Isolation, Identification, and Condition Optimization of an Ammonia-Oxidizing Bacterium and Its Potential Application in Wastewater Treatment

Abstract

Ammonia-oxidizing bacteria (AOB) are the rate-limiting microbial group in biological nitrogen removal. However, difficulties in isolation and cultivation, along with unclear metabolic regulation mechanisms, have long constrained their engineering applications. Existing research has mostly focused on the responses of model strains to environmental factors, while studies on the growth regulation mechanisms driven by key medium components remain scarce. Moreover, there is a lack of efficient strains suitable for complex wastewater with high ammonia and high salinity. To isolate an efficient strain and optimize its culture conditions for high-ammonia wastewater treatment, we collected water samples from a polluted river in Zhongshan City. After enrichment, a strain was isolated via gradient dilution and silica gel plating, identified by scanning electron microscopy and 16S rDNA sequencing as Nitrosomonas europaea W4 (99.93% similarity to the type strain). Single-factor medium optimization examined CaCO₃ and Fe²⁺/Fe³⁺, while temperature and initial ammonia nitrogen effects were tested, and landfill leachate was used for verification. CaCO₃ shortened the lag phase without affecting maximum specific growth rate; replacing Fe³⁺ with Fe²⁺ further reduced lag and enhanced the ammonia oxidation rate. Optimal growth occurred at 25–30 °C and an initial ammonia nitrogen concentration of ~2000 mg/L. In landfill leachate, the strain increased the ammonia degradation rate 6.3-fold, from 4.64 ± 0.11 mg L⁻¹ d⁻¹ in the uninoculated control group to 29.32 ± 0.07 mg L⁻¹ d⁻¹, and importantly, nitrite can be rapidly degraded by indigenous denitrifiers, posing no secondary pollution risk. Overall, N. europaea W4 exhibits high ammonia oxidation efficiency, and the optimized medium and conditions improve its proliferation and metabolic stability, providing a basis for cultivation and application in treating high-strength ammonia nitrogen wastewater.

1. Introduction

Ammonia-oxidizing bacteria (AOB) exhibit high substrate specificity and require stringent environmental conditions for growth. In natural environments, AOB frequently compete with heterotrophic bacteria, resulting in low isolation efficiency [1]. Due to their chemoautotrophic nature, AOB rely on specific substrates and favorable environmental conditions for growth. Even under suitable cultivation conditions, AOB generally exhibit slow growth rates, with doubling times typically exceeding 10 h. Consequently, they are readily outcompeted by fast-growing heterotrophic microorganisms, making it difficult to obtain high-purity strains using conventional isolation methods [2]. The cultivation efficiency of AOB is influenced by multiple environmental factors, among which medium composition, temperature, and the initial ammonia nitrogen concentration are particularly important.

Iron is an essential trace element in the metabolism of AOB, as it directly participates in the active centers of both ammonia monooxygenase (AMO) and hydroxylamine oxidoreductase (HAO) [3]. The chemical form of iron significantly influences the growth and metabolic activity of AOB. Ferrous iron (Fe²⁺) functions as a cofactor for AMO. Insufficient Fe²⁺ concentrations may result in reduced enzyme activity, whereas excessive Fe²⁺ levels can promote the generation of reactive oxygen species (ROS) through the Fenton reaction, thereby inducing oxidative stress [4]. In addition, the sensitivity of AOB to Fe²⁺ varies among species. Compared with Fe²⁺, ferric iron (Fe³⁺) exhibits lower bioavailability and must first be converted into Fe²⁺ through siderophore-mediated transport or cellular reduction systems before microorganisms can utilize it. Moreover, excessive Fe³⁺ concentrations may compete with Fe²⁺ for membrane transporters, thereby inhibiting the catalytic activity of key enzymes involved in ammonia oxidation [5].

Calcium carbonate (CaCO₃) performs multiple functions in the cultivation of AOB. It acts as a pH buffer, thereby maintaining the alkaline conditions required for ammonia oxidation, and simultaneously provides an inorganic carbon source for CO₂ fixation by chemoautotrophic microorganisms [6]. In addition, CaCO₃ particles can serve as attachment sites for bacterial cells [7] and may further enhance quorum-sensing processes [6]. However, the low solubility of CaCO₃ may adversely affect the medium composition when excessive amounts are added. Elevated Ca²⁺ concentrations can promote precipitation with phosphate ions, thereby reducing phosphorus bioavailability in the culture medium [8].

In addition to the medium composition, temperature is a key environmental factor regulating AOB metabolic activity, determining their growth kinetics by influencing enzymatic reaction rates and cell membrane fluidity. Antoniou et al. [9] observed that, within the experimental temperature range of 15–25 °C, the effective maximum specific growth rate (μ_max) of Nitrosomonas increased from 0.12 to 0.97 d⁻¹ with increasing temperature. However, significant inter-strain differences in temperature adaptability exist among various N. europaea isolates, and studies on temperature suitability under practical wastewater treatment scenarios remain scarce.

N. europaea exhibits low affinity toward ammonia but demonstrates strong tolerance to elevated ammonia nitrogen concentrations. In contrast, its growth is inhibited under conditions of low ammonia nitrogen availability [10]. However, excessively high ammonia nitrogen levels can lead to the accumulation of free ammonia (FA). The FA concentration is jointly influenced by pH and ammonia nitrogen levels. Elevated FA concentrations exert substantial inhibitory effects on the growth and metabolic activity of N. europaea [11].

Despite extensive research on ammonia-oxidizing bacteria, three critical knowledge gaps remain regarding Nitrosomonas europaea: (i) studies are almost exclusively based on the type strain ATCC 19718, leaving the genetic diversity of geographically distinct strains largely unknown [12], (ii) systematic medium optimization is lacking, and the mechanistic role of CaCO₃ as well as the differential effects of Fe²⁺/Fe³⁺ have not been elucidated [13]; (iii) environmental adaptability assessments have been largely confined to synthetic media, without validation in real complex wastewaters. Therefore, based on the systematic optimization of culture medium composition, the present study investigates the effects of CaCO₃, Fe²⁺, and Fe³⁺ on the growth and ammonia oxidation performance of the AOB strain N. europaea W4. Furthermore, the roles of temperature and initial ammonia nitrogen concentration in nitrogen removal performance are examined. Finally, the practical application potential of the strain is evaluated through bench-scale treatment experiments using landfill leachate. This study aims to provide support for the application of N. europaea W4 in wastewater bioaugmentation.

2. Materials and Methods

2.1. Sample Source

Water samples were collected from a section of the river affected by domestic sewage discharge in Zhongshan. Its physicochemical properties were as follows: pH 7.39, dissolved oxygen 0.52 mg/L, ammonia nitrogen concentration 3.93 mg/L, and temperature 26.00 °C. The samples were initially pre-filtered through a 0.45 μm nylon membrane to remove large suspended particles. Subsequently, the filtrates were centrifuged at 3000× g for 5 min (Model: Centrifuge5810R, Eppendorf, Hamburg, Germany), and the resulting pellets were collected for further analysis.

2.2. Media and Reagents

AOB liquid medium (pH 7) was modified based on DSMZ Medium 1583 [14]. The medium contained 0.5 g/L MgSO₄·7H₂O, 0.2 g/L NaCl, 0.1 g/L KH₂PO₄, 5.0 g/L CaCO₃, 1.5 g/L NaHCO₃, 0.37 g/L NH₄Cl, 0.05 g/L FeCl₃·6H₂O, and 0.1% trace element solution. Accurately weigh sequentially 0.062 g/L H₃BO₃, 0.017 g/L CuCl₂·2H₂O, 0.1 g/L MnCl₂·4H₂O, 0.036 g/L Na₂MoO₄·2H₂O, 0.07 g/L ZnCl₂, 0.19 g/L CoCl₂·6H₂O, and 0.024 g/L NiCl₂·6H₂O. Dissolve in ultrapure water (Model: UPT-II-10T, Sichuan Youpu Ultra-Pure Technology Co., Ltd., Chengdu, China) and bring the volume to 1 L. Store at 4 °C in the dark until use. The pH of the medium was adjusted to 7.0 with 1 mol/L NaOH and 1 mol/L HCl using a pH meter (METTLER TOLEDO, Mettler Toledo Instruments Co., Ltd., Shanghai, China).

The basal medium was sterilized by autoclaving at 121 °C for 20 min using an automatic sterilizer (Model: GR60DA, ZEALWAY, Wilmington, DE, USA). CaCO₃ was sterilized by dry heat at 160 °C for 2 h. The Fe²⁺/Fe³⁺ stock solution (10 g/L) was filter-sterilized through a 0.22 μm membrane and added aseptically after the medium had cooled to below 50 °C (Model: HH-6, Guohua Dianxi Co., Ltd., Beijing, China). The Fe²⁺ stock solution was prepared fresh immediately before use.

Silica gel plates were prepared according to the classic Winogradsky method for autotrophic bacteria isolation. Briefly, sodium silicate solution (specific gravity 1.10) was mixed with hydrochloric acid solution (specific gravity 1.09) at a volume ratio of 1:2 to form a silica gel. The gel was cut into small pieces and dialyzed against running tap water for 72 h until no chloride ions were detected by 1% silver nitrate solution. The dialyzed gel was autoclaved at 121 °C for 20 min and cooled to 50 °C. One milliliter of filter-sterilized (0.22 μm) AOB isolation medium was added to 10 mL of sterile silica gel solution, mixed gently, and poured into sterile Petri dishes. The plates were dried at 55 °C in a sterile incubator (Model: SHP-250, Jinghong, China) until no free liquid was present on the surface before use.

2.3. Isolation and Purification of AOB

2.3.1. Strain Enrichment

The collected cell pellet was inoculated into a 250 mL conical flask containing 95 mL of AOB liquid medium with an initial ammonia nitrogen concentration of 100 mg/L. The cultures were incubated at 30 °C and 150 r/min in the dark. Ammonia nitrogen and nitrite nitrogen concentrations were measured at 12 h intervals throughout the cultivation period. Once the culture reached the logarithmic growth phase, 5 mL of the enriched culture was transferred into fresh medium for subculturing. This enrichment procedure was repeatedly performed until stable ammonia degradation and nitrite accumulation were achieved. This concentration was chosen because it closely matches the ammonia nitrogen levels in typical municipal wastewater influent (50–150 mg/L) and pretreated landfill leachate (80–120 mg/L), ensuring the enriched strains can directly adapt to actual wastewater environments. It also provides sufficient nitrogen source for AOB growth while effectively eliminating low-tolerance contaminants, thus improving enrichment efficiency.

2.3.2. Strain Isolation

Primary screening: The AOB medium was sterilized by filtration through a 0.22 μm membrane and subsequently dispensed for later use. A 5 mL aliquot of the enrichment culture, with an initial ammonia nitrogen concentration of 100 mg/L, was inoculated into 95 mL of sterile AOB medium. The culture was incubated at 30 °C and 150 r/min in the dark until the logarithmic growth phase was reached (OD600 ≈ 0.1). Subsequently, 0.5 mL of the culture was transferred into 4.5 mL of sterile medium and thoroughly mixed. Serial 10-fold dilutions were then prepared up to 10⁻⁹. Before each gradient dilution, the bacterial suspension was dispersed by pulsed ultrasound (40 kHz, 160 W, 3 s on/5 s off for 1 min) in a KQ-800KDE high-power digital ultrasonic cleaner (Kunshan Ultrasonic Instruments Co., Ltd., Kunshan, China) to break up bacterial aggregates. Aliquots (200 μL) from the 10⁻⁶ to 10⁻⁹ dilutions were spread onto silica gel solid plates, and blank control plates were prepared simultaneously. All plates were sealed and incubated at 30 °C in the dark for 14 days. Single colonies were subsequently selected and inoculated into 5 mL of liquid medium, then incubated statically at 30 °C for 7 days. Nitrite production was quantified using the Griess assay. Cultures exhibiting nitrite concentrations greater than 20 mg/L were selected for further purification.

Secondary screening: The positive culture obtained from the primary screening was serially diluted 10-fold. Subsequently, 500 μL aliquots from the 10⁻⁷ and 10⁻⁸ dilutions were inoculated into two separate 96-well plates. The plates were sealed with Parafilm and incubated at 30 °C in the dark for 30 days. Nitrite concentrations in the wells were determined using the Griess assay. Cultures from nitrite-positive wells were subsequently transferred into beef extract peptone medium and incubated at 30 °C in the dark for 7 days. The absence of contamination was verified based on medium clarity and microscopic observation. The isolate was then preliminarily identified as an AOB strain.

2.4. Identification of AOB

2.4.1. Scanning Electron Microscopy

Cells harvested during the logarithmic growth phase were centrifuged at 3000× g for 10 min, and approximately 1 g of cell pellet was collected. The pellet was fixed with 2.5% glutaraldehyde for 4 h. After fixation, the supernatant was removed, and the cells were rinsed three times with 0.1 mol/L phosphate buffer, with each wash lasting 20 min. The samples were subsequently dehydrated using a graded ethanol series (30%, 50%, 70%, 80%, 90%, and 100%) with each step lasting 15 min. After ethanol removal, the samples were immersed in isoamyl acetate for 10 min under gentle agitation. The treated samples were then dried using a critical-point dryer, sputter-coated with gold, and finally observed and photographed using a field-emission scanning electron microscope (FE-SEM).

2.4.2. 16S rRNA Gene Sequence Analysis

The 16S rRNA gene was amplified using the universal primers 27F (5′-AGAGTTTGATCCTGGCTCAG-3′) and 1492R (5′-GGTTACCTTGTTACGACTT-3′). The polymerase chain reaction (PCR) mixture (25 μL total volume) consisted of 12.5 μL 2× Taq Master Mix (Vazyme, Nanjing, China), 1 μL each of forward and reverse primers (10 μM), 1 μL of template DNA, and 9.5 μL of double-distilled water (ddH₂O). The PCR amplification conditions were as follows: initial denaturation at 95 °C for 5 min; followed by 30 cycles of denaturation at 95 °C for 30 s, annealing at 55 °C for 30 s, and extension at 72 °C for 90 s. A final extension step was performed at 72 °C for 10 min. The purified PCR products were submitted to Sangon Biotech (Shanghai, China) for sequencing. The obtained sequences were analyzed using NCBI BLAST (https://blast.ncbi.nlm.nih.gov/Blast.cgi, accessed on 10 May 2026), and reference sequences with greater than 99% similarity were retrieved for comparative analysis. Phylogenetic analysis was performed using MEGA 11.0 software. A phylogenetic tree was constructed using the neighbor-joining (NJ) method, with bootstrap values set to 1000 replicates.

The 16S rRNA gene sequence of N. europaea W4 obtained in this study has been deposited in the EMBL-EBI database under accession number PRJNA245576.

2.4.3. Bacterial Genomic DNA Extraction Method

To minimize interference from medium impurities, the AOB liquid medium was sterilized and filtered through a 0.22 μm membrane filter. The bacterial cells were inoculated after the medium had cooled to near room temperature. Strain W4 was cultured to the logarithmic growth phase and then centrifuged at 8000 g for 5 min. Approximately 0.5 g of cell pellet was collected and used for DNA extraction with a MagPure Bacterial DNA Kit. The concentration of genomic DNA was measured using TBS380, and DNA fragment integrity was confirmed by pulsed-field gel electrophoresis. The purified genomic DNA was quantified, and high-quality DNA was used for subsequent library construction and sequencing. The raw sequencing data and assembled genome have been deposited in NCBI under the following accession numbers: chromosome sequence, CP157069.1; plasmid sequence, CP157070.1.

2.5. Effects of Medium Components and Environmental Factors on the Nitrogen Removal Performance of the Strain

2.5.1. General Experimental Methods

The tested strain was the ammonia-oxidizing strain isolated and purified above. Prior to inoculation, the pre-culture was grown to mid-exponential phase (OD₆₀₀ ≈ 0.1), and the inoculum density was adjusted based on the optical density at 600 nm (OD₆₀₀) to achieve an initial OD₆₀₀ of approximately 0.001 in the experimental flasks (corresponding to an inoculum size of approximately 1%, v/v). The adjusted bacterial suspension was then inoculated into 250 mL conical flasks containing 100 mL of sterile medium at an inoculum size of 1% (v/v). The flasks were sealed with sterile cotton plugs, and all experimental procedures were conducted under aseptic conditions in a clean bench. Except for the temperature-related experiments, all cultures were incubated at 30 °C and 150 r/min in a constant-temperature shaker. Three replicates were prepared for each treatment, together with a non-inoculated blank control. Ammonia nitrogen and nitrite nitrogen concentrations were measured every 12 h to evaluate the ammonia-oxidation performance. When required, the pH was adjusted to 7.0 using 1 mol/L NaOH and monitored using pH test paper.

2.5.2. Effect of CaCO₃ on the Nitrogen Removal Performance of N. europaea W4

Two treatment groups were established: one without CaCO₃ supplementation and the other with 5 g/L CaCO₃ supplementation. The initial ammonia nitrogen concentration in both groups was maintained at 100 mg/L. The effects of CaCO₃ on the ammonia-oxidizing performance of N. europaea W4 were evaluated by monitoring changes in ammonia nitrogen and nitrite nitrogen concentrations throughout the cultivation period.

2.5.3. Effects of Fe²⁺ and Fe³⁺ on the Nitrogen Removal Performance of N. europaea W4

The concentrations of both Fe²⁺ and Fe³⁺ were adjusted to 0.05 g/L. The effects of iron ions in different valence states on the ammonia-oxidizing performance of N. europaea W4 were compared.

2.5.4. Effect of Temperature on the Nitrogen Removal Performance of N. europaea W4

Five temperature conditions were established at 10, 15, 20, 30, and 40 °C. Following inoculation, the conical flasks were incubated at the respective temperatures with shaking at 150 r/min. The initial ammonia nitrogen concentration in all treatment groups was maintained at 100 mg/L.

2.5.5. Effect of Initial Ammonia Nitrogen Concentration on the Nitrogen Removal Performance of N. europaea W4

The initial ammonia nitrogen concentrations were adjusted to 100, 500, 1000, 1750, and 2500 mg/L. Changes in ammonia nitrogen and nitrite nitrogen concentrations were monitored throughout the cultivation period to evaluate the effects of different ammonia nitrogen concentrations on the ammonia-oxidizing performance of N. europaea W4.

2.6. Bench-Scale Test of N. europaea W4 in Aerobic Tank Wastewater

Wastewater from the aerobic tank at a municipal solid waste transfer station in Yichang, with a treatment capacity of 100 t/d, was selected as the target for treatment. The sampling point was located at the center of the aerobic tank, approximately 1.5 m below the water surface. A composite wastewater sample (25 L) was collected in September 2023 using a sterile sampler. The collected samples were filtered through 0.45 μm membrane filters before analysis. The main water quality parameters of the wastewater are presented in

Table 1. Water quality parameters of the landfill leachate.

Parameter	Concentration (mg/L)
Total nitrogen (TN)	388
Ammonia nitrogen (NH4+-N)	115
Total phosphorus (TP)	112
Chemical oxygen demand (COD)	5.79 × 103
Calcium ion (Ca2+)	31.2
Magnesium ion (Mg2+)	67

.

Raw non-sterilized landfill leachate was used to retain the indigenous microbial community. Collect 100 mL of culture broth of strains in exponential phase and harvest bacterial cells by centrifugation at 3000 g, then prepare bacterial suspension. The resulting cell pellet was resuspended to prepare a bacterial suspension, which was subsequently inoculated into a 250 mL conical flask containing 100 mL of landfill leachate. The flask was incubated at 30 °C and 50 r/min in the dark, with the dissolved oxygen level controlled at 2.0–3.0 mg/L by adjusting the shaking speed. During the experiment, the flask was manually shaken three times per day (1 min each time) to ensure homogeneous mixing. The cultures were incubated at 30 °C and 50 r/min in the dark. A non-inoculated control group was established simultaneously, and all treatments were conducted in triplicate. Ammonia nitrogen and nitrite nitrogen concentrations were measured every 12 h throughout the experiment.

2.7. Analytical Methods and Parameters

The main analytical parameters and corresponding methods are shown in

Table 2. Analytical methods for water quality indicators.

Indicator	Method
Ammonia nitrogen (NH4+-N)	Nessler’s reagent spectrophotometry
Nitrite (NO2−-N)	N-(1-naphthyl)-ethylenediamine spectrophotometry
Nitrate (NO3−-N)	Sulfamic acid-UV spectrophotometry
pH	Universal pH test paper
OD600	UV spectrophotometry

.

2.8. Statistical Analysis

The nitrite accumulation rate (V, mg/(L·h)) was calculated using Equation (1):

$$\mathrm{V} = {\displaystyle \frac{{\mathrm{C}}_{\mathrm{t}2} - {\mathrm{C}}_{\mathrm{t}1}}{{\mathrm{t}}_2 - {\mathrm{t}}_1}}$$

(1)

where C_t1 and C_t2 represent the nitrite concentrations (mg/L) at time t₁ and t₂, respectively. t₁, t₂: two adjacent sampling time points, in hours (h). The sampling interval in this study was 12 h. The specific growth rate (μ, d⁻¹) was calculated using Equation (2):

$$\mathsf{\mu } = {\displaystyle \frac{\ln {(\mathrm{V}}_2) -\ln {(\mathrm{V}}_1)}{{\mathrm{t}}_2 -{ \mathrm{t}}_1}}$$

(2)

where V₂ and V₁ represent the nitrite accumulation rates (mg/(L·h)) measured at days t₂ and t₁, respectively. t₁, t₂: two adjacent sampling time points, in days (d). The generation time (g, h) was calculated using Equation (3):

$$\mathrm{g} = {\displaystyle \frac{\ln 2}{\mathsf{\mu }}} \times 24$$

(3)

where μ represents the specific growth rate (d⁻¹).

ln2: the constant used for calculating doubling time.

24: conversion factor to convert specific growth rate from day⁻¹ to h⁻¹.

$$\surd \mathsf{\mu }={\mathrm{b}}_1(\mathrm{T} -{\mathrm{T}}_{\min }\left)\right[1 -\mathrm{e}\mathrm{x}\mathrm{p}({\mathrm{b}}_2(\mathrm{T}-{\mathrm{T}}_{\max }\left)\right)]$$

μm: Maximum specific growth rate without inhibition, unit: d⁻¹

b₂: Temperature coefficient in the suboptimal temperature range, value = 0.0377 √(d⁻¹)/°C

c₂: Temperature coefficient in the supraoptimal temperature range, value = 0.25 °C⁻¹

T: Incubation temperature, unit: °C

T_min: Theoretical minimum temperature for microbial growth, unit: °C

T_max: Theoretical maximum temperature for microbial growth, unit: °C

Figures and graphs were prepared using Origin, and the error bars represent the standard deviation (SD). Statistical analysis was performed using SPSS 20.0. Normality was assessed using the Shapiro–Wilk test, and homogeneity of variances was evaluated using Levene’s test. For data meeting the assumptions of normality and homogeneity of variance, analysis of variance (ANOVA) was conducted to evaluate statistical significance among different treatment groups, followed by Tukey’s HSD post hoc test for multiple comparisons.

3. Results

3.1. Isolation and Purification Results

Water samples were collected from the Shiqi River. Following enrichment cultivation at 100 mg/L initial ammonia nitrogen, pH 7, and 30 °C in the dark, a pure bacterial strain was successfully isolated using serial dilution combined with silica gel plate spreading techniques. When cultivated in a medium containing an initial ammonia nitrogen concentration of 100 mg/L for 72 h, the isolated strain achieved an ammonia nitrogen removal efficiency of 99.0%, accompanied by a nitrite accumulation of 99.2 mg/L, and an ammonia-to-nitrite conversion efficiency exceeding 99%, and no nitrate accumulation was detected throughout the cultivation process. The culture was subsequently transferred into beef extract peptone medium and incubated at 30 °C for 7 days. No colony growth was observed during incubation, confirming that the isolate was a pure culture with no detectable contamination. The strain was deposited at the China Center for Type Culture Collection under the deposit number CCTCCM 2024693.

3.2. Identification Results

3.2.1. Scanning Electron Microscopy Results

Field-emission scanning electron microscopy (FE-SEM) observation (Figure 1) revealed that cells of N. europaea W4 in the logarithmic growth phase exhibited short rod-shaped to nearly coccoid morphologies. The dimensions of individual cells ranged from approximately 1.0–1.5 μm in length and 0.2–0.4 μm in width, which are consistent with the reported morphological characteristics of the type strain N. europaea ATCC 19718.

3.2.2. 16S rRNA Sequencing Results

The 16S rRNA gene sequence (1486 bp) of strain W4 was obtained through PCR amplification and subsequently compared with sequences in the NCBI RefSeq database using the BLASTn program. The analysis revealed that strain W4 shared the highest sequence similarity (99.93%) with N. europaea ATCC 25978 (EMBL-EBI accession number: PRJNA245576). A phylogenetic tree constructed based on 16S rDNA sequences further demonstrated that strain W4 clustered within the same clade as N. europaea, while forming a clearly distinct lineage from Nitrosospira (Figure 2). Therefore, the isolate was identified as N. europaea and designated as N. europaea W4.

3.2.3. Genome Assembly and Prediction

The genome of N. europaea W4 consists of a circular chromosome (2,892,991 bp, G/C content 51.26%) and a circular plasmid (55,583 bp, G/C content 52.84%). At the genome level, it showed the highest similarity to N. europaea ATCC 19718, with an ANI value of 97.70% (ANI ≥ 95% indicates the same species). This confirms that W4 and ATCC 19718 belong to the same species. GC skew analysis revealed that the replication partition of the strain is nearly symmetrical (Figure 3), and no obvious contamination was detected.

3.3. Effects of Medium Components, Temperature, and Initial Ammonia Nitrogen Concentration on the Nitrogen Removal Performance of the Strain

3.3.1. Effect of CaCO₃ on the Nitrogen Removal Performance of N. europaea W4

As shown in Figure 4, no evident lag phase was observed in the CaCO₃-supplemented group. The initial ammonia nitrogen concentration of 100.76 ± 0.63 mg/L decreased to 25.56 ± 0.49 mg/L on day 3, whereas the nitrite nitrogen concentration increased from 1.24 ± 0.02 mg/L to 72.26 ± 2.01 mg/L over the same period. The maximum specific growth rate (μ_max) reached 1.42 d⁻¹ during days 2–3, and the logarithmic growth phase was observed from day 1 to day 3. By day 4, ammonia nitrogen concentration had declined to 1.30 ± 0.98 mg/L, indicating that ammonia oxidation was nearly complete. The final nitrite nitrogen concentration reached 98.59 ± 0.46 mg/L on day 10. In the group without CaCO₃ supplementation, no significant change in ammonia nitrogen concentration was observed during the first 7 days. After the culture entered the logarithmic growth phase on day 8, ammonia nitrogen concentration rapidly decreased from 86.39 ± 1.15 mg/L to 2.30 ± 1.06 mg/L by day 10. Simultaneously, the nitrite nitrogen concentration increased from 13.90 ± 2.11 mg/L to 99.60 ± 0.69 mg/L. The maximum specific growth rate (μ_max) reached 1.37 d⁻¹ during days 9–10. Although no significant difference in the final nitrite nitrogen concentration was observed between the two treatment groups, the addition of CaCO₃ shortened the lag phase by approximately 7 days, indicating that CaCO₃ supplementation markedly accelerated the initiation of ammonia oxidation by N. europaea W4.

3.3.2. Effects of Fe²⁺ and Fe³⁺ on the Nitrogen Removal Performance of N. europaea W4

As shown in Figure 5, the strain exhibited high ammonia-oxidizing activity in both Fe²⁺ and Fe³⁺ -supplemented media, although ammonia nitrogen consumption occurred slightly faster in the Fe²⁺ treatment group. A lag phase of approximately 1 day was observed in the Fe³⁺ group. Subsequently, the ammonia nitrogen concentration decreased from 109.28 ± 0.41 mg/L on day 1 to 2.15 ± 0.01 mg/L on day 5. In parallel, the nitrite nitrogen concentration increased from 5.26 ± 0.03 mg/L to 105.39 ± 0.01 mg/L. The maximum specific growth rate (μ_max) in the Fe³⁺ group reached 0.865 d⁻¹ during days 2–3, and complete ammonia oxidation required approximately 5 days. In contrast, no apparent lag phase was observed in the Fe²⁺ treatment group. The ammonia nitrogen concentration decreased from 109.94 ± 0.64 mg/L to 90.30 ± 0.44 mg/L on day 1 and further declined to 4.12 ± 0.74 mg/L by day 4. Meanwhile, the nitrite nitrogen concentration increased from 3.25 ± 0.01 mg/L to 20.89 ± 0.94 mg/L on day 1 and remained stable at 108.50 ± 0.12 mg/L during days 4–6. The maximum specific growth rate (μ_max) reached 0.773 d⁻¹ during days 1–2, and complete ammonia oxidation was achieved within only 4 days.

3.3.3. Effect of Temperature on the Nitrogen Removal Performance of N. europaea W4

As shown in Figure 6, within the tested temperature range of 10–40 °C, the strain exhibited the highest ammonia-oxidizing activity at 30 °C. No apparent lag phase was observed under this condition. The ammonia nitrogen concentration decreased from 124.98 ± 0.03 mg/L to 98.9 ± 0.12 mg/L within 24 h. In contrast, negligible growth and ammonia oxidation activity were detected at the other tested temperatures. Regarding nitrite accumulation, the final nitrite nitrogen concentrations at 72 h reached 119.26 ± 0.01 mg/L at 20 °C and 123.45 ± 0.01 mg/L at 30 °C, with ammonia nitrogen conversion efficiencies exceeding 99% under both conditions. In contrast, only 6.02 ± 0.11 mg/L of nitrite nitrogen accumulated at 10 °C, indicating that the strain exhibited minimal proliferation and ammonia oxidation activity at low temperature. At 30 °C, more than 99% of the ammonia nitrogen was converted into nitrite within 60 h. During this period, the ammonia nitrogen concentration decreased from 124.98 ± 0.03 mg/L to 1.23 ± 0.01 mg/L, whereas the nitrite nitrogen concentration increased from 5.47 ± 0.12 mg/L to 122.45 ± 0.18 mg/L. The maximum specific growth rate reached 0.1038 h⁻¹, corresponding to a minimum generation time of 6.7 h. Fitting with the Ratkowsky 2 model showed a good correlation (R² = 0.95, p < 0.05) and indicated that the optimal temperature for nitrite accumulation was approximately 27 °C. In comparison, the generation time at 15 °C increased substantially to 21.7 h.

3.3.4. Effect of Different Ammonia Nitrogen Concentrations on the Nitrogen Removal Performance of N. europaea W4

The effects of different initial ammonia nitrogen concentrations on the ammonia-oxidizing performance of N. europaea W4 are presented in Figure 7. In the treatment groups containing 500, 1000, and 1750 mg/L ammonia nitrogen, nitrite nitrogen accumulation was significantly higher than that observed in the other groups (p < 0.05). The corresponding nitrite nitrogen concentrations on day 9 reached 433.01 ± 0.32, 727.28 ± 0.89, and 760.36 ± 0.56 mg/L, respectively. These results indicate that this concentration range was the most favorable for ammonia-oxidation activity. Fitting with the Edwards 2 model demonstrated a strong correlation (R² = 0.94, p < 0.05) and suggested that the optimal initial ammonia nitrogen concentration for nitrite accumulation was approximately 2000 mg/L. Furthermore, the strain maintained high ammonia-oxidizing efficiency within the ammonia nitrogen concentration range of 1500–2000 mg/L.

Under the experimental conditions of pH 7.0 and 30 °C, the corresponding free ammonia (FA) concentrations were 0.80, 3.98, 7.96, 13.94 and 19.91 mg/L at initial ammonia nitrogen levels of 100, 500, 1000, 1750 and 2500 mg/L, respectively. As the initial ammonia nitrogen concentration increased, the strain’s maximum specific growth rate increased. The maximum specific growth rate observed in the 100 mg/L treatment group (0.92 ± 0.17 d⁻¹) was significantly lower than that measured in the other treatment groups (p < 0.05). In contrast, the maximum specific growth rates in the 1750 and 2500 mg/L groups were significantly higher than those in the low-concentration treatments (p < 0.05). The Edwards 2 model fitting further revealed that the effects of ammonia nitrogen concentration on ammonia oxidation activity and cellular growth in N. europaea W4 were not fully synchronized.

3.4. Bench-Scale Test of N. europaea W4 in Aerobic Tank Wastewater

As shown in Figure 8, inoculation of N. europaea W4 into landfill leachate significantly enhanced ammonia nitrogen degradation. The average ammonia oxidation rate in the inoculated group reached 29.32 ± 0.07 mg·L⁻¹·d⁻¹, whereas the control group exhibited a substantially lower rate of only 4.64 ± 0.11 mg·L⁻¹·d⁻¹. No nitrite accumulation was detected in the control group throughout the experimental period. In contrast, the inoculated group showed continuous nitrite accumulation over the first 4 days, with nitrite nitrogen concentrations increasing from 29.54 ± 0.03 mg/L to a peak of 45.40 ± 0.07 mg/L on day 4. Subsequently, the nitrite concentration rapidly declined following the depletion of ammonia nitrogen.

4. Discussion

The isolation of the N. europaea W4 strain enriches the existing resource pool of ammonia-oxidizing bacteria (AOB) and provides valuable experimental material for investigating AOB diversity in freshwater environments in South China. Compared with the previously reported N. europaea ATCC 19718 [15], N. europaea W4 exhibits short rod-shaped to subspherical cellular morphology and shows highly phylogenetic similarity, with a 16S rRNA sequence similarity of 99.93%. However, strain W4 demonstrates distinct advantages in ammonia oxidation performance and may possess unique metabolic regulatory characteristics and environmental adaptability [16]. Phylogenetic analysis demonstrated that N. europaea W4 clustered within the same clade as N. europaea and was clearly separated from Nitrosospira strains. This finding not only confirms the taxonomic classification of strain W4 but also reflects the functional divergence that has occurred between Nitrosomonas and Nitrosospira during evolutionary adaptation processes. Nitrosomonas species are generally adapted to environments with high ammonia nitrogen concentrations, whereas Nitrosospira species preferentially inhabit environments with relatively low ammonia nitrogen levels [17]. N. europaea W4 demonstrated high ammonia oxidation efficiency in a medium containing an initial ammonia nitrogen concentration of 100 mg/L. Within 72 h, the strain achieved an ammonia nitrogen removal efficiency of 99% and a nitrite nitrogen conversion efficiency exceeding 99%. Notably, no nitrate accumulation was detected throughout the process. This observation is consistent with the typical metabolic characteristics of ammonia-oxidizing bacteria (AOB), which oxidize ammonia to nitrite without further conversion to nitrate under the experimental conditions. The ammonia oxidation-related gene clusters of N. europaea W4 are located within highly conserved LCBs, and their arrangement order is completely consistent with that of N. europaea ATCC 19718. This indicates that the ammonia oxidation function in N. europaea W4 exhibits a high degree of evolutionary conservation, representing an important manifestation of core metabolic functions.

Scanning electron microscopy (SEM) observations revealed that N. europaea W4 exhibited a short-rod-shaped to subspherical morphology with irregularly wrinkled cell surfaces. No cellular appendages, such as flagella, fimbriae, or capsules, were observed. These morphological characteristics may be related to the strain’s planktonic growth state in the liquid medium. In addition, the cells tended to aggregate into clusters rather than disperse uniformly throughout the medium, suggesting that N. europaea W4 possesses an inherent propensity for cell-to-cell aggregation, the specific mechanism of which requires further investigation.

CaCO₃ supplementation significantly reduced the lag phase of N. europaea W4 from 6 days to 0 days, while exerting no substantial effect on the maximum specific growth rate (μ_max). The μ_max values were 1.42 d⁻¹ and 1.37 d⁻¹ in the CaCO₃-supplemented and non-supplemented groups and 1.37 d⁻¹ in the non-supplemented group. The well-known conventional functions of CaCO₃ (neutralizing acidity and providing an inorganic carbon source) cannot explain this decoupling phenomenon, where the lag phase is significantly shortened while μ_max remains essentially unchanged. Wang et al. [18] reported that the combined application of calcium carbonate with specific biochar and organic fertilizers effectively stimulated soil nitrification activity, suggesting that CaCO₃ may exert biological functions beyond simple acid–base neutralization. Furthermore, Kolodkin-Gal et al. [19] noted that Ca²⁺ not only serves as a structural component involved in biomineralization but also acts as a signaling molecule that regulates gene expression and biofilm formation. Based on the above experimental results and previous studies, we propose the following working hypothesis: the mechanism by which CaCO₃ shortens the lag phase of AOB involves the dual effects of physical enrichment and calcium-mediated signaling. On one hand, CaCO₃ particles can provide attachment sites for bacterial cells and accelerate local microbial enrichment; on the other hand, Ca²⁺ released from CaCO₃ dissolution may act as a signaling molecule to independently initiate early-stage colonization and coordinated expression of functional genes, thereby shortening the time required for the establishment of community function. This hypothesis explains why the lag phase is significantly shortened while μ_max is unaffected, as Ca²⁺ primarily regulates community-level initiation behaviors rather than the metabolic growth rate of individual cells. It should be clearly stated that the above mechanism regarding Ca²⁺-mediated signal transduction is currently a speculative conclusion. The potential interaction between Ca²⁺ signaling and the quorum-sensing system of AOB, as well as the specific contribution of CaCO₃ in complex systems, still requires further verification through molecular biological methods such as gene expression analysis and signal molecule detection. Our experimental results clearly confirm that CaCO₃ primarily plays a priming-promoting role rather than a growth-enhancing role in the regulation of nitrifying microorganisms, providing important experimental evidence for understanding AOB growth kinetics and optimizing high-ammonia-nitrogen wastewater treatment processes.

The experimental results demonstrated that the lag phase in the Fe²⁺ treatment group was approximately 1 day shorter than that in the Fe³⁺ treatment group. In addition, the ammonia nitrogen consumption rate in the Fe²⁺-supplemented medium was slightly higher than that observed in the Fe³⁺ medium. Since the transformation and stability of iron oxidation states were not experimentally monitored in this study, the following analysis is proposed as a reasonable speculation based on existing physiological results. Fe²⁺ can be directly absorbed by AOB and incorporated into the active centers of ammonia monooxygenase (AMO) and hydroxylamine oxidoreductase (HAO). In contrast, Fe³⁺ must first undergo chelation, transport, and reduction to Fe²⁺ via siderophore-mediated pathways before it can be utilized. This reduction process imposes an additional energy burden on the cells [20]. The genome of N. europaea encodes a wide range of proteins associated with iron acquisition and transport [4]. The activities of AMO and HAO depend on copper/iron cofactors and heme iron cofactors, respectively. Therefore, iron availability directly affects ammonia oxidation efficiency. Under iron-limited conditions, the ammonia-dependent oxygen consumption activity of AOB decreases significantly [21]. Moreover, the transport and reduction processes required for Fe³⁺ utilization further increase the metabolic burden on the cells [22]. Consequently, under Fe³⁺ supplementation conditions, the strain is likely to allocate additional energy to Fe³⁺ reduction, leading to a lower ammonia oxidation rate and an extended lag phase [5]. Under both iron source treatments, nitrite nitrogen remained the primary end product of the AOB ammonia oxidation pathway in AOB [21],The presence of Fe²⁺ mainly enhanced the rate of nitrite nitrogen accumulation without altering the nature of the metabolic end product.

The optimal growth temperature range for N. europaea W4 was determined to be 25–30 °C. The strain still maintained basal metabolic activity at 15 °C; however, growth arrest occurred at 10 °C, indicating that further acclimation may be required to improve low-temperature tolerance. Temperature influences bacterial growth and metabolism primarily by affecting membrane fluidity, DNA replication, protein synthesis, and enzyme activity. In general, elevated temperatures enhance enzymatic activity, whereas reduced temperatures suppress metabolic processes [23]. Compared with typical AOB strains, N. europaea ATCC 19718 exhibits an optimal growth temperature range of 28–32 °C [24], whereas N. eutropha CZ-4 has an optimal growth temperature of 30.9 °C and is capable of surviving at 40 °C [25,26]. Both strains therefore demonstrate stronger high-temperature adaptability than N. europaea W4. The suitable growth temperature range for AOB is generally considered to be 20–30 °C. The lower temperature tolerance limit of N. europaea W4 was approximately 5 °C lower than that reported for typical AOB strains, indicating that W4 possesses a relatively broad temperature adaptation range within the 15–30 °C. These batch experimental results demonstrate that N. europaea W4 shows preliminary adaptability to temperatures in the 15–30 °C range; however, its engineering applicability for treating fluctuating-temperature wastewater requires further validation through continuous-flow experiments and pilot-scale tests.

N. europaea W4 exhibits low affinity toward ammonia and primarily maintains intracellular homeostasis through a high ammonia metabolic rate. When the initial ammonia nitrogen concentration increased to 2500 mg/L, the lag phase was prolonged to 72 h, approximately 48 h longer than in the moderate- and low-concentration treatment groups (100–1750 mg/L). In addition, the maximum specific growth rate decreased to 2.85 d⁻¹. This phenomenon is associated with multiple stress effects induced by high ammonia nitrogen concentrations, including disruption of intracellular ion homeostasis [27], free ammonia concentrations exceeding the inhibitory threshold for AMO activity, and increased accumulation of hydroxylamine, a by-product of ammonia oxidation that can trigger feedback inhibition [28]. Meanwhile, ammonia transporters encoded by the strain may enhance cellular tolerance to ammonia toxicity [29]. Edwards model fitting (R² = 0.94) indicated that the optimal ammonia nitrogen concentration for N. europaea W4 was approximately 1987 mg/L, which is substantially higher than those reported for related strains. For example, the inhibition constant (K_i) of N. europaea SH-3 is 922.76 mg/L, whereas those of N. eutropha CZ-4 and N. halophila C-19 are 597.88 mg/L and 186.24 mg/L, respectively [30]. Similarly, the optimal ammonia nitrogen concentration for the N. mobilis NZ13 enrichment culture is only 200–400 mg/L, with a half-maximal inhibitory concentration of 840 mg/L [25]. In addition, strains belonging to N. cluster 7 achieve their maximum growth rates at ammonium concentrations of ≥5 mM (approximately 70 mg/L NH₄⁺) [31].

Free ammonia (FA) concentrations below 23.64 mg NH₃-N/L have been reported to promote the proliferation of AOB and stimulate extracellular polymeric substance secretion. In contrast, FA concentrations exceeding 687.1 mg/L can completely inhibit microbial activity [32,33]. N. europaea W4 may therefore possess unique adaptive mechanisms that enable tolerance to high nitrogen loading conditions during its evolutionary development. N. europaea W4 can achieve an ammonia oxidation load of 1.28–1.55 kg-N·m⁻³·d⁻¹ at ammonia nitrogen concentrations ranging from 1500 to 2000 mg/L. Under ammonia nitrogen concentrations ranging from 1500 to 2000 mg/L, strain W4 achieved an ammonia oxidation load of 1.28–1.55 kg-N/m³/d. In future engineering applications, when treating wastewater with ammonia nitrogen concentrations exceeding 2500 mg/L, measures such as staged dilution or pH regulation may be considered to keep the free ammonia (FA) concentration below the inhibitory threshold, thereby ensuring efficient operation of the strain.

N. europaea W4 exhibited measurable ammonia oxidation capacity in landfill leachate. In the inoculated treatment group, the ammonia nitrogen degradation rate was significantly enhanced, and the nitrite nitrogen concentration reached a maximum value of 45.40 mg/L. In contrast, no detectable nitrite accumulation was observed in the control group following ammonia nitrogen degradation. The high chemical oxygen demand (COD) environment (5.79 × 10³ mg/L) did not completely inhibit the strain’s activity. This observation may be attributable to the inherent resistance mechanisms of autotrophic AOB toward elevated organic matter concentrations [34]. Park and Noguera reported that certain AOB strains can reduce the toxicity of organic compounds by expressing efflux pump proteins [35,36]. This capability is particularly important for the biological treatment of complex wastewater matrices such as landfill leachate. The rapid decline in nitrite concentration following ammonia nitrogen depletion is likely due to synergistic interactions among multiple microbial groups. Future studies employing high-throughput sequencing technologies could elucidate the underlying mechanisms. Notably, the primary objective of this study was to evaluate the tolerance of the AOB strain W4 in actual wastewater, with a focus on its ammonia-nitrogen conversion function. During the experiment, the decreasing trend of ammonia nitrogen and the concurrent production of nitrite nitrogen were significant and well-correlated, which fully reflected the ammonia oxidation activity of the strain and strongly supported the core conclusions of this study. Therefore, nitrate was not included as a measurement indicator in this study, and this limitation has been explicitly stated in the limitations section of the manuscript. Additionally, long-term continuous reactor trials are warranted to evaluate the stability of ammonia oxidation by strain W4 under practical operating conditions.

5. Conclusions

In the present study, a highly efficient ammonia-oxidizing bacterial strain was successfully isolated from the Shiqi River. Based on morphological characterization, Gram staining, scanning electron microscopy, and 16S rDNA sequence analysis, the isolate was identified as N. europaea and designated N. europaea W4. Under an initial ammonia nitrogen concentration of 100 mg/L, strain W4 achieved an ammonia nitrogen removal efficiency of 99% and a nitrite nitrogen conversion efficiency exceeding 99% within 72 h. The cells of N. europaea W4 exhibited short rod-shaped to subspherical morphology, with wrinkled cell surfaces and no observable appendages. In addition, the cells tended to aggregate into clusters. Medium optimization experiments demonstrated that supplementation with 5 g/L CaCO₃ significantly reduced the strain’s lag phase. The optimal growth temperature range was determined to be 25–30 °C. Although the strain maintained basal metabolic activity at 15 °C, complete growth inhibition occurred at 10 °C. Furthermore, the optimal initial ammonia nitrogen concentration was approximately 1987 mg/L, and efficient high-load ammonia treatment could be achieved within the 1500–2000 mg/L range. The optimal initial ammonia nitrogen concentration was approximately 1987 mg/L, and a high treatment load could be achieved within the range of 1500–2000 mg/L. Pilot-scale experiments using landfill leachate demonstrated that inoculation with strain W4 increased the ammonia nitrogen degradation rate by approximately 6.3-fold. In addition, the peak nitrite nitrogen concentration reached 45.40 mg/L. It should be noted, however, that these results were obtained from bench-scale shaken-flask experiments, and the engineering applicability of the strain requires further validation in pilot-scale continuous-flow treatment systems.