Next Article in Journal
Effective Removal of Reactive and Direct Dyes from Colored Wastewater Using Low-Cost Novel Bentonite Nanocomposites
Previous Article in Journal
Implementation of FeSO4·H2O as an Eco-Friendly Coagulant for the Elimination of Organic Pollutants from Tertiary Palm Oil Mill Effluent: Process Optimization, Kinetics, and Thermodynamics Studies
 
 
Search for Articles:
Title / Keyword
Author / Affiliation / Email
Journal
Article Type
 
 
Advanced Search
 
Section
Special Issue
Volume
Issue
Number
Page
 
Journals
Water
Volume 14
Issue 22
10.3390/w14223603
water-logo
Submit to this Journal Review for this Journal Propose a Special Issue
► Article Menu

Article Menu

share Share announcement Help format_quote Cite question_answer Discuss in SciProfiles thumb_up
...
Endorse textsms
...
Comment
first_page
settings
Order Article Reprints
Font Type:
Arial Georgia Verdana
Font Size:
Aa Aa Aa
Line Spacing:
Column Width:
Background:
Article

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

by
Xinghao Ren
1,
Yan Wang
1,2,*,
Jinquan Wan
1,2,
Zhicheng Yan
1,
Yongwen Ma
1,2,
Guihua Zhang
3 and
Bin Zhu
4
1
College of Environment and Energy, South China University of Technology, Guangzhou 510006, China
2
Guangdong Plant Fiber High-Valued Cleaning Utilization Engineering Technology Research Center, Guangzhou 510006, China
3
Huizhou Institute of Environmental Science, Huizhou 516000, China
4
Guangdong Zihua Technology Co., Ltd., Foshan 528300, China
*
Author to whom correspondence should be addressed.
Water 2022, 14(22), 3603; https://doi.org/10.3390/w14223603
Submission received: 10 October 2022 / Revised: 4 November 2022 / Accepted: 4 November 2022 / Published: 8 November 2022
(This article belongs to the Section Wastewater Treatment and Reuse)
Browse Figures
Versions Notes

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.
2 N O 3 − + 1.25 C H 3 C O O − → 2.5 C O 2 + N 2 + 0.25 H 2 O + 3.25 O H −  
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 − + 0.4 N O 3 − + 2.4 H + → S 0 + 0.2 N 2 + 1.2 H 2 O  
S 2 − + 1.6 N O 3 − + 1.6 H + → S O 4 2 − + 0.8 N 2 + 0.8 H 2 O  
S O 4 2 − + C H 3 C O O − + 3 H + → 2 C O 2 + H 2 S + 2 H 2 O
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. 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, S2−, 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. 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.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w14223603/s1, Table S1: the concentrations of micronutrients in this experiment; Test S1: high-throughput sequencing method; Test S2: metagenomic sequencing Method.

Author Contributions

Conceptualization, X.R. and Y.W.; methodology, X.R. and Y.W.; validation, X.R. and Y.W.; formal analysis, X.R.; investigation, X.R. and J.W.; resources, J.W., Y.W., Y.M., Z.Y. and G.Z.; writing—original draft preparation, X.R.; writing—review and editing, X.R., Y.W. and B.Z.; visualization, X.R. and Z.Y.; supervision, Y.W.; project administration, X.R., Y.W. and J.W.; funding acquisition, J.W., Y.W., Y.M., B.Z. and G.Z. All authors have read and agreed to the published version of the manuscript.

Funding

This study was supported by the National Key Research and Development Project (No. 2018YFE0110400), Foshan Science and Technology Innovation Project of Guangdong Province (No. 2130218003140), Guangdong Special Support Program Project (No. 2021JC060580) and National Natural Science Foundation of China (No. 21978102 and No. 22278156).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Not applicable.

Conflicts of Interest

The authors declare no conflict of interest.

References

  1. Li, W.; Lin, J.G.; Zhang, L.B.; Zhu, K.P.; Liu, X.Y.; Wang, Z.; Cao, B.X.; Guo, P. Mixotrophic Denitrification Desulfurization Wastewater Treatment Process: Bioreactor Performance and Analysis of the Microbial Community. Pol. J. Environ. Stud. 2016, 25, 2491–2497. [Google Scholar] [CrossRef] [Green Version]
  2. Moussavi, G.; Jafari, S.J.; Yaghmaeian, K. Enhanced biological denitrification in the cyclic rotating bed reactor with catechol as carbon source. Bioresour. Technol. 2015, 189, 266–272. [Google Scholar] [CrossRef]
  3. Wang, D.B.; Wang, Y.L.; Liu, Y.W.; Ngo, H.H.; Lian, Y.; Zhao, J.W.; Chen, F.; Yang, Q.; Zeng, G.M.; Li, X.M. Is denitrifying anaerobic methane oxidation-centered technologies a solution for the sustainable operation of wastewater treatment Plants? Bioresour. Technol. 2017, 234, 456–465. [Google Scholar] [CrossRef] [PubMed]
  4. Fu, X.R.; Hou, R.R.; Yang, P.; Qian, S.T.; Feng, Z.Q.; Chen, Z.B.; Wang, F.; Yuan, R.F.; Chen, H.L.; Zhou, B.H. Application of external carbon source in heterotrophic denitrification of domestic sewage: A review. Sci. Total Environ. 2022, 817, 153061. [Google Scholar] [CrossRef] [PubMed]
  5. Xu, G.H.; Peng, J.J.; Feng, C.J.; Fang, F.; Chen, S.H.; Xu, Y.J.; Wang, X.Z. Evaluation of simultaneous autotrophic and heterotrophic denitrification processes and bacterial community structure analysis. Appl. Microbiol. Biotechnol. 2015, 99, 6527–6536. [Google Scholar] [CrossRef]
  6. Pang, Y.M.; Wang, J.L. Various electron donors for biological nitrate removal: A review. Sci. Total Environ. 2021, 794, 148699. [Google Scholar] [CrossRef]
  7. Sun, S.P.; Nacher, C.P.I.; Merkey, B.; Zhou, Q.; Xia, S.Q.; Yang, D.H.; Sun, J.H.; Smets, B.F. Effective Biological Nitrogen Removal Treatment Processes for Domestic Wastewaters with Low C/N Ratios: A Review. Environ. Eng. Sci. 2010, 27, 111–126. [Google Scholar] [CrossRef] [Green Version]
  8. Huang, X.; Duan, C.S.; Yu, J.H.; Dong, W.Y. Transforming heterotrophic to autotrophic denitrification process: Insights into microbial community, interspecific interaction and nitrogen metabolism. Bioresour. Technol. 2022, 345, 126471. [Google Scholar] [CrossRef]
  9. Hellman, M.; Bonilla-Rosso, G.; Widerlund, A.; Juhanson, J.; Hallin, S. External carbon addition for enhancing denitrification modifies bacterial community composition and affects CH4 and N2O production in sub-arctic mining pond sediments. Water Res. 2019, 158, 22–33. [Google Scholar] [CrossRef]
  10. Chen, C.; Ren, N.Q.; Wang, A.J.; Yu, Z.G.; Lee, D.J. Simultaneous biological removal of sulfur, nitrogen and carbon using EGSB reactor. Appl. Microbiol. Biotechnol. 2008, 78, 1057–1063. [Google Scholar] [CrossRef]
  11. Tang, K.; An, S.J.; Nemati, M. Evaluation of autotrophic and heterotrophic processes in biofilm reactors used for removal of sulphide, nitrate and COD. Bioresour. Technol. 2010, 101, 8109–8118. [Google Scholar] [CrossRef] [PubMed]
  12. Huang, S.; Yu, D.S.; Chen, G.H.; Wang, Y.Y.; Tang, P.; Liu, C.C.; Tian, Y.; Zhang, M. Realization of nitrite accumulation in a sulfide-driven autotrophic denitrification process: Simultaneous nitrate and sulfur removal. Chemosphere 2021, 278, 130413. [Google Scholar] [CrossRef] [PubMed]
  13. Show, K.Y.; Lee, D.J.; Pan, X.L. Simultaneous biological removal of nitrogen-sulfur-carbon: Recent advances and challenges. Biotechnol. Adv. 2013, 31, 409–420. [Google Scholar] [CrossRef]
  14. Wang, B.; Wu, D.; Dai, J.; Ekama, G.A.; Hao, X.D.; Chen, G.H. Elucidating the effects of starvation and reactivation on anaerobic sulfidogenic granular sludge: Reactor performance and granular sludge transformation. Water Res. 2019, 151, 44–53. [Google Scholar] [CrossRef] [PubMed]
  15. Goorissen, H.P.; Stams, A.J.M.; Hansen, T.A. Methanol utilization in defined mixed cultures of thermophilic anaerobes in the presence of sulfate. FEMS Microbiol. Ecol. 2004, 49, 489–494. [Google Scholar] [CrossRef] [PubMed]
  16. Jiang, G.M.; Gutierrez, O.; Sharma, K.R.; Keller, J.; Yuan, Z.G. Optimization of intermittent, simultaneous dosage of nitrite and hydrochloric acid to control sulfide and methane productions in sewers. Water Res. 2011, 45, 6163–6172. [Google Scholar] [CrossRef]
  17. Mohanakrishnan, J.; Gutierrez, O.; Sharma, K.R.; Guisasola, A.; Werner, U.; Meyer, R.L.; Keller, J.; Yuan, Z. Impact of nitrate addition on biofilm properties and activities in rising main sewers. Water Res. 2009, 43, 4225–4237. [Google Scholar] [CrossRef]
  18. Rathnayake, D.; Sathasivan, A.; Kastl, G.; Krishna, K.C.B. Hydrogen sulphide control in sewers by catalysing the reaction with oxygen. Sci. Total Environ. 2019, 689, 1192–1200. [Google Scholar] [CrossRef]
  19. Yan, X.F.; Sun, J.; Kenjiahan, A.; Dai, X.H.; Ni, B.J.; Yuan, Z.G. Rapid and strong biocidal effect of ferrate on sulfidogenic and methanogenic sewer biofilms. Water Res. 2020, 169, 115208. [Google Scholar] [CrossRef]
  20. Zhang, L.; De Schryver, P.; De Gusseme, B.; De Muynck, W.; Boon, N.; Verstraete, W. Chemical and biological technologies for hydrogen sulfide emission control in sewer systems: A review. Water Res. 2008, 42, 1–12. [Google Scholar] [CrossRef]
  21. Nielsen, A.H.; Lens, P.; Vollertsen, J.; Hvitved-Jacobsen, T. Sulfide-iron interactions in domestic wastewater from a gravity sewer. Water Res. 2005, 39, 2747–2755. [Google Scholar]
  22. Park, K.; Lee, H.; Phelan, S.; Liyanaarachchi, S.; Marleni, N.; Navaratna, D.; Jegatheesan, V.; Shu, L. Mitigation strategies of hydrogen sulphide emission in sewer networks—A review. Int. Biodeterior. Biodegrad. 2014, 95, 251–261. [Google Scholar] [CrossRef]
  23. Wisniewski, K.; di Biase, A.; Munz, G.; Oleszkiewicz, J.A.; Makinia, J. Kinetic characterization of hydrogen sulfide inhibition of suspended anammox biomass from a membrane bioreactor. Biochem. Eng. J. 2019, 143, 48–57. [Google Scholar] [CrossRef]
  24. Joo, H.S.; Hirai, M.; Shoda, M. Characteristics of ammonium removal by heterotrophic nitrification-aerobic denitrification by Alcaligenes faecalis no. 4. J. Biosci. Bioeng. 2005, 100, 184–191. [Google Scholar] [CrossRef]
  25. Huang, C.; Liu, Q.; Li, Z.L.; Ma, X.D.; Hou, Y.N.; Ren, N.Q.; Wang, A.J. Relationship between functional bacteria in a denitrification desulfurization system under autotrophic, heterotrophic, and mixotrophic conditions. Water Res. 2021, 188, 116526. [Google Scholar] [CrossRef]
  26. Huang, C.; Liu, Q.; Chen, C.; Chen, F.; Zhao, Y.K.; Gao, L.F.; Liu, W.Z.; Zhou, J.Z.; Li, Z.L.; Wang, A.J. Elemental sulfur recovery and spatial distribution of functional bacteria and expressed genes under different carbon/nitrate/sulfide loadings in up-flow anaerobic sludge blanket reactors. J. Hazard. Mater. 2017, 324, 48–53. [Google Scholar] [CrossRef] [Green Version]
  27. APHA. Standard Methods for the Examination of Water and Wastewater, 21st ed.; American Public Health Association: Washington, DC, USA, 2012. [Google Scholar]
  28. Adav, S.S.; Lee, D.J. Characterization of extracellular polymeric substances (EPS) from phenol degrading aerobic granules. J. Taiwan Inst. Chem. Eng. 2011, 42, 645–651. [Google Scholar] [CrossRef]
  29. Adav, S.S.; Lee, D.J. Extraction of extracellular polymeric substances from aerobic granule with compact interior structure. J. Hazard. Mater. 2008, 154, 1120–1126. [Google Scholar] [CrossRef]
  30. Luo, J.H.; Hao, T.W.; Wei, L.; Mackey, H.R.; Lin, Z.Q.; Chen, G.H. Impact of influent COD/N ratio on disintegration of aerobic granular sludge. Water Res. 2014, 62, 127–135. [Google Scholar] [CrossRef]
  31. US Environmental Protection Agency (US EPA). Drinking Water Advisory: Consumer Acceptability Advice and Health Effects Analysis on Sulfate; US Environmental Protection Agency: Washington, DC, USA, 2003.
  32. Bichai, F.; Barbeau, B.; Dullemont, Y.; Hijnen, W. Role of predation by zooplankton in transport and fate of protozoan (oo)cysts in granular activated carbon filtration. Water Res. 2010, 44, 1072–1081. [Google Scholar] [CrossRef]
  33. More, T.T.; Yadav, J.S.S.; Yan, S.; Tyagi, R.D.; Surampalli, R.Y. Extracellular polymeric substances of bacteria and their potential environmental applications. J. Environ. Manag. 2014, 144, 1–25. [Google Scholar] [CrossRef] [PubMed]
  34. Han, F.; Ye, W.; Wei, D.; Xu, W.Y.; Du, B.; Wei, Q. Simultaneous nitrification-denitrification and membrane fouling alleviation in a submerged biofilm membrane bioreactor with coupling of sponge and biodegradable PBS carrier. Bioresour. Technol. 2018, 270, 156–165. [Google Scholar] [CrossRef] [PubMed]
  35. Wilen, B.M.; Jin, B.; Lant, P. The influence of key chemical constituents in activated sludge on surface and flocculating properties. Water Res. 2003, 37, 2127–2139. [Google Scholar] [CrossRef]
  36. Laspidou, C.S.; Rittmann, B.E. A unified theory for extracellular polymeric substances, soluble microbial products, and active and inert biomass. Water Res. 2002, 36, 2711–2720. [Google Scholar] [CrossRef]
  37. Chu, L.B.; Wang, J.L. Denitrification of groundwater using PHBV blends in packed bed reactors and the microbial diversity. Chemosphere 2016, 155, 463–470. [Google Scholar] [CrossRef]
  38. Liao, R.H.; Shen, K.; Li, A.M.; Shi, P.; Li, Y.; Shin, Q.Q.; Wang, Z. High-nitrate wastewater treatment in an expanded granular sludge bed reactor and microbial diversity using 454 pyrosequencing analysis. Bioresour. Technol. 2013, 134, 190–197. [Google Scholar] [CrossRef]
  39. Yamada, T.; Sekiguchi, Y.; Imachi, H.; Kamagata, Y.; Ohashi, A.; Harada, H. Diversity, localization, and physiological properties of filamentous microbes belonging to Chloroflexi subphylum I in mesophilic and thermophilic methanogenic sludge granules. Appl. Environ. Microbiol. 2005, 71, 7493–7503. [Google Scholar] [CrossRef] [PubMed] [Green Version]
  40. Xiao, Y.; Zheng, Y.; Wu, S.; Yang, Z.H.; Zhao, F. Bacterial Community Structure of Autotrophic Denitrification Biocathode by 454 Pyrosequencing of the 16S rRNA Gene. Microb. Ecol. 2015, 69, 492–499. [Google Scholar] [CrossRef]
  41. Hille, A.; He, M.; Ochmann, C.; Neu, T.; Horn, H. Application of two component biodegradable carriers in a particle-fixed biofilm airlift suspension reactor: Development and structure of biofilms. Bioprocess Biosyst. Eng. 2009, 32, 31–39. [Google Scholar] [CrossRef]
  42. Kim, K.H.; Ten, L.N.; Liu, Q.M.; Im, W.T.; Lee, S.T. Comamonas granuli sp nov., isolated from granules used in a wastewater treatment plant. J. Microbiol. 2008, 46, 390–395. [Google Scholar] [CrossRef]
  43. Liao, R.H.; Li, Y.; Yu, X.M.; Shi, P.; Wang, Z.; Shen, K.; Shi, Q.Q.; Miao, Y.; Li, W.T.; Li, A.M. Performance and microbial diversity of an expanded granular sludge bed reactor for high sulfate and nitrate waste brine treatment. J. Environ. Sci. 2014, 26, 717–725. [Google Scholar] [CrossRef]
  44. Lu, H.J.; Chandran, K.; Stensel, D. Microbial ecology of denitrification in biological wastewater treatment. Water Res. 2014, 64, 237–254. [Google Scholar] [CrossRef]
  45. Coates, J.D.; Chakraborty, R.; Lack, J.G.; O’Connor, S.M.; Cole, K.A.; Bender, K.S.; Achenbach, L.A. Anaerobic benzene oxidation coupled to nitrate reduction in pure culture by two strains of Dechloromonas. Nature 2001, 411, 1039–1043. [Google Scholar] [CrossRef] [PubMed]
  46. Zhou, W.L.; Li, Y.; Liu, X.; He, S.B.; Huang, J.C. Comparison of microbial communities in different sulfur-based autotrophic denitrification reactors. Appl. Microbiol. Biotechnol. 2017, 101, 447–453. [Google Scholar] [CrossRef] [PubMed]
  47. Yang, Y.; Gerrity, S.; Collins, G.; Chen, T.H.; Li, R.H.; Xie, S.H.; Zhan, X.M. Enrichment and characterization of autotrophic Thiobacillus denitrifiers from anaerobic sludge for nitrate removal. Process Biochem. 2018, 68, 165–170. [Google Scholar] [CrossRef]
  48. Zhang, R.C.; Xu, X.J.; Chen, C.; Xing, D.F.; Shao, B.; Liu, W.Z.; Wang, A.J.; Lee, D.J.; Ren, N.Q. Interactions of functional bacteria and their contributions to the performance in integrated autotrophic and heterotrophic denitrification. Water Res. 2018, 143, 355–366. [Google Scholar] [CrossRef]
Water 14 03603 g001 550
Figure 1. Schematic diagram of heterotrophic-autotrophic denitrification reactor.
Figure 1. Schematic diagram of heterotrophic-autotrophic denitrification reactor.
Water 14 03603 g001
Water 14 03603 g002 550
Figure 2. Performance of the system during the different stages: (a) the concentration changes of NO3−-N, NO2−-N, NH4+-N in the effluent, and the removal of TN; (b) the concentration changes of S2− and SO42− in the effluent, and the removal of S2−; (c) the concentration changes of COD in the effluent, and the removal of COD.
Figure 2. Performance of the system during the different stages: (a) the concentration changes of NO3−-N, NO2−-N, NH4+-N in the effluent, and the removal of TN; (b) the concentration changes of S2− and SO42− in the effluent, and the removal of S2−; (c) the concentration changes of COD in the effluent, and the removal of COD.
Water 14 03603 g002
Water 14 03603 g003 550
Figure 3. Changes in EPS contents and PN/PS ratio at the different stages.
Figure 3. Changes in EPS contents and PN/PS ratio at the different stages.
Water 14 03603 g003
Water 14 03603 g004 550
Figure 4. Bacterial community structures at the phylum (a), and at the genus level (b), during the operation.
Figure 4. Bacterial community structures at the phylum (a), and at the genus level (b), during the operation.
Water 14 03603 g004
Water 14 03603 g005 550
Figure 5. Nitrogen metabolic pathways (a) and the functional genes of nitrogen metabolism at different stages (b).
Figure 5. Nitrogen metabolic pathways (a) and the functional genes of nitrogen metabolism at different stages (b).
Water 14 03603 g005
Table
Table 1. Influent conditions of the UASB reactor during different stages.
Table 1. Influent conditions of the UASB reactor during different stages.
StageNitrate-N (mg/L)COD
(mg/L)		S2− (mg/L)TP (mg/L)HRT
(h)		C/NS/N
I (1–30 d)301000543.33-
II (31–60 d)301005543.330.17
III (61–90 d)3010010543.330.33
IV (91–120 d)3010020543.330.67
Table
Table 2. Alpha diversity index at the different stages.
Table 2. Alpha diversity index at the different stages.
SampleAceChao1ShannonSimpsonCoverage
SI2762764.8470.9381.000
SII2602604.2910.9501.000
SIII2302293.5580.9480.999
SIV1981983.2180.9641.000
Publisher’s Note: MDPI stays neutral with regard to jurisdictional claims in published maps and institutional affiliations.

Share and Cite

MDPI and ACS Style

Ren, X.; Wang, Y.; Wan, J.; Yan, Z.; Ma, Y.; Zhang, G.; Zhu, B. The Nitrogen Removal Performance and Functional Bacteria in Heterotrophic Denitrification and Mixotrophic Denitrification Process. Water 2022, 14, 3603. https://doi.org/10.3390/w14223603

AMA Style

Ren X, Wang Y, Wan J, Yan Z, Ma Y, Zhang G, Zhu B. The Nitrogen Removal Performance and Functional Bacteria in Heterotrophic Denitrification and Mixotrophic Denitrification Process. Water. 2022; 14(22):3603. https://doi.org/10.3390/w14223603

Chicago/Turabian Style

Ren, Xinghao, Yan Wang, Jinquan Wan, Zhicheng Yan, Yongwen Ma, Guihua Zhang, and Bin Zhu. 2022. "The Nitrogen Removal Performance and Functional Bacteria in Heterotrophic Denitrification and Mixotrophic Denitrification Process" Water 14, no. 22: 3603. https://doi.org/10.3390/w14223603

APA Style

Ren, X., Wang, Y., Wan, J., Yan, Z., Ma, Y., Zhang, G., & Zhu, B. (2022). The Nitrogen Removal Performance and Functional Bacteria in Heterotrophic Denitrification and Mixotrophic Denitrification Process. Water, 14(22), 3603. https://doi.org/10.3390/w14223603

Note that from the first issue of 2016, this journal uses article numbers instead of page numbers. See further details here.

Article Metrics

Back to TopTop