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Article

Enhancement of Microbial and Metabolic Mechanisms in an Aerobic Bioreactor with Immobilized Microflora by Simple and Complex Electron Donors

1
Shenzhen Key Laboratory of Marine Bioresource & Eco-Environmental Sciences, College of Life Sciences and Oceanography, Shenzhen University, Shenzhen 518071, China
2
College of Physics and Optoelectronic Engineering, Shenzhen University, Shenzhen 518060, China
3
Shenzhen Academy of Environmental Sciences, Shenzhen 518001, China
4
Shenzhen Ecological and Environmental Monitoring Center of Guangdong Province, Shenzhen 518049, China
*
Author to whom correspondence should be addressed.
Water 2023, 15(14), 2548; https://doi.org/10.3390/w15142548
Submission received: 7 June 2023 / Revised: 1 July 2023 / Accepted: 5 July 2023 / Published: 12 July 2023

Abstract

:
Microflora immobilization is promising for nutrient removal applications in sewage; however, the metabolic and microbial mechanism needs to be further explored. Heterotrophic nitrification-aerobic denitrification (HN-AD) bacterium and efficient nitrogen (N) removal bacteria were selected and immobilized on corncob particles using alginate polymer to prepare microbe–organic complex beads. The complex beads were then added into activated sludge under a continuous-flow aerobic bioreactor with sufficient sodium acetate also applied as a simple electron donor. The role of polymer electron donors under carbon-rich conditions was then studied. Results showed that the total nitrogen removal rate improved by 8.3% (reaching 91.2%) and ammonium nitrogen removal rates were approximately 98%. Only 0.59 mg/L of nitrate nitrogen was detected in the treatment group. 16S rRNA gene sequencing results showed that bacterial richness in activated sludge within the treatment group was significantly higher than within the control group (p < 0.05), and KEGG pathways analysis indicated that carbon (C) metabolism gene and N-cycle-related genes were also improved. This suggested that polymer electron donors generated complex C sources that nourished diverse bacterial species related to N cycles so that the N removal rate could be strengthened and further improved by simple electron donors and the microflora.

Graphical Abstract

1. Introduction

Nitrogen (N) removal is one of the most important processes in the biochemical tank of wastewater treatment plants (WWTPs) [1]. In conventional heterotrophic N removal processes, oxygen supply and carbon (C) source are required for nitrification, denitrification, and N assimilation [2,3,4]. The C source not only supports microorganisms to grow but also act as important electron donors that drive nitrate reduction processes and N removal in WWTPs [5]. However, WWTPs face a shortage of C source as the decreasing ratio of chemical oxygen demand (COD) to N in influents and the increasing discharge standards for total nitrogen (TN) [6].
Nowadays, C source shortages are common in the denitrification process in WWTPs, and a supply of dissolved organic carbon such as methanol, acetate, glucose, ethanol, etc. is the priority choice for a solution to this problem [5,7]. Simple organics are highly bioavailable electron donors that can be directly used by denitrifiers. Additionally, liquid phase or soluble electron donors have a high contact area; thus they have the advantage of a short lag period and rapid denitrification [8]. However, some simple electron donors such as methanol are toxic to humans, and even the cheapest results in high operating costs due to the high demand for wastewater treatment [9]. Moreover, the supply of simple electron donors increases the risk of secondary contamination as unused electron donors may remain in the effluent (carbon breakthrough) and require post-treatments such as coagulation, adsorption, and disinfection [5].
As an alternative to simple organics, a variety of cheap biopolymer materials which contain cellulose, hemicellulose, or lignin (e.g., starch, cotton, corncob, wheat straw, rice husk, sawdust, and woodchip) have been developed as alternative electron donors for solid-phase denitrification [10,11,12,13,14,15]. Additionally, corncob is considered a better biopolymer electron donor with a higher N removal rate than straw, rice husks, rice straw, wheat straw, corn stalk, soybean stalk, and soybean hull in different studies [16,17]. Simultaneously, biopolymers are used as supporting material for microbes and biofilm formation due to their porous structure and large surface areas [18]. When corncob is used as an electron donor and biofilm carrier, the reactor startup becomes faster, and almost half of the simple organics get saved [19,20].
It was studied that synthetic polymer and corncob complex showed a better performance in driving denitrification than corncob alone, with a shorter start-up period and no COD pollution [21]. Alginate polymer can also apply as an electron donor and perform similarly to corncob in carbon release and shows excellent denitrification performance [22]. Additionally, alginate has excellent biocompatibility and can be used to immobilize microorganisms and prepare environmentally friendly materials of different forms such as microspheres, microcapsules, sponges, foams, and fibers [23,24]. However, the denitrification rates attained when using a bio-material or synthetic polymer as an electron donor is not comparable to simple electron donors such as methanol, which limits the utilization of polymer electron donors [8,25]. Therefore, the combined usage of simple and natural complex electron donors in wastewater treatment will be promising research, which has not been adequately studied and the mechanisms behind remained unclear.
To achieve a fast startup and to further enhance the N removal, three bacteria species including one HN-AD bacterium on corncob particles were immobilized using alginate polymer. The immobilized bacteria can perform solid-phase denitrification on the carriers and receive extra electron donors from the liquid phase. On the other hand, the immobilized bacteria help break down the complex polymer materials and release the C source to the reactors. HN-AD bacteria have proven very advantageous in N removal rate compared with traditional autotrophic nitrification-heterotrophic denitrification processes since autotrophic nitrifiers are easily affected by the high loads of ammonium and organic matter, which lead to low nitrification rates [26,27]. In the HN-AD process, NH4+-N is oxidized to hydroxylamine and then oxidized to nitrite nitrogen (NO2-N) and nitrate nitrogen (NO3-N), which are denitrified into N2 under the pathway NH4+-N→NH2OH→NO2-N→NO3-N→NO2-N→NO→N2O→N2 [28].
In this study, three strains of N-removal bacteria named Kerstersia sp. S9 (S9), Rhodococcus sp. S2 (S2), and Acinetobacter sp. MZ-5 (MZ-5) were isolated, of which the MZ-5 was able to drive the HN-AD processes. Five dissolved organics (disodium succinate, sodium acetate, trisodium citrate, glucose, and sucrose) were tested to give the most suitable simple electron donor. Corncob particles and calcium alginate were prepared into hydrogel beads and used as polymer electron donors as well as carriers for the bacteria. Two aerobic bioreactor systems were established, and the prepared hydrogel beads were added into the activated sludge in one of the bioreactors as a treatment group and the other as a control group. Sodium acetate was selected as a simple electron donor and added to the simulated wastewater. Thereafter, the N removal performance of the treatment and the control groups were compared and the differences in microbial composition and C and N metabolism were explored to address the following questions:
(1) how did the polymer electron donors interact with the simple electron donor in driving N removal? (2) how did the immobilized microflora improve the system in N removal? and (3) what are the microbial and metabolic mechanisms behind the multi-electron-donor bioreactors?

2. Materials and Methods

2.1. Screening and Isolation of N Removal Microbes (MZ-5, S2, and S9)

The N removal bacteria were isolated from sediments collected from Maozhou River, located in Shenzhen, China. After collection, the sediment samples were immediately stored under the ice and transferred back to the laboratory for screening and isolation of N removal bacteria. The screening and isolation processes followed the procedure described previously by Ouyang [29]. In short, the diluted sediment sludge was pre-incubated for 8 h, and then was added to the enrichment medium for 7 days of enrichment culture. The enrichment medium was then diluted 100 times by sterilized heterotrophic nitrification medium (HNM) and incubated 3 days for the secondary enrichment and repeated 3 times. The enriched bacteria suspension was spread onto HNM plates and incubated 3 days before isolating the intended target bacteria. The isolates were transferred into HNM. Bromothymol blue plates were used in further screening for the aerobic denitrifiers. From each of the plates, each strain isolate marked with blue circles were transferred and conserved at −80 °C in glycerol solution. Three strains capable of effective N removal were obtained and named MZ-5, S2, and S9. Luria-Bertani media was utilized to activate the strains in subsequent experiments.

2.2. Preparation of Polymer Complex Beads Carrying Microflora

Considering that the synthesis conditions of calcium alginate and the embedded amount of bacteria and natural electron donors have important effects on the mass transfer efficiency, mechanical and chemical stability, and final N removal performance of the material, a series of controlled experimental variables were previously performed to determine the optimal conditions for the preparation of calcium alginate based immobilized bacteria for N removal. The optimal synthesis conditions and the number of bacteria and corncob added were then determined. The specific processes involved the activation of the MZ-5, S2, and S9 strains, followed by culturing in the LB medium. The OD600 of each of the three cultured bacteria was then adjusted to 0.85 with pure water. Thereafter, 33 mL of each of the bacterial cultures were taken, mixed, and then centrifuged under conditions of 4 °C and 8000 rpm. The supernatant was discarded, the pellet was washed three times with sterilized normal saline, then 5 mL of normal saline was added and resuspended by vortexing. Next, 15 mL of 4% sodium alginate solution and 0.05 g of 30 mesh corncob particles were added into the resuspension and then vortexed to mix evenly. The mixture was then added in a dropwise manner to 3% CaCl2 solution and stirred on a magnetic stirrer at a speed of 200 r/min for 12 h. The CaCl2 solution was discarded, and the obtained hydrogel beads were rinsed three times with normal saline. The products were stored at 4 °C in the refrigerator for further analysis.

2.3. Assessment of N Removal Potential of the Microflora Using Different Electron Donors

The three strains activated by the LB medium were inoculated into multi-N source medium at a rate of 2%, cultured at 25 °C and 150 rpm, and sampled at 0 h, 4 h, 8 h, 12 h, 24 h, and 48 h. Samples were centrifuged at 8000 rpm for 10 min and the supernatants were collected to determine the concentrations of TN, NH4+-N, NO3-N, and NO2-N.
The effect of different types of electron donors on N removal by the microflora (MZ-5, S2, and S9) was investigated. Based on the C content, the sodium succinate in the multi-N source medium was replaced by sodium citrate, sodium acetate, glucose, and sucrose, respectively. Three activated bacterial cultures were co-inoculated into the above medium each with 2% inoculation amount, then cultured under 25 °C and 150 rpm, and followed by sampling at 0 h, 4 h, 8 h, 12 h, 24 h, and 48 h. Samples were centrifuged at 8000 rpm for 10 min and the supernatants were collected and used to detect the concentrations of TN, NH4+-N, NO3-N, and NO2-N. The multi-N medium had a composition of 0.10 g/L (NH4)2SO4, 0.10 g/L NaNO2, 0.15 g/L KNO3, 0.86 g/L disodium succinate, 0.25 g/L K2HPO4, 0.125 g/L NaCl, 2.5 g/L MgSO4 7H2O, 0.05 g/L MnSO4 4H2O and 0.05 g/L FeSO4 7H2O.

2.4. Establishment of the Aerobic Reactor Device and Operation Conditions

As shown in Figure 1, the main body of the reactor consisted of a 5 L acrylic container connected to a sharp-bottomed overflow vessel. The aeration strip at the bottom of the reactor was connected to a 24-h blower pump (Figure 1). The inlet and outlet water as well as the sludge return were controlled by peristaltic pumps at a constant continuous speed (Figure 1). Two systems were then designed to serve as control and treatment groups 5 L of the sludge collected from the biochemical tank of the wastewater treatment plant was added to the reactors. The mixed liquid suspended solids (MLSS) and the sludge settling velocity (SV) were 4900 mg/L and 31%, respectively. Oxygenation of the water was maintained by the blower pump, and the gas flow rate was controlled at 3 L/min. The sludge retention time (SRT) was set at 30 days to regularly discharge 200 mL of sludge every day. By adjusting the speed of the peristaltic pump, the hydraulic retention time (HRT) was set at 16 h in stage 1 and adjusted to 8 h in stage 2. Within stage 1, the simulated wastewater underwent full reaction for 7 days with a relatively longer HRT in the control and treatment reactors to domesticate and stabilize the bacterial community in the system. In stage 2, the inflow rate of the wastewater was doubled, and the prepared microflora-carrying polymer complex beads (30 g) were added to the treatment group on day 7, while the other system was left as the control group. The effluent and influent of the treatment and control groups were collected daily for the follow-up testing. The formula composition of the simulated wastewater is provided in Table 1.

2.5. Metagenomic DNA Extraction and 16S rRNA Sequencing

For metagenomic DNA extraction and 16S rRNA sequencing, 50 mL of activated sludge was sampled from the control and treatment groups on the first and seventh days of adding immobilized microflora beads to the system. Samples were named AS1, BS1, AS7, and BS7, and each was taken in triplicate. Supernatants were discarded after centrifuging the sludge samples. The genomes of obtained sludge samples were extracted using the EZNA DNA kit (OMEGA Bio-Tek, Norcross, GA, USA). The bacterial 16S rRNA V3-V4 hypervariable region was then amplified with primers 341F (5’-CCTAYGGGRBGCASCAG-3’) and 806R (5’-GGACTACNNGGGTATCTAAT-3’). The PCR reactions and the platform used were described by [30].

2.6. Analytical Methods

A microplate spectrophotometer (Epoch 2, BioTeK, Winooski, VT, USA) was used to detect the growth of bacteria at 600 nm. Samples were centrifuged to obtain supernatants for water chemical analyses, and measurements of the concentrations of NH4+-N, NO3-N, and NO2-N and TN followed the international standard method [31]. Genetic, metabolic, and functional analysis were mainly based on the Kyoto Encyclopedia of Genes and Genomes (KEGG) database.

3. Results and Discussions

3.1. Selection of Simple Electron Donor for the Microflora

The MZ-5, S2, and S9 can remove various forms of N, including NH4+-N, NO3-N, and NO2-N. The MZ-5 was an HN-AD bacterium capable of effectively treating landfill leachate wastewater [29], and the NH4+-N removal rate by the MZ-5 reached 98.97 ± 0.09% under the culture medium in which disodium succinate was used as an electron donor (Table 2). The S2 and S9 bacteria were also capable of removing NH4+-N up to levels of 99.26 ± 0.13% and 99.15 ± 0.03% respectively. The NO3-N removal rate was higher in the S2 culture medium and reached 93.06 ± 0.16%, while NO2-N removal rate was only 7.74 ± 0.17% (Table 2). However, the TN removal rate of the three bacteria did not exceed 70% when using disodium succinate as a C source (Table 2). It is important to choose an electron donor in a bioreactor system, as the type of electron donor affected the metabolisms of microbial products and denitrification in a bioreactor [32]. Additionally, co-culture of different HN-AD bacteria had been recorded to increase the NO3-N and NO2-N removal rates up to 99.85% and 96.94%, respectively [33]. Also, fungi and bacteria with poor denitrification ability can achieve a 95.78% NO3-N removal rate when co-cultured [34]. Therefore, to increase the TN removal rate, a co-cultured microflora containing the three efficient N removal bacteria (MZ-5, S2, and S9) was designed and the N removal performance under the different simple electron donors was investigated.
Five simple organics were chosen, which showed that trisodium citrate was not a suitable electron donor for the microflora, with almost no or very little N being removed (Table 3), which was most likely because of the lack of citrate transporter in the bacteria [35]. A small amount of NO3-N accumulation was found in the medium when saccharides including glucose and sucrose were used as C sources (Table 3). The glucose feeding might have produced more soluble microbial products that were detrimental to functional bacteria, thus inhibiting the denitrification performance [32]. Sodium acetate was the most suitable electron donor that drove the microflora growth, triggered the denitrification process, and reached 98.58 ± 0.32% in NO3-N removal rate (Table 3).

3.2. Effects of Multi-Electron Donors on the Performance of Aerobic Bioreactors

A series of research has shown that corncob is a great biopolymer material in supporting denitrification and N removal as compared to straw, rice husk, peanut shell, obsolescent rice, etc. due to its higher C release ability [17,22]. Therefore, corncob was chosen as a polymer electron donor source to supplement sodium acetate in the aerobic bioreactors. The alginate, applied to immobilize the microflora and the corncob particles, was also serving as a polymer electron donor. Results showed that the TN removal rate increased within two days after the addition of the microflora-carrying polymer complex beads. TN removal rate reached the peak (91.2%) in four days, which increased by an average of 8.3% compared to the control group. In addition, the performance of the treatment group did not show any significant downward trend during the experiment (Figure 2). NH4+-N was the main form of N in the system, which had a high removal rate through assimilation and the coupled nitrification-denitrification process (Figure 2). It had been reported that NH4+-N has preferential uptake by microorganisms and green algae [29,36]. However, in this study, the NH4+-N removal rate started to decrease in both the control and the treatment groups after the 13th day of operation (Figure 2). This might be related to the fact that the HRT of activated sludge in the system was too long, which led to the retention of excess sludge in the systems.
NO2-N did not accumulate in both the control and the treatment groups during the whole operation (Figure 3B). At stage 1, NO3-N concentrations in the effluent of the control and the treatment group were stable and 1.16 mg/L on average (Figure 3B) due to the relatively long HRT. Like most aeration basins of WWTPs, the NO3-N in the effluent of the control group gradually accumulated after the shortening of the HRT within stage 2. Whereas NO3-N in the effluent of the treatment group started to decrease to an average of 0.59 mg/L (Figure 3B) and reached the first-class standard which is 1 mg/L for NO3-N in the Surface Water Environmental Quality Standard (GB3838-2002) of China [33]. The increased NO3-N in the control group was mainly generated from nitrification or residue of the denitrification. The adjustment of HRT at stage 2 might have caused insufficient denitrification and led to the accumulation of NO3-N in the control group. Additionally, the acceleration of inlet water input might have also increased the nitrification rate due to more NH4+-N flowing into the system. Therefore, the limitation of denitrification rate was one of the important reasons for the N removal (especially the NO3-N) within the systems. Hence, the addition of microflora-carrying polymer complex beads was shown to be effective in strengthening and maintaining the removal rate of TN, NH4+-N, and NO3-N in an aerobic bioreactor.

3.3. The Relationships between the Microbial Diversity and Electron Donor

The optimal performance and functional stability of WWTPs greatly depend on the microbial community and composition of activated sludge [37]. The 16S rRNA gene sequencing showed that microbial diversity of activated sludge increased in both the control and the treatment groups during stage 2 (Figure 4). However, there was no significant difference in the index of observed species in the control group during stage 2 (Figure 5). Therefore, in the control group, simple electron donors promoted a minority of the microbes that were already activated in the domestication period of stage 1 and improved the microbial evenness, which enhanced the overall microbial diversity. In the treatment group, the observed species index was significantly increased, and microbial diversity was greatly improved (Figure 5). The addition of the polymer electron donors improved the microbial richness of the treatment group. In field studies of denitrification beds, polymer electron donor also determines the microbial community diversity [38]. Although microbial diversity does not reflect the functions of the ecosystems directly, shifts in functionality can well be explained by the changes in microbial community structure [39]. It could be inferred from Figure 3A that the quantity of released COD from the polymer complex beads was negligible as compared to the sodium acetate. However, enzymatic hydrolysis of the polymers produced various kinds of organics such as organic acids and sugars, which greatly enriched the types of electron donors in the system [21,40], which might decrease the threshold for the life of the microorganisms. Therefore, more species could have existed due to the condition created by the main simple electron donor in the treatment group.

3.4. Role of Non-Dominant Bacteria and The Characteristic Genus of the Treatment Groups

The study indicated that in the treatment group, the number of observed species significantly increased, and the minority species were observed to thrive more (Figure 4 and Figure 5). It should be noted that Aquaspirillum in the AS1 and AS7 might be due to the contamination during the initial cleaning of the system or the sequencing error when using the Illumina platform. In the control group of AS7, the relative abundance of dominant species decreased, while almost all less dominant specie increased in the treatment group of BS7 (Table S5). This greatly enhanced the microbial diversity and richness of the activated sludge, which implied that the relative abundance of dominant species might not have an impact on the final N removal improvement in this case. It was reported that non-dominant flanking bacteria affected the N metabolisms within the denitrifying enhanced biological phosphorus removal bioreactors [41]. By avoiding the number-increased species in Table S2, four species were chosen as the characteristic genera of the treatment group from Table S5 as their relative abundance was much higher in BS7 than in AS7. Specifically, Pedobacter, Sphingopyxis, Fusibacter, and Hyphomonas were the characteristic genus of the treatment group that resulted from the addition of the polymer electron donors. Pedobacter, Fusibacter, and Hyphomonas had been reported to carry denitrification functions in specific species [42,43,44], and Sphingopyxis was mostly utilized as a degrader of organic pollutants [45,46,47]. These results suggested that the C and N metabolism genes might have interacted in the treatment group.

3.5. C Metabolism and N Cycle Genes of Activated Sludge

It has been reported that C metabolisms interact with denitrification and affect N removal in various systems [13,48]. The up-regulated C metabolism allows the system to generate more energy, thereby promoting the HN-AD-related N removal pathways [49]. On the other hand, the modulated C metabolism optimized electron distribution among denitrifying enzymes by adsorbing the extracellular metabolites [50]. In this study, carbohydrate, lipid as well as amino acid metabolism genes were increased in both the control and treatment groups in stage 2, which indicated the promotion of organic C utilization efficiency, and these three representative C metabolism genes were higher in BS7 than in AS7 (Figure 6, Figures S1 and S2). The results showed that the combined use of simple and complex electron donors had a stronger contribution in shaping microbial communities and improving the C metabolic potential than simple electron donors alone.
From the perspective of N metabolisms, the nirK, norB, and nosZ were key genes related to the generation of gaseous N and N oxides (N2, N2O, and NO), which played a leading role in N removal functions in the denitrification processes. Indeed, the results from this study showed that the OTU numbers of nirK, norB, and nosZ were significantly higher in BS7 than in AS7 (p < 0.05) (Figure 7). In addition to the denitrification process, the nasA and nirB genes were also significantly higher in the treatment group than in the control group during the stage 2 (p < 0.05) (Figure 7), which regulated assimilatory nitrate reduction processes that transformed nitrate into intercellular organic N and removed NO3-N from the water phase. However, genes related to nitrification did not change during the experiment (Figure 7). Thereby, the improved removal of NO3-N was driven by different processes of N cycles, which greatly contributed to the final TN removal. However, the differences in NO3-N concentrations in the effluents between the treatment and the control groups could not fully explain the improved TN removal, which suggested that the assimilation of ammonium and organic N might have also been improved in the treatment group by the improvement of microbial richness. Many studies have proved that the use of complex electron donors to facilitate N removal rates is restricted. While in our study, the complex electron donors enriched the variety of organics and provided the ability to continuously supply the organics to ensure that microorganisms can obtain abundant organics uninterruptedly, thus increasing the microbial richness and metabolic activity of the microorganisms. Therefore, the utilization efficiency of simple organics was improved in the system. As a result, the cost of the wastewater treatment system can be reduced, and the system performance efficiency can be improved.

4. Conclusions

The combined use of simple and polymer electron donors significantly increased the observed microbial species and thus improved the microbial richness. The enhanced C metabolic potential in the treatment group indicated that the polymer complex beads generated various forms of organics which supported diverse species and re-shaped the microbial community. The N cycle genes related to denitrification and assimilatory nitrate reduction were significantly upregulated, which contributed to the improved N removal rates in the treatment group. With the help of the immobilized microflora, TN and NO3-N removal rate was further increased in the aerobic bioreactor. Therefore, this study proposes the use of polymers as microbial carriers and the complex organics supply sources, in combination with simple electron donors, rather than separate applications for improved microbial and metabolic mechanisms in aerobic bioreactors.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w15142548/s1, Figure S1: Heatmap showing clustering among different samples based on relative abundance of lipid metabolisms from KEGG database; Figure S2: Heatmap showing clustering among different samples based on relative abundance of amino acid metabolisms from KEGG database; Figure S3: Stacked column chart of relative abundance in the activated sludges at the order level; Figure S4: PCoA analysis of different sample groups based on weighted unifrac; Table S1: Alpha Diversity Statistics of different samples; Table S2: Microbial difference between AS1 group and AS7 group in genus level; Table S3: Microbial difference between BS1 group and BS7 group in genus level; Table S4: Microbial difference between AS1 group and BS1 group in genus level; Table S5: Microbial difference between AS7 group and BS7 group in genus level.

Author Contributions

Conceptualization, Q.D. and H.C.; data curation, S.L.; formal analysis, Q.D.; funding acquisition, S.L.; investigation, K.W.; methodology, Q.D., K.W. and W.X.; project administration, S.L.; resources, S.L.; Software, X.Y.; supervision, H.C.; validation, J.F.; visualization, Q.D.; writing—original draft, Q.D.; writing—review and editing, S.L. and H.C. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by [National Key Research and Development Program of China] grant number [2020YFD0901003] and [Shenzhen Science and Technology Program] grant number [KCXFZ20201221173404012] & [KCXST20221021111206015].

Data Availability Statement

The data that support the findings of this study are available from the corresponding author upon reasonable request.

Conflicts of Interest

The authors have no relevant financial or non-financial interests to disclose.

References

  1. Wang, X.; Xia, S.; Chen, L.; Zhao, J.; Renault, N.; Chovelon, J. Nutrients removal from municipal wastewater by chemical precipitation in a moving bed biofilm reactor. Process. Biochem. 2006, 41, 824–828. [Google Scholar] [CrossRef]
  2. Kuai, L.P.; Verstraete, W. Ammonium removal by the oxygen-limited autotrophic nitrification-denitrification system. Appl. Environ. Microb. 1998, 64, 4500–4506. [Google Scholar] [CrossRef] [PubMed] [Green Version]
  3. Yang, Q.; Peng, Y.; Liu, X.; Zeng, W.; Mino, T.; Satoh, H. Nitrogen Removal via Nitrite from Municipal Wastewater at Low Temperatures Using Real-Time Control to Optimize Nitrifying Communities. Environ. Sci. Technol. 2007, 41, 8159–8164. [Google Scholar] [CrossRef] [PubMed]
  4. Zheng, K.; Li, H.; Wang, S.; Wang, Y.; Li, A.; Feng, X.; Li, J. Enhanced proteins and amino acids production based on ammonia nitrogen assimilation and sludge increment by the integration of bioadsorption with anaerobic-anoxic-oxic (AAO) process. Chemosphere 2021, 280, 130721. [Google Scholar] [CrossRef]
  5. Park, J.Y.; Yoo, Y.J. Biological nitrate removal in industrial wastewater treatment: Which electron donor we can choose. Appl. Microbiol. Biotechnol. 2009, 82, 415–429. [Google Scholar] [CrossRef]
  6. Isaacs, S.; Henze, M.; Søeberg, H.; Kümmel, M. External carbon source addition as a means to control an activated sludge nutrient removal process. Water Res. 1994, 28, 511–520. [Google Scholar] [CrossRef]
  7. Hu, W.; Tian, J.; Li, X.; Chen, L. Wastewater treatment system optimization with an industrial symbiosis model: A case study of a Chinese eco-industrial park. J. Ind. Ecol. 2020, 24, 1338–1351. [Google Scholar] [CrossRef]
  8. Pang, Y.; Wang, J. Various electron donors for biological nitrate removal: A review. Sci. Total Environ. 2021, 794, 148699. [Google Scholar] [CrossRef]
  9. Castellar, J.; Formosa, J.; Fernandez, A.I.; Jové, P.; Bosch, M.G.; Morató, J.; Brix, H.; Arias, C.A. Cork as a sustainable carbon source for nature-based solutions treating hydroponic wastewaters—Preliminary batch studies. Sci. Total Environ. 2019, 650, 267–276. [Google Scholar] [CrossRef]
  10. Fowdar, H.S.; Hatt, B.E.; Breen, P.; Cook, P.L.; Deletic, A. Evaluation of sustainable electron donors for nitrate removal in different water media. Water Res. 2015, 85, 487–496. [Google Scholar] [CrossRef]
  11. Hu, R.; Zheng, X.; Xin, J.; Sun, Z.; Zheng, T. Selective enhancement and verification of woody biomass digestibility as a denitrification carbon source. Bioresour. Technol. 2017, 244, 313–319. [Google Scholar] [CrossRef] [PubMed]
  12. Si, Z.; Song, X.; Wang, Y.; Cao, X.; Zhao, Y.; Wang, B.; Chen, Y.; Arefe, A. Intensified heterotrophic denitrification in constructed wetlands using four solid carbon sources: Denitrification efficiency and bacterial community structure. Bioresour. Technol. 2018, 267, 416–425. [Google Scholar] [CrossRef]
  13. Deng, Q.; Wan, L.; Li, X.; Cao, X.; Zhou, Y.; Song, C. Metagenomic evidence reveals denitrifying community diversity rather than abundance drives nitrate removal in stormwater biofilters amended with different organic and inorganic electron donors. Chemosphere 2020, 257, 127269. [Google Scholar] [CrossRef] [PubMed]
  14. Sun, F.; Deng, Q.; Li, X.; Tang, M.; Ma, X.; Cao, X.; Zhou, Y.; Song, C. Organic carbon quantity and quality jointly triggered the switch between dissimilatory nitrate reduction to ammonium (DNRA) and denitrification in biofilters. Chemosphere 2021, 280, 130917. [Google Scholar] [CrossRef] [PubMed]
  15. Zhang, F.; Ma, C.; Huang, X.; Liu, J.; Lu, L.; Peng, K.; Li, S. Research progress in solid carbon source–based denitrification technologies for different target water bodies. Sci. Total Environ. 2021, 782, 146669. [Google Scholar] [CrossRef]
  16. Ling, Y.; Yan, G.; Wang, H.; Dong, W.; Wang, H.; Chang, Y.; Chang, M.; Li, C. Release Mechanism, Secondary Pollutants and Denitrification Performance Comparison of Six Kinds of Agricultural Wastes as Solid Carbon Sources for Nitrate Removal. Int. J. Environ. Res. Public Health 2021, 18, 1232. [Google Scholar] [CrossRef]
  17. Tao, M.; Jing, Z.; Tao, Z.; Luo, H.; Zuo, S. Improvements of nitrogen removal and electricity generation in microbial fuel cell-constructed wetland with extra corncob for carbon-limited wastewater treatment. J. Clean. Prod. 2021, 297, 126639. [Google Scholar] [CrossRef]
  18. Zhong, H.; Cheng, Y.; Ahmad, Z.; Shao, Y.; Zhang, H.; Lu, Q.; Shim, H. Solid-phase denitrification for water remediation: Processes, limitations, and new aspects. Crit. Rev. Biotechnol. 2020, 40, 1113–1130. [Google Scholar] [CrossRef] [PubMed]
  19. Xu, Z.-X.; Shao, L.; Yin, H.-L.; Chu, H.-Q.; Yao, Y.-J. Biological Denitrification Using Corncobs as a Carbon Source and Biofilm Carrier. Water Environ. Res. 2009, 81, 242–247. [Google Scholar] [CrossRef]
  20. Li, G.; Chen, J.; Yang, T.; Sun, J.; Yu, S. Denitrification with corncob as carbon source and biofilm carriers. Water Sci. Technol. 2012, 65, 1238–1243. [Google Scholar] [CrossRef]
  21. Li, C.; Wang, H.; Yan, G.; Dong, W.; Chu, Z.; Wang, H.; Chang, Y.; Ling, Y.; Zhang, Y. Initial carbon release characteristics, mechanisms and denitrification performance of a novel slow release carbon source. J. Environ. Sci. 2022, 118, 32–45. [Google Scholar] [CrossRef] [PubMed]
  22. Xiong, R.; Yu, X.; Zhang, Y.; Peng, Z.; Yu, L.; Cheng, L.; Li, T. Comparison of agricultural wastes and synthetic macromolecules as solid carbon source in treating low carbon nitrogen wastewater. Sci. Total Environ. 2020, 739, 139885. [Google Scholar] [CrossRef] [PubMed]
  23. Sun, J.; Tan, H. Alginate-Based Biomaterials for Regenerative Medicine Applications. Materials 2013, 6, 1285–1309. [Google Scholar] [CrossRef] [PubMed]
  24. Chang, Q.; Ali, A.; Su, J.; Wen, Q.; Bai, Y.; Gao, Z.; Xiong, R. Efficient removal of nitrate, manganese, and tetracycline by a polyvinyl alcohol/sodium alginate with sponge cube immobilized bioreactor. Bioresour. Technol. 2021, 331, 125065. [Google Scholar] [CrossRef]
  25. Warneke, S.; Schipper, L.A.; Matiasek, M.G.; Scow, K.M.; Cameron, S.; Bruesewitz, D.A.; McDonald, I.R. Nitrate removal, communities of denitrifiers and adverse effects in different carbon substrates for use in denitrification beds. Water Res. 2011, 45, 5463–5475. [Google Scholar] [CrossRef] [Green Version]
  26. Khin, T.; Annachhatre, A.P. Novel microbial nitrogen removal processes. Biotechnol. Adv. 2004, 22, 519–532. [Google Scholar] [CrossRef]
  27. 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]
  28. Xu, N.; Liao, M.; Liang, Y.; Guo, J.; Zhang, Y.; Xie, X.; Fan, Q.; Zhu, Y. Biological nitrogen removal capability and pathways analysis of a novel low C/N ratio heterotrophic nitrifying and aerobic denitrifying bacterium (Bacillus thuringiensis strain WXN-23). Environ. Res. 2021, 195, 110797. [Google Scholar] [CrossRef]
  29. Ouyang, L.; Wang, K.; Liu, X.; Wong, M.H.; Hu, Z.; Chen, H.; Yang, X.; Li, S. A study on the nitrogen removal efficacy of bacterium Acinetobacter tandoii MZ-5 from a contaminated river of Shenzhen, Guangdong Province, China. Bioresour. Technol. 2020, 315, 123888. [Google Scholar] [CrossRef]
  30. Xie, N.; Zhong, L.; Ouyang, L.; Xu, W.; Zeng, Q.; Wang, K.; Zaynab, M.; Chen, H.; Xu, F.; Li, S. Community Composition and Function of Bacteria in Activated Sludge of Municipal Wastewater Treatment Plants. Water 2021, 13, 852. [Google Scholar] [CrossRef]
  31. APHA. Standard Methods for the Examination of Water and Wastewater, 22nd ed.; American Public Health Association (APHA): Washington, DC, USA, 2012. [Google Scholar]
  32. Jiang, F.; Qi, Y.; Shi, X. Effect of liquid carbon sources on nitrate removal, characteristics of soluble microbial products and microbial community in denitrification biofilters. J. Clean. Prod. 2022, 339, 130776. [Google Scholar] [CrossRef]
  33. Zhang, Y.; Xu, Z.; Li, J.; Liu, D.; Yuan, Y.; Chen, Z.; Wang, G. Cooperation between two strains of Enterobacter and Klebsiella in the simultaneous nitrogen removal and phosphate accumulation processes. Bioresour. Technol. 2019, 291, 121854. [Google Scholar] [CrossRef] [PubMed]
  34. Chen, C.; Wang, Z.; Zhao, M.; Yuan, B.; Yao, J.; Chen, J.; Hrynshpan, D.; Savitskaya, T. A fungus–bacterium co-culture synergistically promoted nitrogen removal by enhancing enzyme activity and electron transfer. Sci. Total Environ. 2021, 754, 142109. [Google Scholar] [CrossRef] [PubMed]
  35. Van Hofwegen, D.J.; Hovde, C.J.; Minnich, S.A. Rapid Evolution of Citrate Utilization by Escherichia coli by Direct Selection Requires citT and dctA. J. Bacteriol. 2016, 198, 1022–1034. [Google Scholar] [CrossRef] [PubMed] [Green Version]
  36. Xiu, B.; Liang, S.-K.; He, X.-L.; Wang, X.-K.; Cui, Z.-G.; Jiang, Z.-J. Bioavailability of dissolved organic nitrogen and its uptake by Ulva prolifera: Implications in the outbreak of a green bloom off the coast of Qingdao, China. Mar. Pollut. Bull. 2019, 140, 563–572. [Google Scholar] [CrossRef]
  37. Sun, C.; Zhang, B.; Ning, D.; Zhang, Y.; Dai, T.; Wu, L.; Li, T.; Liu, W.; Zhou, J.; Wen, X. Seasonal dynamics of the microbial community in two full-scale wastewater treatment plants: Diversity, composition, phylogenetic group based assembly and co-occurrence pattern. Water Res. 2021, 200, 117295. [Google Scholar] [CrossRef]
  38. Grießmeier, V.; Gescher, J. Influence of the Potential Carbon Sources for Field Denitrification Beds on Their Microbial Diversity and the Fate of Carbon and Nitrate. Front. Microbiol. 2018, 9, 1313. [Google Scholar] [CrossRef]
  39. Zhou, Z.; Wang, C.; Luo, Y. Meta-analysis of the impacts of global change factors on soil microbial diversity and functionality. Nat. Commun. 2020, 11, 3072. [Google Scholar] [CrossRef]
  40. Gong, X.; Li, Q.; Li, T.; Li, C.; Huang, J.; Zhou, N.; Jia, X. Chemical composition and monolignin in alkali and acid treated corncob affect sugar release. Ind. Crop. Prod. 2022, 176, 114317. [Google Scholar] [CrossRef]
  41. Gao, H.; Mao, Y.; Zhao, X.; Liu, W.-T.; Zhang, T.; Wells, G. Genome-centric metagenomics resolves microbial diversity and prevalent truncated denitrification pathways in a denitrifying PAO-enriched bioprocess. Water Res. 2019, 155, 275–287. [Google Scholar] [CrossRef]
  42. Fesefeldt, A.; Kloos, K.; Bothe, H.; Lemmer, H.; Gliesche, C. Distribution of denitrification and nitrogen fixation genes in Hyphomicrobium spp. and other budding bacteria. Can. J. Microbiol. 1998, 44, 181–186. [Google Scholar] [CrossRef]
  43. Harter, J.; Weigold, P.; El-Hadidi, M.; Huson, D.H.; Kappler, A.; Behrens, S. Soil biochar amendment shapes the composition of N2O-reducing microbial communities. Sci. Total Environ. 2016, 562, 379–390. [Google Scholar] [CrossRef] [PubMed]
  44. Han, F.; Li, X.; Zhang, M.; Liu, Z.; Han, Y.; Li, Q.; Zhou, W. Solid-phase denitrification in high salinity and low-temperature wastewater treatment. Bioresour. Technol. 2021, 341, 125801. [Google Scholar] [CrossRef]
  45. Aranda, C.; Godoy, F.; Becerra, J.; Barra, R.O.; Martinez, M. Aerobic secondary utilization of a non-growth and inhibitory substrate 2,4,6-trichlorophenol by Sphingopyxis chilensis S37 and Sphingopyxis-like strain S32. Biodegradation 2003, 14, 265–274. [Google Scholar] [CrossRef] [PubMed]
  46. Hu, X.; Mamoto, R.; Tang, M.; Kawai, F. Polyvinyl alcohol degradation by Sphingopyxis sp. strain 113P3. J. Biotechnol. 2008, 136, S308. [Google Scholar] [CrossRef]
  47. Kaminski, M.A.; Sobczak, A.; Dziembowski, A.; Lipinski, L. Genomic Analysis of γ-Hexachlorocyclohexane-Degrading Sphingopyxis lindanitolerans WS5A3p Strain in the Context of the Pangenome of Sphingopyxis. Genes 2019, 10, 688. [Google Scholar] [CrossRef] [Green Version]
  48. Huang, X.; Duan, C.; Yu, J.; Dong, W. Transforming heterotrophic to autotrophic denitrification process: Insights into microbial community, interspecific interaction and nitrogen metabolism. Bioresour. Technol. 2022, 345, 126471. [Google Scholar] [CrossRef]
  49. Lin, Z.; Zhou, J.; He, L.; He, X.; Pan, Z.; Wang, Y.; He, Q. High-temperature biofilm system based on heterotrophic nitrification and aerobic denitrification treating high-strength ammonia wastewater: Nitrogen removal performances and temperature-regulated metabolic pathways. Bioresour. Technol. 2022, 344, 126184. [Google Scholar] [CrossRef]
  50. Zhang, Y.; Zhang, Z.; Chen, Y. Biochar Mitigates N2O Emission of Microbial Denitrification through Modulating Carbon Metabolism and Allocation of Reducing Power. Environ. Sci. Technol. 2021, 55, 8068–8078. [Google Scholar] [CrossRef]
Figure 1. Design and connection of the aerobic bioreactor devices.
Figure 1. Design and connection of the aerobic bioreactor devices.
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Figure 2. Changes in concentrations and removal rates of NH4+-N and TN within the reactors. Stage 1 represented the domestication period and stage 2 was the treatment period. The arrows showed the moment adding the microflora-carrying polymer complex beads.
Figure 2. Changes in concentrations and removal rates of NH4+-N and TN within the reactors. Stage 1 represented the domestication period and stage 2 was the treatment period. The arrows showed the moment adding the microflora-carrying polymer complex beads.
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Figure 3. Changes in concentrations and removal rates of NH4+-N (A) and TN (B) within the reactors. Stage 1 represented the domestication period and stage 2 was the treatment period. The arrows showed the moment adding the microflora-carrying polymer complex beads.
Figure 3. Changes in concentrations and removal rates of NH4+-N (A) and TN (B) within the reactors. Stage 1 represented the domestication period and stage 2 was the treatment period. The arrows showed the moment adding the microflora-carrying polymer complex beads.
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Figure 4. Stacked column chart of relative abundance of top 35 species in the activated sludges at the genus level. AS1 (control samples taken on the first day after the treatment group added the materials), AS7 (control samples taken on the seventh day after the treatment group added the materials), BS1 (treat samples taken on the first day after the treatment group added the materials), and BS7 (treat samples taken on the seventh day after the treatment group added the materials). ‘Others’ means unclassified.
Figure 4. Stacked column chart of relative abundance of top 35 species in the activated sludges at the genus level. AS1 (control samples taken on the first day after the treatment group added the materials), AS7 (control samples taken on the seventh day after the treatment group added the materials), BS1 (treat samples taken on the first day after the treatment group added the materials), and BS7 (treat samples taken on the seventh day after the treatment group added the materials). ‘Others’ means unclassified.
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Figure 5. Box plots of Alpha diversity index in the sample groups.
Figure 5. Box plots of Alpha diversity index in the sample groups.
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Figure 6. Heatmap showing clustering among different samples based on relative gene abundance of lipid metabolism from KEGG database.
Figure 6. Heatmap showing clustering among different samples based on relative gene abundance of lipid metabolism from KEGG database.
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Figure 7. Differences in N cycle-related genes between treatment and control group activated sludge samples. The asterisks of different colors represent different treatment groups. Lowercase letters indicated significant differences exist between the corresponding groups.
Figure 7. Differences in N cycle-related genes between treatment and control group activated sludge samples. The asterisks of different colors represent different treatment groups. Lowercase letters indicated significant differences exist between the corresponding groups.
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Table 1. The composition of the simulated wastewater.
Table 1. The composition of the simulated wastewater.
ComponentConcentration (g/L)ComponentConcentration (g/L)
CH3COONa3.13MgSO4·7H2O0.05
(NH4)2SO40.24MnSO4·4H2O0.01
K2HPO40.25FeSO4·7H2O0.01
NaCl0.125C:N18
Table 2. N removal rates by MZ-5, S2, and S9 using disodium succinate as C source.
Table 2. N removal rates by MZ-5, S2, and S9 using disodium succinate as C source.
StrainsNH4+-NNO2-NNO3-NTN
MZ-598.97 ± 0.09%67.44 ± 0.12%65.21 ± 0.12%68.52 ± 0.26%
S299.26 ± 0.13%7.74 ± 0.17%93.06 ± 0.16%69.69 ± 0.09%
S999.15 ± 0.03%17.51 ± 0.22%55.10 ± 0.15%60.05 ± 0.09%
Table 3. The N removal rates by microflora under five simple electron donors.
Table 3. The N removal rates by microflora under five simple electron donors.
Carbon SourceNH4+-NNO2-NNO3-NTN
Disodium succinate99.58 ± 0.23%4.10 ± 0.08%81.14 ± 0.12%65.91 ± 0.22%
Sodium acetate99.34 ± 0.31%46.06 ± 0.15%98.58 ± 0.32%84.64 ± 0.16%
Trisodium citrate3.51 ± 0.11%−4.15 ± 0.26%2.14 ± 0.12%10.24 ± 0.17%
Glucose56.41 ± 0.21%3.60 ± 0.23%−4.79 ± 0.22%7.89 ± 0.13%
Sucrose41.84 ± 0.27%4.18 ± 0.14%−9.43 ± 0.26%8.47 ± 0.22%
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Deng, Q.; Wang, K.; Xu, W.; Yu, X.; Feng, J.; Li, S.; Chen, H. Enhancement of Microbial and Metabolic Mechanisms in an Aerobic Bioreactor with Immobilized Microflora by Simple and Complex Electron Donors. Water 2023, 15, 2548. https://doi.org/10.3390/w15142548

AMA Style

Deng Q, Wang K, Xu W, Yu X, Feng J, Li S, Chen H. Enhancement of Microbial and Metabolic Mechanisms in an Aerobic Bioreactor with Immobilized Microflora by Simple and Complex Electron Donors. Water. 2023; 15(14):2548. https://doi.org/10.3390/w15142548

Chicago/Turabian Style

Deng, Qinghui, Keju Wang, Wang Xu, Xinfan Yu, Jie Feng, Shuangfei Li, and Huirong Chen. 2023. "Enhancement of Microbial and Metabolic Mechanisms in an Aerobic Bioreactor with Immobilized Microflora by Simple and Complex Electron Donors" Water 15, no. 14: 2548. https://doi.org/10.3390/w15142548

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