying microorganisms, so that the abundances of denitrifying genes were increased. On the other hand, the addition of S²⁻ enhanced the synergy among microorganisms and accelerated nitrogenous pollutants degradation by microorganisms, therefore stimulating the up-regulation of denitrifying gene abundances. Overall, the abundances of Nar G, Nap A, Nir S, Nor B, and Nos Z were significantly enhanced through added S²⁻ to supplement the electron donors required by the denitrification process, which promoted the whole process of nitrate reduced to nitrogen and increased the nitrogen metabolism potential of the system.

4. Conclusions

In this study, the supplementation of S²⁻ successfully switched heterotrophic denitrification to autotrophic-heterotrophic denitrification, and significantly enhanced the TN removal rate of the system. When the mass ratio of COD/N/S was 3.33:1:0.67, the TN removal rate of the system could reach 99.6%. Under mixotrophic conditions, the efficient denitrification capacity of the system was achieved with the interaction of h-so NRB (Thauera, Dechloromonas), a-so NRB (Thiobacillus) and h-NRB (Rubrivivax, Acidovorax, Simplicispira, Alicycliphilus). In addition, the abundances of denitrifying genes, such as Nar G, Nap A, Nir S, Nor B, and Nos Z, were significantly enhanced under mixotrophic conditions, which indicated the nitrogen metabolism potential of the system was improved after adding S²⁻. These results confirmed the efficient pollutants removal capacity of the heterotrophic and autotrophic denitrification (HAD) system and revealed the influence of complex bacteria on the removal of TN in the system, which provided a theoretical basis for the establishment of an efficient synergistic denitrification system.