Screening and Application of High-Efficiency Ammonia Nitrogen Degrading Bacteria

Abstract

There is a lack of research on screening new strains of high-efficiency ammonia nitrogen degrading bacteria and treating high-concentration ammonia nitrogen aquaculture wastewater using immobilized composite bacteria. In this study, two strains capable of degrading ammonia nitrogen and nitrite were isolated from surface water. The species of the strains were accurately identified using ITS sequencing technology. Scp1 was identified as Pseudomonas and Scr1 as Rhodococcus erythropolis. Both strains were preserved. When the initial concentration of ammonia nitrogen was 1.50 mg/L, the degradation efficiency of ammonia nitrogen after 4 days of inoculation with Scp1, Scr1, and a combination of Scp1 and Scr1 was 90%, 93.3%, and 99.99%, respectively. Similarly, when the initial concentration of nitrite was 0.25 mg/L, the degradation efficiency after 4 days of inoculation with Scp1, Scr1, and a combination of Scp1 and Scr1 was 60%, 82%, and 97.2%, respectively. In addition, when the initial concentration of COD was 20 mg/L, the degradation efficiency after 6 days of inoculation with Scp1, Scr1, and a combination of Scp1 and Scr1 was 59%, 59.4%, and 93.75%, respectively. The results demonstrated that the combined bacteria, Scp1 and Scr1, had a better degradation effect on ammonia nitrogen, nitrite, and COD. Furthermore, a degradation test was conducted in a Penaeus vannamei breeding base, which showed good degradation effects. These findings provide theoretical support for the treatment of high ammonia nitrogen wastewater in aquaculture and have important practical applications.

1. Introduction

China’s aquaculture industry has experienced rapid growth in recent years. However, this growth has come at the expense of the ecological environment, resulting in increased concentrations of ammonia nitrogen and nitrite. These pollutants have detrimental effects on animal health and have caused significant losses within the industry [1].

Despite these challenges, the aquaculture industry remains vital for increasing fishermen’s income and contributing to the country’s foreign exchange. To improve efficiency and quality, the industry has adopted various feed and drug innovations. Unfortunately, these advancements have also contributed to the rise in ammonia nitrogen and nitrite levels in aquaculture water [2].

High concentrations of ammonia nitrogen in aquaculture water can lead to decreased dissolved oxygen levels and an increase in harmful substances. This poses a significant threat to the safety of aquatic animals. Moreover, these harmful substances can be transmitted through the food chain, potentially endangering human health [3,4]. In 2023, China experienced a serious red tide disaster. Throughout the year, there were a total of 77 red tides, covering an area of 15,000 square kilometers. This resulted in economic losses of about 1 billion yuan and had a significant impact on the marine ecological environment. One of the major causes of pollution is ammonia nitrogen. Currently, the main pollutants found in aquaculture wastewater include ammonia nitrogen, nitrite, organic pollutants, and phosphorus [5]. During aquaculture, organic waste such as feed residues and feces enter the water, causing water turbidity and providing nutrients, which in turn leads to eutrophication and problems like blooms [5]. The decomposition of organic waste also produces compounds such as nitrate and nitrite, which have a serious impact on the nitrogen cycle and ecosystem of water bodies [5,6]. Furthermore, aquaculture may involve the use of feed containing heavy metals, pesticides, and veterinary drugs, which are released into the wastewater and pose toxic effects on aquatic organisms. These harmful substances also accumulate in water, posing potential threats to the ecological environment [7,8]. These pollutants have the potential to affect aquatic organisms and human health through the food chain.

Ammonia nitrogen exists in the form of molecular ammonia (NH₃) and ionic ammonium (NH₄⁺). Among them, NH₃ is extremely toxic, and the proportion increases with the increase in water temperature and pH (such as toxicity doubling when pH > 9). NH₃ can penetrate the gill mucosa of fish, inhibit ammonia excretion, increase blood pH, reduce the oxygen carrying capacity of red blood cells, and lead to physiological hypoxia (even if the dissolved oxygen in water is sufficient). It can also corrode the gill filaments and intestinal mucosa, causing congestion and bleeding, which is manifested by purple-black gills and increased mucus. Long-term exposure of aquatic organisms to low concentrations of NH₃ (>0.05 mg/L) can lead to reduced food intake, growth retardation, and decreased immunity. For example, the maximum tolerated concentration of NH₃ in grass carp fry was 0.054–0.099 mg/L; the growth of tilapia was inhibited at 0.035–0.171 mg/L. When the concentration of NH₃ increases, the harm is greater. NH₃ concentration > 0.66 mg/L can cause acute poisoning of carp species. The symptoms include frenzied swimming, spasm, exophthalmos, and many deaths within 24–48 h.

Ammonia nitrogen is excreted by aquatic organisms and is a final product of the decomposition of nitrogen-containing organic matter such as leftover bait, feces, and dead organisms [9]. Hydrogen sulfide is produced by the breakdown of sulfur-containing organic matter under anaerobic conditions [10]. Nitrite, on the other hand, is an unstable intermediate product found in aquaculture wastewater. Various techniques are used to degrade ammonia nitrogen and nitrite, including physical, chemical, and biological methods [5,10]. While physical and chemical methods like ion exchange and adsorption have their downsides such as sensitivity, short lifespan, possible secondary pollution, and high costs, the biological nitrogen removal method is not suitable for widespread application due to its long processing time, large land requirement, high energy consumption, and high investment and operation costs. In contrast, the development and utilization of new ammonia nitrogen degrading bacteria as compound microbial inoculants has shown promise in reducing treatment time, lowering operation and maintenance costs, and improving nitrogen removal efficiency [11].

Ammonia degradation and ammonia assimilation are both methods for sewage treatment. Ammonia degradation is carried out by nitrification and denitrification of bacteria. In the process of nitrification, high oxygen consumption leads to the decrease in Chemical Oxygen Demand (COD) in wastewater. Ammonia assimilation does not lead to a redox reaction. Bacteria directly absorb NH₃ to synthesize nitrogen-containing organic compounds such as amino acids and produce a large amount of microbial sludge. In this study, the COD is content in the wastewater decreased significantly, so the mechanism of reducing ammonia in this study was mainly ammonia degradation. COD represents the total amount of organic matter that can be oxidized by chemical oxidants in the wastewater. The organic matter in the wastewater is mainly degraded by heterotrophic bacteria as carbon source and energy. The process of degrading organic matter requires a large amount of dissolved oxygen (DO). The degradation of ammonia nitrogen (nitrification process) is completed by autotrophic nitrifying bacteria, which also need to consume a large amount of dissolved oxygen. In the case of elevated COD, heterotrophic bacteria will rapidly proliferate and consume a large amount of DO, which may lead to insufficient DO concentration in the system, thereby inhibiting the activity of nitrifying bacteria with higher DO demand and slower growth, directly affecting the removal efficiency of ammonia nitrogen. The decrease in COD level is helpful to improve the activity of ammonia nitrogen degrading bacteria, so as improve the removal efficiency of ammonia nitrogen.

For example, Zhang et al. isolated a novel ammonia-degrading bacterium, identified as Klebsiella pneumoniae TN-1, from fermentation residue (TFDs). It achieved an ammonia nitrogen removal rate of 83.26% when the initial concentration was 300 mg/L [12]. Huang et al. screened three dominant bacteria (X1, X2, and X3) from farmland sludge, and the X123 mixed bacteria proved to be the most effective with a degradation rate of 96.4% when the initial NH3-N concentration was 20 mg/L, pH was 7, and C/N ratio was 10 [13]. Wang et al. isolated a strain of bacteria with strong ammonia nitrogen degradation ability from pig farm wastewater, identified as Bacillus tequilensis. It achieved a 95% ammonia nitrogen degradation rate under specific conditions: culture temperature of 42 °C, initial pH of 7, seed volume of 5%, rotation speed of 160 rpm, and C/N ratio of 10:1 [14]. However, there is limited research on the immobilization of complex flora, particularly regarding the proportion and construction of these complex flora. In this study, we screened two new strains from surface water that have a strong ability to degrade ammonia nitrogen, nitrite, and COD. We then tested their degradation ability under laboratory conditions, specifically simulating high ammonia nitrogen wastewater. Based on the efficiency of various ammonia nitrogen degrading bacteria, we selected the best ratio and constructed a composite flora with a 1:1 ratio. This composite flora was also used in the aquaculture water of Penaeus vannamei as a reference for future theoretical research and practical application. Therefore, this study fills the gaps in the screening of high-efficiency ammonia nitrogen degrading bacteria and the treatment of high-concentration ammonia nitrogen aquaculture wastewater by immobilized composite bacteria.

2. Materials and Methods

2.1. Screening of Pseudomonas Scp1

Surface water samples were collected from Qihe County in Shandong Province and the main campus of Jinan University. One milliliter of surface water was weighed and added to 100 mL of LB enrichment culture. The mixture was then shaken in a shaker at 180 rpm and 30 °C until bacterial turbidity was observed. The enriched culture medium was diluted using a gradient dilution of 10⁻¹ to 10⁻⁴ four times. Raw water, 10⁻², and 10⁻⁴ water samples (100 µL) were coated on M9 solid medium containing the following components: Na₂HPO₄·12H₂O (128 g/L), KH₂PO₄ (30 g/L), NaCl (5 g/L), NH₄Cl (10 g/L), and agar (15 g/L). The solid medium also contained ampicillin (Amp, 50 µg/mL) and other components, namely 2 mL of 1 M MgSO₄ solution and 25 mL of 1 M glucose solution. The coated solid medium was then incubated at 28 °C to allow for colony growth. Single colonies were picked and confirmed by cultivating them in 1 mL of liquid medium.

2.2. Screening of Rhodococcus Scr1

Surface water was diluted in gradients of 100, 10⁻⁴, 10⁻⁵, 10⁻⁶, 10⁻⁷, 10⁻⁸, and 10⁻⁹. One milliliter of each diluted water sample was added to 100 mL of a thiosulfate liquid medium. The medium contained various components: Na₂S₂O₃ (2.5 g), MgCl₂·6H₂O (0.2 g), CaCl₂·2H₂O (0.1 g), NH₄Cl (0.1 g), K₂HPO₄ (0.1 g), and KCl (0.1 g). The liquid medium was sterilized and then supplemented with 1 mL of a trace element mixed solution, 1 mL of a vitamin mixed solution, 1 mL of a selenite–tungstate mixed solution, and 30 mL of a 2.4 mM NaHCO₃ solution. The selenite–tungstate mixed solution was prepared with NaOH (0.5 g), Na₂SeO₅·H₂O (0.3 mg), Na₂WO₄·2H₂O (0.4 mg), and ddH₂O (1 L), and was sterilized by filtration. The mixture was placed in a shaker at 28 °C and 100 rpm for cultivation.

2.3. Identification and Preservation of Microbial Strains

PCR was used to amplify the 16S rRNA genes of Pseudomonas and Rhodococcus erythropolis. The primers used were 27F (5′-GTTTGATCCTGGCTCA-3′) and 1492R (5′-TACGGCTACCTTGTTACGACTT-3′). The PCR conditions consisted of an initial denaturation step at 95 °C for 3 min, followed by 34 cycles of denaturation at 95 °C for 30 s, annealing at 55 °C for 30 s, and extension at 72 °C for 1 min. A final extension step was performed at 72 °C for 5 min. The amplification products were analyzed using 1% agarose gel electrophoresis, purified, and sequenced by Qingdao Qingke Biotechnology Co., Ltd. (Qingdao, China). The obtained sequences were compared with other published 16S rRNA gene sequences in NCBI using the BLAST tool. A phylogenetic tree was constructed using MEGA 7.0. The newly identified strains were preserved in the China Center For Type Culture Collection (CCTCC).

2.4. Laboratory Pilot Experiment

Dipped the preserved bacteria liquid from the preservation tube and streaked it on LB solid medium. Waited for a single colony to grow out. Picked single colonies into 1 mL of LB liquid medium to prepare the seed liquid. Prepared and sterilized two bottles of 100 mL LB liquid medium at 121 °C. Inoculated 1 mL of seed liquid into 100 mL of LB liquid medium at 28 °C and 150 rpm to obtain the activated bacterial liquid.

To test the effect of mixed bacteria on ammonia nitrogen degradation, we used the initial concentration of 2.5 mg/L ammonia nitrogen wastewater. The experiment was divided into a control group and three experimental groups. The control group was high ammonia nitrogen water without adding any strains. The inoculation amounts of high ammonia nitrogen water in the experimental group was 1%. The experimental group 1 was 100 mL high ammonia nitrogen water. Inoculated 1 mL Scp1 bacteria liquid, experimental group 2 was inoculated with 1 mL Scr1 bacteria liquid in 100 mL high ammonia nitrogen water, and experimental group 3 was inoculated with 0.5 mL Scp1 bacteria liquid and 0.5 mL Scr1 bacteria liquid in 100 mL high ammonia nitrogen water. The specific experimental mixing ratio was determined according to

Table 1. Proportion experiments of different experimental groups (unit: mL).

No	Bacterial Strain (Inoculation Amount 1%)	Water Sample
1	Scp1	100 mL
2	Scr1	100 mL
3	Scp1 and Scr1	100 mL
CK	——	100 mL
Total	——	100 mL

.

The activated bacterial solution was diluted 10⁻⁴, and the ammonia nitrogen degradation experiment was conducted according to the groups mentioned above. Take 100 mL of high ammonia nitrogen wastewater and add 1 mL of the diluted single or mixed bacterial solution to achieve a final concentration of 10⁻⁶ for the bacterial solution. The water samples were cultured at room temperature (20 °C). Ammonia nitrogen degradation was tested on the 0th, 1st, 2nd, 3rd, 4th, 5th, and 6th days. The test method was determined using the salicylic acid–hypochlorite photometric method. Dissolved oxygen was determined using the iodometric method, pH value was determined using a precision pH meter, and nitrite nitrogen content was determined using the Nessler colorimetric method. The COD content was determined using the national environmental protection standard HJ 828-2017 dichromate method.

2.5. Specific Experimental Steps

Determination of ammonia nitrogen by salicylic acid–hypochlorite spectrophotometry: 1. Sample pretreatment: 1 mL 10 % ZnSO₄ and 0.1–0.2 mL 50% NaOH were added to the wastewater, and the supernatant was taken after precipitation. Add 0.1 g activated carbon adsorption 10 min, filtration. 2. Color reaction: Take 25 mL treated water sample in 50 mL colorimetric tube. Add in turn: 1.0 mL 10% potassium sodium tartrate, 1.5 mL salicylic acid–sodium citrate solution, 0.3 mL sodium hypochlorite solution, 0.25 mL sodium nitroferricyanide, distilled water to 50 mL, mixed well and reacted in the dark for 60 min (20–25 °C). 3. Determination: The absorbance was measured at a wavelength of 697 nm with a 10 mm cuvette. The standard curve was made with the ammonia nitrogen concentration of 0, 0.10, 0.50, 1.00, and 2.00 mg/L with the abscissa absorbance as the ordinate. The absorbance measured in different treatment groups at 0–6 d was brought into the standard curve and a line chart was made.

Determination of dissolved oxygen (DO) by an iodimetry: 1. On-site fixation: water samples were collected with dissolved oxygen bottles, overflowing 2–3 times the volume (avoiding bubbles). Immediately insert the pipette into the liquid surface to add: 1 mL MnSO₄ solution (480 g/L) and 1 mL alkaline potassium iodide solution (150 g NaOH + 50 g KI/L). Cover the bottle stopper tightly, mix upside down for 15 times, and stand until the precipitate drops to the bottom of the bottle. 2. Acidification titration: After adding 1.5 mL concentrated H₂SO₃ (slowly injected along the bottle wall), the bottle body was reversed until the precipitate was completely dissolved (the solution was brownish yellow). A total of 100 mL solution was taken in a conical flask and titrated with 0.025 mol/L Na₂SO₃ to light yellow, and then 1 mL 0.5% starch solution (blue) was added and titrated until the blue color disappeared. 3. Calculation:

DO (mg/L) = C × V × 8 × 1000VsDO (mg/L) = VsC × V × 8 × 1000.

C: Na₂S₂O₃ concentration (mol/L), V: titration volume (mL), Vs.: water volume (mL).

Determination of nitrite nitrogen by Nessler’s reagent colorimetric method: 1. Sample treatment: The filtrate was removed after filtration with 0.45 μm filter membrane. 2. Colorimetric determination: 50 mL water sample was put into a 50 mL colorimetric tube, added with 1.0 mL p-aminobenzenesulfonic acid solution (10 g/L), mixed well and allowed to stand for 3–5 min, then added with 1.0 mL naphthylethylenediamine hydrochloride solution (1 g/L), mixed well and allowed to stand in the dark for 30 min. The absorbance was measured at 540 nm with a 10 mm cuvette. 3. Standard curve: The standard curve was made with the horizontal coordinate absorbance of nitrite nitrogen concentration of 0, 0.10, 0.20, 0.50 and 1.00 mg/L as the vertical coordinate. The absorbance measured in different treatment groups at 0–6 d was brought into the standard curve and a line chart was made.

Determination of chemical oxygen demand (COD) by dichromate method: 1. Digestion reaction: 20.00 mL of water sample was taken in 250 mL conical flasks. Add: 10.00 mol/L K₂Cr₂O₇ standard solution, several explosion-proof boiling glass beads, slowly add 30 mL H₂SO₃-Ag₂SO₃ solution (containing 10 g/L Ag₂SO₃). Connect the reflux condenser tube and heat the reflux for 2 h (timing from boiling). 2. Titration determination: After cooling, the condenser tube was rinsed with 40 mL pure water, and 3 drops of ferrous iron indicator (1,10-phenanthroline ferrous sulfate) was added, and 0.1 mol/L ammonium ferrous sulfate ((NH₄)2Fe(SO₄)₂) was titrated to the end point of reddish brown → blue-green → reddish brown. 3. Calculation:

COD (mg/L) = (V₀ − V₁) × C × 8000VsCOD (mg/L) = Vs(V₀ − V₁) × C × 8000.

V₀: blank titration volume (mL), V₁: sample titration volume (mL), C: ammonium ferrous sulfate concentration (mol/L), Vs.: water volume (mL).

2.6. Shrimp Pond Experiment

For the shrimp pond experiment, the optimal combination was chosen from the previous experiments. The shrimp pond is located at the Penaeus vannamei breeding base in the coastal village of Ju Town, Rudong County, Nantong City, Jiangsu Province. In the three test ponds, samples were taken from the water surface, 60 cm below the water surface, and the sediment. Each sample was 500 mL in volume. The ammonia nitrogen, pH, nitrite, and COD content in the three pond samples were detected. The detection methods for ammonia nitrogen, nitrite, pH, and COD were performed using laboratory flask experiments.

3. Results

3.1. Screening and Identification Results

After separation and screening, we obtained two strains. The 16S rDNA sequence was identified using the 27F (5′-AGAGTTTGATCCTGGCTCAG-3′) and 1492R (5′-TACGGCTACCTTGTTACGACTT-3′) primers. The results of the BLAST alignment showed that one of the strains belonged to Pseudomonas and was named Scp1. It had a 99.93% similarity to Pseudomonas sp. strain 382-a white. The other strain belonged to the genus Rhodococcus and was named Scr1. It had a 99.79% similarity to Rhodococcus erythropolis strain RE-1. On LB solid medium, Scp1 colonies were round, light yellow transparent, smooth, and moist. Scr1 colonies were round, with a dry and rough surface that was opaque. Phylogenetic trees based on the 16S RNA sequences of Scp1 and Scr1 were constructed using the neighbor joining algorithm. Please refer to Figure 1 and Figure 2 for the alignment results and phylogenetic trees. The preservation numbers of Pseudomonas sp. SCP1 (CCTCC NO: M 20231698) and Rhodococcus erythropolis SCR1 (CCTCC NO: M 20231697) comply with the requirements of the China Center For Type Culture Collection (CCTCC) and have been preserved according to the regulations.

3.2. Results of Ammonia Nitrogen Degradation Experiment in Polluted River Water

Figure 3 shows the results of the laboratory experiment where wastewater with an initial concentration of 2.5 mg/L ammonia nitrogen was selected for degradation. In the control group (CK) without inoculation, the concentration of ammonia nitrogen decreased. In the experimental groups with Scp1, Scr1, and a combination of Scp1 and Scr1, the concentration of ammonia nitrogen decreased significantly within 4 days. In the control group (CK), the ammonia nitrogen concentration decreased to 0.4 mg/L. In the group inoculated with Scp1, the ammonia nitrogen concentration decreased to 0.2 mg/L, as well as in the group inoculated with Scr1. In the combined group of Scp1 and Scr1, the ammonia nitrogen concentration decreased to 0, indicating a significant degradation effect. In summary, the composite bacteria Scp1 and Scr1 showed a rapid removal effect on ammonia nitrogen, with Scr1 and Scp1 together achieving the best removal effect.

3.3. Results of Laboratory Nitrogen Degradation Experiment

Pseudomonas, a Gram-negative bacteria, and Rhodococcus, a Gram-positive bacteria, exhibit diverse metabolic capabilities and environmental adaptability. These bacteria have garnered much attention in the fields of wastewater treatment and biodegradation due to their ability to degrade a wide array of organic and inorganic pollutants. Pseudomonas, in particular, plays a crucial role in the nitrogen cycle by degrading nitrite (NO2⁻) through its nitrite reductase (NiR), facilitating biological denitrification in the environment [15]. In Figure 4, it is evident that the initial concentration of nitrite in the selected water sample was 0.25 mg/L. After 4 days, the nitrite concentration in the control group (CK) increased to 0.3 mg/L. However, the water samples inoculated with Scp1 and Scr1 exhibited a decrease in nitrite content to 0.005 mg/L, while the water samples inoculated with Scr1 alone showed a decrease to 0.01 mg/L.

3.4. Results of Laboratory Vial pH Change Experiment

The pH changes observed in the laboratory vial experiment are displayed in Figure 5. The initial pH of the wastewater was found to be around 7.2. After four days, the pH of the control group decreased slightly to about 7.4. In contrast, the water samples inoculated with Scp1 and Scr1 exhibited an increase in pH to about 8.2, while the water sample inoculated with Scp1 alone showed a pH increase to about 8.0, and the water sample inoculated with Scr1 alone had a pH increase to about 8.0. There was no correlation between the degradation of ammonia nitrogen and the change in pH value. The reason why the pH value of wastewater was detected was to show that the pH value of the three experimental groups in this study was within a reasonable range during the degradation of ammonia nitrogen. It is worth noting that the appropriate pH range for fish and shrimp is typically 6.5–8.5, which is weakly alkaline. Outside of this range, fish and shrimp may experience negative effects. When Scp1 and Scr1 absorb NH as a nitrogen source through assimilation, they discharge H ions into the environment and directly consume the acid (H) in the environment, resulting in an increase in pH. This process is particularly significant in the presence of high ammonia nitrogen and available organic carbon (BOD). For South American white shrimp, the ideal pH range for growth and development is between 7.8 and 8.6. This slightly alkaline condition is favorable for their physiological processes, including digestion, immune response, and molting [15].

3.5. Results of COD in Laboratory Vials

The results of the laboratory vials also included the measurements of COD, but they are not provided in the given text. The change in COD is shown in Figure 6. The initial COD concentration of the wastewater is approximately 20 mg/L. After four days, the control group’s COD concentration decreased to about 18.9 mg/L, the water sample inoculated with Scp1 decreased to about 15.01 mg/L, the water sample inoculated with Scr1 decreased to about 13.28 mg/L, and the water sample inoculated with both Scp1 and Scr1 decreased to about 8.2 mg/L. After six days, the water sample inoculated with both Scp1 and Scr1 decreased further to about 1.25 mg/L. The normal range for aquaculture COD is 0.1–10 mg/L. Higher COD levels typically indicate more organic matter in the water, which can lead to a decrease in dissolved oxygen content. This occurs because microorganisms consume a large amount of dissolved oxygen when decomposing organic matter [16]. Aquaculture requires sufficient dissolved oxygen for the metabolism and growth of shrimp. In severe cases, insufficient dissolved oxygen can even lead to the death of fish and shrimp [17].

3.6. Results of Penaeus Vannamei Aquaculture Water Test

When Scp1 and Scr1 are combined to treat high ammonia nitrogen wastewater, synergistic denitrification is achieved through complementary metabolic functions and microenvironment reciprocity. Scp1 utilizes ammonia monooxygenase (AMO) to efficiently oxidize ammonia (NH₃) to nitrite (NO₂⁻), providing denitrification substrates [18,19]; Scr1 was photo-driven denitrification to generate N₂ [18] with the nitrite as an electron acceptor under anaerobic light, and directly assimilated ammonia nitrogen to synthesize biomass [20]. The two form a cross-nutritional cascade cycle: Scr1 cooperates with the release of O₂ to support the continuous ammonia oxidation of Pseudomonas [21], while the CO₂ produced by Scp1 metabolism acts as a photosynthetic carbon source for Rhodobacter [19]. The system significantly improves stress resistance by diverting ammonia nitrogen load (Scp1 treats acute ammonia shock, Scr1 eliminates nitrite accumulation) [22], and uses photosynthetic reduction force reducing energy consumption, and finally realizes high ammonia nitrogen and low C/N ratio wastewater. Efficient denitrification and system stability [20,22].

We used the optimal experimental conditions identified in the laboratory to inoculate the mixed bacteria of Scp1 and Scr1 at a ratio of 1/100,000 of the total water volume. The degradation test was conducted in the shrimp ponds at the Litopenaeus vannamei breeding base in the coastal village of Juzhen Street, Rudong County, Nantong City, Jiangsu Province. Prior to inoculation, the concentration of ammonia nitrogen was 0.3 mg/L, the concentration of nitrite was 0.19 mg/L, and the COD content was 9.50 mg/L. After four days, we determined the concentrations of ammonia nitrogen, nitrite, pH, and COD using laboratory methods, as shown in Figure 7. The concentration of ammonia nitrogen decreased to 0 mg/L, achieving a degradation rate of 100%. The concentration of nitrite decreased to 0.005 mg/L, achieving a degradation rate of 97.3%. The concentration of COD decreased to 1.2 mg/L, achieving a degradation rate of 87.4%. These results demonstrate that the compound microbial agent can significantly reduce the concentrations of ammonia nitrogen, nitrite, and COD in aquaculture water. The high levels of ammonia nitrogen, nitrite, and COD in shrimp ponds can have detrimental effects on the growth and health of Penaeus vannamei [23]. Shrimp may experience shortness of breath, loss of appetite, slow growth, and even death when exposed to high concentrations of ammonia nitrogen and nitrite. Prolonged exposure to a high ammonia nitrogen environment can weaken shrimp immunity, making them more susceptible to pathogen invasion [24]. Additionally, high COD levels can deplete dissolved oxygen, leading to water hypoxia. This oxygen-deprived environment can negatively impact shrimp metabolism and growth, resulting in a significant number of deaths in severe cases [25]. Furthermore, the decomposition of organic matter can produce harmful substances like hydrogen sulfide, which can harm shrimp.

In this study, two strains capable of degrading ammonia nitrogen and nitrite were isolated from surface water: Scp1 and Scr1. When the initial concentration of ammonia nitrogen was 1.50 mg/L, the degradation efficiency of ammonia nitrogen was 90%, 93.3%, and 99.99%, respectively, after 4 days of inoculation with Scp1, Scr1, and the combination of Scp1 and Scr1. Similarly, when the initial concentration of nitrite was 0.25 mg/L, the degradation efficiency of Scp1, Scr1, and the combination of Scp1 and Scr1 was 60%, 82%, and 97.2%, respectively, after 4 days of inoculation. In addition, when the initial concentration of chemical oxygen demand (COD) was 20 mg/L, the degradation efficiencies of Scp1, Scr1 and the combination of Scp1 and Scr1 were 59%, 59.4%, and 93.75%, respectively, after 4 days of inoculation. The results showed that the composite bacteria of Scp1 and Scr1 had better degradation effect on ammonia nitrogen, nitrite, and chemical oxygen demand (COD). In the degradation test of Litopenaeus vannamei breeding base, it also showed good degradation effect. After 4 days, the degradation rate of ammonia nitrogen was 100%, the degradation rate of nitrite was 97.3%, and the COD was reduced by 87.4%.

4. Discussion

The results of this study demonstrate the efficacy of a composite microbial agent in degrading ammonia nitrogen, a significant pollutant in aquaculture. These findings highlight the potential of microbial agents as a viable strategy for nitrogen removal, aligning with the growing interest in high ammonia nitrogen aquaculture water treatment methods [26].

4.1. Comparison of Ammonia Nitrogen Degradation Performance

Our study utilized a novel composite microbial agent that proved highly effective in reducing ammonia nitrogen levels. The synergistic action of the two strains within the agent likely contributed to the observed enhanced degradation efficiency. This is consistent with previous research indicating that microbial consortia often outperform single strains in biodegradation processes due to their complementary metabolic capabilities [27].

Li et al.in the in fluent ammonia nitrogen concentration of 400.00–500.00 mg/L, water temperature 30 °C, pH value of 8 under the condition of shaking culture for 36 h, the ammonia nitrogen degradation rate of the composite bacteria was 85.43–89.84% [28]; Wu removed ammonia nitrogen by adding compound microbial agents. After 21 days of operation, the removal rate of ammonia nitrogen finally reached 98.83%, which was 350.00 mg/L [29]. Ding screened a highly efficient ammonia nitrogen and nitrate-nitrogen degrading bacteria from the culture water of Litopenaeus vannamei, and the ammonia nitrogen degradation rate reached 90.00% [30]. The results of this experiment provide two new high-efficiency ammonia nitrogen degrading bacteria. The two bacteria are combined into a composite strain. After 4 days of operation, the degradation rate of ammonia nitrogen is 99.99%, which is better than the first three studies.

The degradation ability of the composite microbial agent was demonstrated through a rapid reduction in ammonia nitrogen levels within four days. Within six days, the content of ammonia nitrogen was close to zero, demonstrating the agent’s effective degradation capabilities. Further studies are needed to elucidate the specific mechanisms involved and to optimize the operational conditions for sustained degradation.

4.2. Metabolic Mechanism of Ammonia Nitrogen Degradation

The degradation process of ammonia nitrogen undergoes nitrification and denitrification. During the nitrification process, ammonia nitrogen finally generates nitrate through the nitrification of Rhodococcus, and the nitrate in the denitrification process is released as N₂ through the denitrification of Pseudomonas. The specific reaction is as follows:

Nitrification: 2NH₄⁺ + 3O₂ → 2NO₂⁻ +4H⁺ + 2H₂O⁺ energy

Denitrification: 5C₆H₁₂O₆ + 24NO₃⁻ → 12N₂↑+30CO₂ + 18H₂O + 24OH⁻

Microorganisms have metabolic complementarity. Scp1 oxidizes ammonia nitrogen (NH₃) to hydroxylamine (NH₂OH) through ammonia monooxygenase (AMO), and then converts to nitrite (NO₂⁻) through hydroxylamine oxidoreductase (HAO). Scr1 expresses nitrite reductase (NirS/NirK) to reduce NO₂ to NO, which is finally converted to nitrogen (N₂) by nitric oxide reductase (Nor) and nitrous oxide reductase (Nos). The synergistic effect of microorganisms is as follows: the nitrification of Scp1 provides a reaction substrate for Scr1 to avoid the toxicity of nitrite accumulation; the denitrogenation effect of Scr1 alleviates the product inhibition of Scp1.

Dissolved oxygen (DO) limitation, low DO (<2 mg/L) inhibits nitrification, and high DO hinders denitrification. A graded aeration biofilm reactor can be designed to achieve partition oxygen control. Heavy metal toxicity, Cu²⁺/Zn²⁺ competitive enzyme activity sites lead to metabolic stagnation. Pre-adsorption treatment (such as biochar immobilization of heavy metals) is required.

4.3. Limitations and Challenges

One critical factor influencing the performance of the composite microbial agent is the environmental conditions, such as pH and temperature [31]. Our findings indicate that the agent performs optimally under neutral pH conditions and moderate temperatures, which are commonly found in many wastewater treatment settings. However, the agent’s sensitivity to extreme conditions highlights the need for further research to enhance its robustness [32].

Economic analysis of the treatment process showed that the use of the composite microbial agent is cost-effective compared to conventional chemical methods. This is a significant advantage considering the high costs associated with ammonia nitrogen removal in aquaculture wastewater treatment.

Despite the promising results, there are limitations to our study [33]. The degradation effects of the compound microbial agent largely depend on the activity and stability of its internal microorganisms. Environmental conditions such as temperature, pH, and salinity can affect the growth and metabolic activities of microorganisms. For example, high salinity environments may inhibit the growth of certain strains, thereby reducing the overall degradation ability of the compound microbial agent [33,34]. It is therefore crucial to find ways to maintain the activity and stability of microorganisms under different environmental conditions, which requires further verification.

The scale-up of the microbial agent application to large-scale aquaculture wastewater treatment plants requires further validation. Additionally, the long-term ecological impact of introducing a composite microbial agent into natural ecosystems needs to be carefully assessed [35,36].

In conclusion, the composite microbial agent shows great promise for ammonia nitrogen degradation in wastewater treatment. Future research should focus on optimizing the agent’s performance under various environmental conditions and assessing its long-term environmental impact.