A Review of Plant-Mediated and Fertilization-Induced Shifts in Ammonia Oxidizers: Implications for Nitrogen Cycling in Agroecosystems

Abstract

Nitrogen (N) cycling in agroecosystems is a complex process regulated by both biological and agronomic factors, with ammonia-oxidizing archaea (AOA) and bacteria (AOB) playing pivotal roles in nitrification. Despite extensive fertilizer applications to achieve maximum crop yields, nitrogen use efficiency (NUE) remains less than ideal, with substantial losses contributing to environmental degradation. This review synthesizes current knowledge on plant-mediated and fertilization-induced shifts in ammonia-oxidizer communities and their implications on nitrogen cycling. We highlight the differential ecological niches of AOA and AOB, emphasizing their responses to plant community composition, root exudates, and allelopathic compounds. Fertilization regimes of inorganic nitrogen inputs and biological nitrification inhibition (BNI) are examined in the context of microbial adaptation and ammonia tolerance. Our review highlights the need for integrated nitrogen management strategies comprising optimized fertilization timing, nitrification inhibitors, and plant–microbe interactions in order to optimize NUE and mitigate nitrogen losses. Future research directions must involve applications of metagenomic and isotopic tracing techniques to unravel the mechanistic AOA and AOB pathways that are involved in regulating these dynamics. An improved understanding of these microbial interactions will inform the creation of more sustainable agricultural systems that aim to optimize nitrogen retention and reduce environmental footprint.

1. Introduction

The global nitrogen (N) cycle is a fundamental biogeochemical process that regulates the availability of N in various forms, essential for all living organisms [1]. Nitrogen, a component of amino acids, proteins, and nucleic acids, is indispensable for agricultural crop growth and productivity. Critically, soil microorganisms mediate most of the transformations in the N cycle, such as N fixation, mineralization, nitrification, and denitrification, thus governing how much plant-available N remains in the soils [2]. However, this global N cycle has been significantly altered by human activities, particularly through the use of synthetic fertilizers and fossil fuel combustion, leading to an increase in N in terrestrial systems [3]. Anthropogenic production of N has increased dramatically over the past ~150 years, rising from about 15 Tg N yr⁻¹ in the 1860s to over 220 Tg N yr⁻¹ in recent decades [1], with 120 Tg N yr⁻¹ of this from fertilizers [2]. This anthropogenic input has increased by about 5% annually over the past decades, driven by the escalating need to enhance food production to support the expanding world population [1]. Despite the extensive use of fertilizers in agricultural systems, the nitrogen use efficiency (NUE) remains suboptimal. Typically, crops absorb only about 50% or less of the applied N fertilizer during the growing season [4,5]. The resultant disruption of the N cycle has profound environmental and ecological consequences. Anthropogenic pressures have led to potentially irreversible changes in terrestrial ecosystems, including global warming, elevated carbon dioxide levels, variations in rainfall, nutrient enrichment, and combinations of these factors [6]. These alterations can modify plant growth, soil microclimatic conditions, soil substrate availability, and the structure and function of soil microbial communities, profoundly impacting soil N cycling [7]. A comprehensive understanding of N dynamics within agroecosystems is crucial for enhancing sustainable agricultural productivity and minimizing negative environmental consequences.

Nitrogen exists in various oxidation states and chemical forms and can be transformed on timescales of hours to days in both terrestrial and marine environments [8,9]. Previously, the global N cycle was considered to be approximately linear, i.e., moving from the atmosphere and soil. The largest component of global N is triple-bonded N₂, comprising 78% of the atmosphere, which must be “fixed”—i.e., converted into bioavailable forms, such as ammonia (NH₃), by microorganisms in order to become accessible to other organisms [10]. Nitrogen cycling encompasses four microbially mediated biochemical reactions: N-fixation, mineralization, nitrification, and denitrification [11]. The application of substantial amounts of N fertilizers to agricultural fields significantly affects these processes, particularly nitrification and denitrification, leading to increased N₂O production [12].

In soils, nitrification converts NH₃, produced from the decomposition and ammonification of organic matter or added as NH₃-containing fertilizers, into the oxidized N species nitrite (NO₂⁻) and nitrate (NO₃⁻) [13]. Ammonia oxidation, considered to be the rate-limiting step of nitrification, is catalyzed by ammonia monooxygenase (AMO), which is encoded by the amoA gene [14]. Ammonium (NH₄⁺) exists in a pH and moisture-dependent equilibrium with NH₃, which is converted into NO₂⁻, some of which can be converted to NO₃⁻. This conversion occurs either in two steps through a mutualistic symbiosis involving AOB, AOA, and NOB, or directly by the recently discovered complete ammonia-oxidizing bacteria (comammox) [14,15,16].

Ammonia oxidizing bacteria initiate ammonia oxidation by converting NH₃ to hydroxylamine (NH₂OH) utilizing the enzyme AMO [11]. This enzyme is membrane-bound and consists of three subunits: amoA, amoB, and amoC [17]. The subsequent oxidation of NH₂OH to NO₂⁻ is catalyzed by hydroxylamine oxidoreductase (HAO), which contains a ferrous heme P460 active site [18]. Recent studies suggest that nitric oxide (NO) might also be an intermediate in this pathway, necessitating a third, as yet unidentified, enzyme to convert NO to NO₂⁻ [19]. Ammonia oxidizing archaea possess homologs of the bacterial AMO subunits, along with a potential fourth subunit, amoX [20,21]. Unlike AOB, AOA lack HAO homologs, but still produces and consume NH₂OH during NH₃ oxidation, indicating that it is a true intermediate [22]. The exact mechanism remains unclear, but it is hypothesized that alternative oxidoreductases, such as multi-copper oxidases (MCOs) or the cofactor F420, may play a role [23,24]. Nitrous oxide is also essential to AOA for the oxidation of NH₂OH, likely to be produced by a nitrite reductase, like nirK [24]. Understanding the distinct ammonia oxidation pathways in AOB, AOA, and comammox bacteria is crucial for developing effective nitrification inhibitors and optimizing N use in agricultural soils. Each group’s unique enzymatic mechanisms and ecological roles highlight the complexity of the N cycle and the need for targeted mitigation strategies to reduce N losses and environmental impacts.

Recent advances in microbial ecology have revealed that complete ammonia-oxidizing (comammox) bacteria, primarily belonging to the Nitrospira genus, can perform the full nitrification process from NH₃ to NO₃⁻ within a single organism. This challenges the long-held paradigm that nitrification is always a two-step process mediated by separate functional groups: AOA and AOB for ammonia oxidation, and NOB for nitrite oxidation [15,16]. Comammox organisms possess the full complement of amo and nxr genes, enabling them to couple ammonia and nitrite oxidation under aerobic conditions.

The main factors controlling the abundance and activity of AOA and AOB in agricultural soils include substrate availability, plant community composition, fertilization practices, and the interplay between AOA and AOB [25]. These factors influence nitrifying organism both directly—at the cellular level—and indirectly—by modifying the soil environment [25]. For instance, plants profoundly influence ammonia-oxidizing microbial communities through root exudation patterns, alterations in rhizosphere pH, and allelopathic interactions, critically shaping within agroecosystems. Root exudates, including organic acids, sugars, amino acids, and phenolic compounds, selectively modulate ammonia oxidizer populations by providing specific substrates and altering microbial habitat conditions [26,27]. Moreover, root-induced shifts in rhizosphere pH profoundly impact the bioavailability of NH₃, directly influencing ammonia oxidation rates and niche differentiation between AOA and AOB [28,29]. Plants also release allelopathic secondary metabolites, such as isothiocyanates and sorgoleone, which specifically inhibit AOB activity by disrupting AMO enzymatic pathways, thereby acting as biological nitrification inhibitors and altering soil nitrification dynamics [30,31]. Consequently, understanding plant-driven mechanisms and their selective influences on ammonia oxidizers is crucial for improving NUE, reducing N losses, and informing sustainable agricultural management practices.

The effects of these factors over time range from rapid changes, such as nitrification rates occurring within minutes to hours, to long-term alterations spanning years to decades, e.g., the effects of coupling to the soil carbon (C) cycle. Significant interactions and feedback mechanisms exist among these controlling factors, as the populations of active nitrifying microbes are not only determined by substrate availability but also shaped by soil properties (e.g., pH, moisture, and structure) and management practices (e.g., tillage, irrigation, and crop rotation) that influence microbial habitat conditions.

This review aims to synthesize current knowledge on how AOA and AOB respond to plant-driven processes and fertilization practices in agroecosystems. Specifically, the review examines plant-mediated shifts in microbial communities through mechanisms such as root exudation, allelopathic interactions, and differences in plant functional traits and resource-use strategies. It further explores how various fertilization regimes, including inorganic N applications, nitrification inhibitors, and biological nitrification inhibition (BNI), influence the community structure and ecological roles of AOA and AOB. This review also focuses on agronomic strategies aimed at optimizing NUE and mitigating environmental impacts through improved management of microbial nitrification processes. By consolidating these aspects, the review provides an integrated perspective essential for developing sustainable N management strategies in agricultural systems.

2. Above Ground Plant Community Composition’s Influence on Ammonia Oxidizers

Ammonia oxidizing archaea and AOB coexist across most terrestrial ecosystems, but differ in abundance and function, reflecting distinct ecological niches [32]. An important factor shaping these niche specializations in agricultural systems is the composition of plant communities. For instance, crop diversity influences soil microbes, as each plant species uniquely shapes the microbial community. Plants profoundly impact soil characteristics, e.g., nutrient dynamics, rhizosphere pH, and soil structure, via root activity and plant tissue decomposition, thereby affecting microbial community structure [33]. Multiple studies indicate that different plant species can shape soil microbial communities through root exudates and functional traits, though the mechanisms and extent are still being studied [34].

2.1. Differences in Ammonia Oxidizers Due to Plant Composition

Understanding plant functional traits is key to enhancing beneficial plant–microbe interactions in agroecosystems. Diverse plant communities can enhance primary productivity and increase belowground C and N inputs, thereby influencing AOA and AOB abundance. Consequently, it has been demonstrated that plant communities can modify N cycling by influencing the abundance and activity of ammonia-oxidizing microbial communities in the soil [14,35,36]. Ammonium availability driven by plant functional groups plays a key role in shaping AOB community composition, highlighting the influence of legumes on nitrification dynamics in soil microbial ecosystems. For instance, Malchair et al. (2010) found AOB structure differed under legumes versus other groups, especially after 16 months [37]. These shifts reflect changes in Nitrosospira, with cluster 3b (NH₄⁺- tolerant) lost under forbs and grasses, and 3a (NH₄⁺- sensitive) absent in legumes and bare soils. Furthermore, Thion et al. (2016) reported that AOA communities were positively linked to N rich grass species, such as Lolium perenne and Festuca rubra [38]. In contrast, AOB groups’ abundance was not influenced by plant groups. No facilitative or antagonistic interactions were observed, indicating independent regulation by plant and soil factors.

2.2. Root Exudates and Rhizosphere Dynamics

Plants alter soil microbiomes through the secretion of root exudates. These exudates provide C and energy to microbes, shaping community structure and function. The vast diversity of plant species and their root-derived secondary metabolites plays a critical role in shaping below-ground microbial communities [39,40]. Plant roots release ~40% of photosynthates as exudates, including sugars, amino acids, organic acids, phytohormones, proteins, and mucilage [41].

Antimicrobial metabolites, like phenolics (e.g., flavonoids, tannins) and amino sugars, play key roles in shaping root-associated microbiota and extracellular enzyme activity [26]. Phenolic compounds influence N cycling by recruiting specific microbial taxa and, in some cases (e.g., Fallopia japonica) [42], inhibiting nitrification to reduce nitrate availability for competing plants [43]. Additionally, phenolic acids released in large quantities during litter turnover can impact the persistence and symbiotic efficiency of rhizobia in the soil, affecting N fixation in legumes [44]. Furthermore, changes in root exudation rates due to N additions can lead to alterations in carbon efflux, soil organic matter decomposition, and nitrogen cycling in alpine shrub ecosystems [45]. Organic acids, like oxalic acid, enhance N mineralization by mobilizing N-bound compounds from clay surfaces [46]. Additionally, they can influence the release of dissolved organic carbon (DOC) and alter bacterial communities, thus affecting organic matter decomposition and nutrient cycling in the rhizosphere [47]. Ai et al. (2015) showed that maize exudates favored AOB, while tomato and cucumber promoted AOA [27]. Huang et al. (2014) found similar patterns in wheat, linked to specific exudate components, like organic and amino acids [48].

2.3. Allelopathic Compounds

Beyond exudates, allelopathic secondary metabolites released by plants can modulate nitrification by inhibiting or stimulating ammonia oxidizers. Secondary metabolites are small molecular metabolites, produced by non-essential metabolic pathways [49]. In plants, secondary metabolites (PSMs) are derived from primary metabolites or their biosynthetic intermediates. Structurally diverse, PSMs are categorized into major molecular families based on their biosynthetic origins, including phenolics, terpenes, steroids, alkaloids, and flavonoids [50]. Though not essential for plant survival, PSMs serve key ecological and physiological functions. Zakir et al. (2008) found that isothiocyanates from brassicaceous residues inhibited AOB and nitrification by targeting the AMO enzyme, similar to methyl 3-(4-hydroxyphenyl) propionate (MHPP) but unlike synthetic inhibitors that do not affect HAO [51]. This suggests differential tolerance mechanisms, with AOA exhibiting lesser impacts. Tesfamariam et al. (2014) similarly reported AOB inhibition by sorgoleone, a benzoquinone-rich compound in sorghum exudates [31]. This inhibition likely alters N cycling in sorghum systems by affecting nitrifier activity under natural conditions and different N regimes.

2.4. Plant–Microbiome Signaling

Plants emit volatile organic compounds (VOCs) as defenses against herbivory, pests, and pathogens (Figure 1) [52]. These VOCs mediate belowground signaling and may recruit beneficial microbes to support plant health [53]. Volatile organic compounds, characterized by their small lipophilic structures, high vapor pressure, and low boiling points, diffuse efficiently through water and gas filled pores [52,54]. These properties allow VOCs to function as chemical messengers in plant–microbe interactions [52]. These compounds act as antimicrobials, nutrients, attractants, or signals, shaping soil microbial diversity and function [52]. For instance, signaling between heterotrophic and autotrophic AOB enhances cytokinin transport, promoting Italian ryegrass growth. Heterotrophic AOB strains, by modulating soil N dynamics and cytokinin pathways within the rhizosphere, significantly increase cytokinin synthesis and its subsequent translocation to the aerial parts of the plant. This mechanism is pivotal for promoting regrowth and biomass accumulation in plants such as Italian ryegrass, with potential implications for other agricultural crops. Further research is necessary to elucidate the detailed pathways and interactions at play, particularly the roles of HAOB in the soil–rhizosphere environment [55].

2.5. Plant Nitrophily Gradients

Plant species exhibit varying affinities for N, known as nitrophily, allowing them to be categorized along an oligotrophic to nitrophilic gradient [56]. Oligotrophic species, often found in N-limited soils, exhibit minimal growth response to increased soil N [57], and tend to allocate C more to root biomass, contrasting with nitrophilic species that prioritize C allocation to root activity, such as N uptake. These variations in nitrophily underscore the functional diversity among plant species in N utilization strategies, thereby driving the composition and function of ammonia-oxidizing microbial communities [58].

2.6. Photosynthetic Pathways

The divergence in photosynthetic mechanisms between C4 (four C compounds through the Calvin cycle) and C3 (three C compounds through the Calvin cycle) plants affects rhizosphere microbes, driven by distinct root exudation patterns [59]. C4 plants, which have a higher N use efficiency and lower photorespiratory rates compared to C3 plants, tend to release a smaller number of root exudates [60,61]. Additionally, C4 exudates are rich in organic acids (e.g., malate, oxalate), whereas C3 exudates contain more sugars (e.g., maltose, mannose, ribose) [62,63]. Several studies have reported that wheat roots exhibit higher C efflux through respiration compared to corn roots [64]. C4 plants (e.g., corn, sugarcane, sorghum) display a consequence of their enhanced photosynthetic efficiency and minimal photorespiration. In contrast, C3 crop plants (e.g., wheat, rye, oats, rice, cotton, sunflower) exhibit lower net C assimilation due to energy costs associated with photorespiration [65]. These traits suggest that C4 plants may harbor more diverse and abundant rhizosphere microbes than C3 plants [66].

In a study by Moreau et al. (2015), eleven plant species were examined across an oligotrophic–nitrophilic gradient to explore the role of nitrogen-related traits in plant–microbe interactions [56]. Among these, Echinochloa crus-galli (L.) Beauv. stood out with a distinctly different rhizosphere bacterial community, likely attributed to its C4 photosynthetic pathway. Overall, it was concluded that plant N-use traits can shape rhizosphere microbial structure. Nevertheless, certain plant species have been observed to adversely affect N cycling microorganisms. These effects are attributed to inhibitory compounds present in plant exudates, rather than nutrition traits, resulting in decreased abundance or activity of nitrifying microorganisms. Furthermore, recent studies reported that C4 plants may also engage in biological N fixation [67]. They may release more exudates to recruit N-fixers, consistent with rhizosphere dynamics. Under N-limited conditions, this strategy may increase reliance on symbiotic N fixation relative to C3 plants.

To advance our understanding, future research should compare AOA and AOB diversity in C3 vs. C4 rhizospheres, examining how exudates shape their populations. Identifying and characterizing exudate compounds that influence AOA and AOB will clarify how chemical signals promote or suppress their activity.

2.7. Plant Resource Use Strategies Influence on AOA/AOB

Plants occupy diverse ecological niches and employ various strategies for acquiring, utilizing, and conserving nutrients [68]. The strategies plants employ for resource utilization are grounded in physiological attributes quantifiable as plant functional traits [69]. These traits influence growth, reproduction, and survival, forming the ‘leaf economics spectrum’ [70,71]. Classifying plants by functional traits helps reveal their impact on soil biological processes, with evidence across spatial scales [72]. The continuum of leaf functionality corresponds to a spectrum ranging from plants with rapid growth to those with slower growth rates. Rapidly growing plants, hereafter referred to as exploitative plants, exhibit swift N uptake and turnover, as evidenced by elevated specific leaf area (SLA), leaf N concentration (LNC), and root N concentration (RNC). In contrast, conservative plants exhibit slow N uptake and turnover, with higher leaf dry matter content (LDMC), and lower SLA, RNC, and specific root length (SRL).

Ecosystem-level studies show that plant traits, like leaf N, relative growth rate (RGR) and LDMC, explain variation in soil C, N, and litter decomposition in grasslands [73,74,75]. For instance, plant species with exploitative traits add more N through root exudates [76]. Studying plant functional traits provides a useful framework for studying plant–microbe interactions and their effects on ecosystem functions [77]. The literature shows associations between these plant functional traits and plant productivity, decomposition rates, N leaching, and microbial community composition [72,78,79,80].

Functional plant ecology classifies species by their ecological roles and contributions to ecosystems [81]. Plants exhibit a continuum of resource-use strategies and can adjust allocation in response to climate, nutrients, and competition [82]. Furthermore, N absorption and sequestration strategies vary with plant developmental stage [83]. The influence of various biotic factors on plant resource utilization is considerable. At the ecosystem level, N-use variability increases with dynamic plant community composition across spatial and temporal dynamics. These dynamics may explain variation in ammonia oxidizer abundance and activity in grasslands and agricultural soils. For instance, findings from Thion et al. (2016) showed that AOA had limited response to fertilization in barley rhizosphere, while AOB growth depended on NH₄⁺ fertilization and was higher in bulk soil [38].

2.8. Plant-Nitrifier Competition for Substrate

Nutrient availability, particularly that of N, plays a pivotal role in the C cycling of terrestrial ecosystems by modulating the biological activities of plants and soil microorganisms [84,85,86]. It is widely acknowledged that natural terrestrial ecosystems frequently experience limitations in N [87,88], with a prevailing view suggesting that temperate and boreal ecosystems predominantly face N scarcity [89]. These limitations results from direct factors (e.g., clay immobilization, low moisture diffusion) and indirect ones (e.g., low litter N, temperature, and microbial C/P constraints) [90]. These N-limited conditions engender a pronounced competitive dynamic between flora and microbial entities for nutrient acquisition [91]. Consequently, this competition results in two significant outcomes: (1) the actual nutrient assimilation by individual organisms frequently falls short of their physiological requirements due to the limited availability of these resources; (2) the nutrient uptake by one group of organisms (e.g., plants) can detrimentally impact the functional capacity of other groups (e.g., soil microbial decomposers). This competition is influenced by various factors, including the initial concentrations of NH₄⁺ and NO₃⁻, with AOB showing dominance at higher NH₄⁺ concentrations, while AOA are more competitive under oligotrophic conditions [92].

Plants and mycorrhizae compete with heterotrophs for NH₄⁺ and NO₃⁻, and with nitrifiers for NH₄⁺. The role of mycorrhizae in this competition is particularly noteworthy, as these symbiotic associations can enhance plant N acquisition [90]. It is widely hypothesized that heterotrophs are C-limited and do not directly compete with plants for N [93]. This hypothesis is supported by the observation that heterotrophic bacteria and fungi typically have lower C/N ratios than detritus, suggesting that they are more often C-limited [90]. However, this generalization is contingent upon the C/N ratios of detrital inputs, which can vary widely. Heterotrophic bacteria and fungi exhibit carbon to nitrogen (C/N) ratios within the range of 4:1 to 12:1. However, about 50% of assimilated C is respired as CO₂, and compounds like lignin and polyphenols are poorly degraded [94,95]. Thus, decomposition and N dynamics depend on detrital C/N ratios.

A functional threshold in residue C/N ratios determines whether N is mineralized or immobilized. While theoretical models use a 30:1 C/N breakpoint [90], mineralization often occurs below 20–25 and immobilization above 35:1 [96,97,98]. Intermediate C/N ratios (~25–35) often lead to a balance between mineralization and immobilization, influenced by litter quality and microbial traits.

In systems where heterotrophic microbes are C-limited (typically under lower C/N inputs), net N mineralization is enhanced, increasing the pool of plant-available inorganic N. This boosts plant N uptake and may lower residue C/N ratios over time, reinforcing N availability and decomposition [97]. Conversely, high C/N residues can trigger microbial N immobilization, potentially limiting plant N availability, especially in agroecosystems with low fertilizer inputs or limited microbial diversity. This microbial-mediated nutrient cycling also affects nitrifying communities, including AOB and AOA. In agricultural systems, AOB respond to high inorganic N from mineralization, while AOA dominate in low-N or acidic soils [28,99]. Therefore, shifts in litter quality and C/N ratios not only influence net N dynamics but also potentially alter the relative abundance and activity of nitrifiers, with downstream consequences for nitrification rates, NUE, and N₂O emissions. While plant traits and rhizosphere dynamics significantly affect ammonia oxidizer communities, fertilization remains a dominant agronomic factor. The following section explores how different fertilizer regimes, including type, rate, and composition, influences the abundance, activity, and community structures of AOA and AOB.

3. Fertilization Influence on Community Structure of amoA

Fertilization is a central component of modern agricultural systems, essential for enhancing crop productivity and sustaining soil fertility. With ~50% of habitable land under agriculture, intensified fertilization is vital to meet rising food, feed, and biofuel demands [100]. Chemical fertilizers (e.g., ammonium nitrate, urea) provide rapid N, P, and K availability, promoting fast growth and high yields. In contrast, organic fertilizers, including compost, manure, and plant-based residues, release nutrients more slowly and enhance soil structure, microbial diversity, and organic matter content [27].

Fertilization strongly affects soil microbes, especially AOA and AOB, which initiate nitrification via the amoA gene. The response of AOA and AOB to fertilization is often mediated by shifts in ammonium availability, soil pH, total carbon (TC), and electrical conductivity. AOB typically proliferate in ammonium-rich, neutral to alkaline soils, especially under high synthetic N inputs [99,101], while AOA often dominate in acidic or low-N environments due to their higher substrate affinity and tolerance to environmental stressors [28,45]. Organic inputs may balance AOA/AOB ratios by supporting microbial diversity and stabilizing substrate and pH levels. Understanding fertilization effects on amoA harboring microbes is key to managing nitrification, enhancing NUE, and reducing environmental impacts.

3.1. Substrate Concentrations (NH₃⁺, NH₄⁺ and NO₃⁻)

Ammonia oxidation serves as the primary energy generating source for both AOB and AOA under aerobic conditions, supporting ATP production through electron transport and oxidative phosphorylation [102]. Niche specialization among AOA and AOB arises from differences in ammonia substrate affinity, tolerance to ammonium toxicity, and the wide variability of ammonia concentrations and sources in soils [63]. These traits minimize competition and allow AOA and AOB to coexist in varied environments.

Laboratory-cultivated AOA exhibit significantly higher substrate affinity for ammonia compared to AOB. For instance, Nitrosopumilus maritimus SCM1 has a Km(app) of ~133 nM NH₃, compared to 1000–1500 nM for N. europaea (AOB), reflecting AOA adaptation to low NH₃ [103,104]. This difference may be partially attributed to the significantly small cell size and greater surface area-to-volume ratio of AOA [104]. AOB are generally copiotrophic, thriving in nutrient rich environments/surface soils [105]. In contrast, AOA are adapted to oligotrophic soils with low nutrients and redox potential [106,107]. Research indicates that the abundance of AOB decreases more significantly than that of AOA as one moves from the surface to subsoil, resulting in higher AOA/AOB ratios in subsoil layers. The reduced NH₄⁺ and organic C levels in deeper soil layers favor AOA over AOB [108]. In the Loess Plateau soils of China, the abundance of AOB significantly diminishes with increasing soil depth, whereas AOA populations increase until reaching a depth of 60 cm [107]. Additionally, previous studies have shown that ammonium-based fertilizers and organic N components from composted manures exhibit considerable differences in their ability to percolate through soil profiles, thereby affecting nitrogen distribution within the soil [107]. Li et al. (2021) found mineral N fertilizers increased AOB abundance 2–10 fold and shifted their community structure [109]. In contrast, AOA populations remained relatively stable, showing minimal change in abundance or community structure. Similar trends occur in neutral and calcareous soils, where AOB respond more to high inorganic N, especially NH₄⁺ fertilizers [110,111]. This differential response between AOB and AOA is closely tied to their eco physiological traits: AOB are generally copiotrophic, thriving in nutrient-rich conditions and responding quickly to fertilizer-induced substrate availability. Conversely, AOA exhibit characteristics of oligotrophic organisms, capable of sustaining activity under low ammonium concentrations and fluctuating nutrient conditions, often found in soils with lower fertilizer input or greater organic matter complexity [28,106].

Substrate availability is also significantly impacted by CO₂ levels, as it stimulates plant growth and enhances C inputs into the soil. This process may indirectly lead to the depletion of soil substrates, such as NH₄⁺ and NO₃⁻, for nitrifying and denitrifying microorganisms [112].

The meta-analysis conducted by Carey et al. (2016), Ouyang et al. (2018) and Zhong et al. (2023) demonstrated that N addition significantly increased the abundance of AOB, but not that of AOA [113,114,115]. These findings suggest that AOB exhibit greater sensitivity to nitrogen inputs compared to AOA, which can be attributed to differences in their affinities for NH₃ and their ammonia oxidation kinetics [20]. Specifically, the amoA gene in AOB has a higher maximum activity (V_m) and half-saturation constant (K_m) than the amoA gene in AOA [114,116]. However, numerous studies have indicated that the abundance AOA/AOB do not consistently correlate with nitrification activity or actual ammonia concentrations in soils. Additionally, the abundance of these functional genes offers limited insight into the actual function or activity of AOA and AOB.

3.2. Soil pH

Contemporary agricultural practices are characterized by input-intensive cropping systems, predominantly reliant on substantial applications of synthetic fertilizers [88]. These systems often feature monoculture with restricted crop rotation, which frequently results in the alterations of soil physicochemical characteristics [117]. The acidification of agricultural soils is a continuous phenomenon primarily attributed to fertilization practices, especially due to the inefficient use of ammonium-based N fertilizers [118,119].

Soil pH exerts a substantial impact on microbial communities and the biogeochemical cycles they facilitate. Previous research has demonstrated that the structure of ammonia-oxidizing microbial communities is particularly sensitive to pH variations [120]. Ammonia (NH₃), rather than NH₄⁺, is considered the primary substrate for the AMO enzyme, which catalyzes the first step in ammonia oxidation. The availability of ammonia is pH-dependent (pKa = 9.25; at 25 °C), and the conversion between its ionic and cationic forms may occur near or at the cell membrane (Norton and Stark (2011) [121]). For instance, research has indicated that increasing soil pH within the range of 4.8 to 8.5 markedly enhances soil nitrification, as observed in both field studies and microcosm experiments [122,123]. Additionally, De Boer and Laanbroek (1989) observed that nitrification rates in pure cultures of Nitrosomonas europaea decreased by approximately 50% at pH 6.5, and were reduced to just 20–30% of optimal levels at pH 6.0 [124]. No detectable ammonia oxidation or cell growth occurred below pH 5.5, indicating strong inhibition of AOB under acidic conditions. In contrast, AOA have been shown to remain active at pH values below 5, suggesting a broader pH tolerance and ecological advantage in acidic soils [29]. At low pH, NH₃ is converted to NH₄⁺, which is not directly used by ammonia-oxidizing microbes. This shift reduces the availability of NH₃, the actual substrate for ammonia oxidation. For example, at pH 7.5, approximately 4% of total ammoniacal N exists as NH₃, but this drops to only 0.4% at pH 6.5 and <0.04% at pH 5.5, illustrating an exponential decline in substrate availability [28,125]. The sharp decline in NH₃ availability at low pH reduces passive diffusion into microbial cells, while the dominant form, NH₄, must be actively transported across the membrane. This active uptake increases the energy cost of ammonia oxidation, leading to a lower energy yield per mole of ammonia oxidized. As a result, the growth efficiency and activity of ammonia oxidizers are reduced under acidic conditions [29,126].

Yao et al. (2011) investigated nitrification in tea orchard soils with pH levels ranging from 3.6–6.3 [127]. They found that these acidic soils exhibited higher nitrification of 4.3 to 28.2 µg NO₃⁻-N g⁻¹ dry soil day⁻¹. These rates were significantly higher (up to 5-fold) than those measured in adjacent forest and wasteland soils, which had lower AOA abundance. Molecular and inhibition analyses indicated that AOA, rather than AOB, were primarily responsible for the observed nitrification activity under these low-pH conditions.

Ammonia-oxidizing archaea are prevalent under both alkaline and acidic conditions and are often more abundant than AOB at higher pH levels [128,129,130]. Shen et al. (2008) found no significant correlation between AOA abundance and pH in alkaline soils (pH 8.3–8.7), although archaeal amoA gene copy numbers remained stable with increasing pH [129]. In contrast, Wessén et al. [131] observed a positive correlation between pH and AOA abundance in Cambisol soils (pH 6–6.5), suggesting that local soil type and pH range can influence microbial responses. Nitrification driven by AOA in alkaline soils (pH 7.5) was demonstrated by Zhang et al. (2010), who observed that, after soil incubation with carbon dioxide, only archaeal DNA was detected, and archaeal amoA genes outnumbered bacterial amoA genes [130]. Bates et al. (2011) also found a positive correlation between AOA abundance and soil pH across 146 soils from across the globe, particularly in forests and shrublands [128]. These contrasting observations highlight that the relationship between pH and the abundance of AOA or AOB is highly site-specific. Some studies report positive correlations, others show no relationship, and a few reports negative trends. This variability underscores the limitations of single-site studies and emphasizes the need to consider soil type, vegetation, and land use history when interpreting how pH influences ammonia-oxidizing communities.

3.3. Differential Responses of AOA and AOB to Fertilization

The influence of inorganic inputs on AOA and AOB has been extensively studied, revealing varied responses based on the type of fertilizer, its application rate, and the soil conditions. Here, we synthesize findings from several studies to provide a coherent understanding of how inorganic N impacts these microbial communities. Wessén et al. (2010) reported variable ammonia oxidation rates across treatments [131]. Though inorganic N regulates ammonia oxidizers, recent work highlights the combined effects of mineral N and organic matter. The quality and degradability of the organic substrate can alter microbial energy availability, C/N ratios, and subsequently, the activity of ammonia oxidizers. For example, Wessén et al. (2010) found that N + degradable straw enhanced oxidation more than N + recalcitrant peat [131]. This shows that both presence and quality of organics affect N cycling efficiency. Muema et al. (2016) reported a pH drop from 5.3 to 4.8 when mineral N was added to manure [132]. This acidification likely alters the microbial niche, potentially suppressing AOA and total microbial communities, including bacteria and archaea. Additionally, Wu et al. (2011) showed that 22 years of N fertilization altered ammonia oxidizers in soils [133]. In their long-term field experiment, AOB amoA gene copy numbers increased by over 50-fold in soils treated with NPK and NPK combined with organic matter, compared to unfertilized control plots. In contrast, AOA abundance remained relatively stable, regardless of fertilization treatment. These results suggest AOB dominate nitrification under long-term N input, especially with inorganic fertilizers. Notably, the inclusion of organic matter alongside NPK did not suppress AOB proliferation; in fact, it maintained similarly high AOB abundance, indicating that combined mineral and organic inputs may synergistically support AOB populations.

Fan et al. (2011) further corroborated these observations by showing that long-term application of either mineral or mineral-based fertilizers reduced AOA abundance while increasing that of AOB [134]. In these studies, soil nitrate concentrations increased linearly over time, even in the absence of detectable microbial population growth. This pattern suggests that nitrification activity continued despite static or declining AOA abundance, indicating a decoupling between activity and growth. One possible explanation is that AOA maintained enzymatic activity (via ammonia monooxygenase) without significant replication, a phenomenon common under energy-limited or nutrient-stable conditions [29]. Additionally, the lack of nitrate uptake by plants may be attributed to either non-growing (post-harvest or fallow) crop conditions, or a mismatch between microbial N transformation rates and plant N demand. We synthesized few field studies and provide their conclusions in Supplementary Table S1.

On the other hand, ecologically, comammox bacteria appear to occupy distinct niches compared to AOA and AOB. They are often found in environments characterized by low ammonia concentrations and slow nutrient turnover, such as subsurface soils, rhizosphere zones, and engineered systems, like drinking water biofilters [135,136]. Their substrate affinity for ammonia is exceptionally high (Km ≈ 0.01–0.06 µM NH₃), even surpassing many known AOA strains, suggesting a strong competitive advantage under oligotrophic or diffusion-limited conditions [136]. In terms of fertilization response, emerging studies suggest that comammox organisms tend to be less responsive to short-term N additions compared to AOB, but more resilient in low-input or organically managed soils. For instance, studies in agricultural systems have shown that comammox abundance may remain stable or even increase under low or moderate fertilization, whereas AOB dominate under high ammonium supply [137,138]. Furthermore, unlike AOB, comammox show a slower growth rate and may be inhibited under high NH₄⁺ concentrations, reinforcing their niche specialization in stable, low-N environments.

3.4. Microbial Adaptation and Ammonia Tolerance

Microcosm studies reveal complexities in ammonia tolerance and soil dynamics. Within both AOB and AOA, tolerance to ammonia varies. For example, Verhamme et al. (2011) [101] demonstrated that AOA can grow in a pH 7.5 soil amended with varying ammonium concentrations (0, 1.1, or 11 mM), indicating high substrate affinity and adaptability. In contrast, AOB growth was only observed at the highest ammonium concentration (11 mM), reflecting their requirement for elevated substrate levels. While 11 mM NH₄ exceeds typical field concentrations, it may occur temporarily following heavy fertilization. These results support the view that AOB dominate in ammonium rich conditions, whereas AOA are better adapted to lower N availability. Furthermore, the response to ammonia sources is influenced by physicochemical soil heterogeneity and microbial interactions. In a study of an acid peat soil, AOA-driven ammonia oxidation occurred with organic N mineralization rather than direct ammonium addition, which had no effect [139]. However, the addition of mineralizable N forms, such as urea, glutamate, and yeast extract stimulated AOA activity [140]. This highlights the essential role of nitrogen mineralizers and suggests the need for a deeper understanding of ammonia’s origins and movement within soils [141]. Although cultivated AOA and AOB display distinct physiological traits—particularly in their affinity for ammonia—these differences do not consistently translate into clear niche differentiation in soil environments. This is likely due to the complex interplay of other factors, such as pH, micro-scale heterogeneity, substrate diffusion, and microbial competition, which can override simple kinetic predictions. Therefore, while affinity data from pure cultures provide useful insights, they do not fully explain the spatial and functional distribution of AOA and AOB in situ [101,141]. These complexities underscore the limitations of relying solely on culture-based kinetic data to infer microbial function in heterogenous soils. Future investigations should integrate metagenomic, meta-transcriptomic, and stable isotope tracing approaches to link taxonomic composition with actual functional activity. Such high-resolution tools can clarify the ecological roles and interactions with AOA, AOB, and commamox organisms under different fertilization regimes and soil environments [28,142]. This mechanistic understanding is critical for designing targeted interventions to enhance N use efficiency and minimize losses.

Given the variable responses of AOA and AOB to fertilization, agronomic strategies must be optimized to enhance N use efficiency while minimizing environmental losses. The next section discusses integrated management approaches, such as timing of fertilizer application, use of nitrification inhibitors, and biological suppression of nitrification, to improve N retention in agroecosystem.

4. Agronomic Strategies for Improving Nitrogen Fertilization in Agroecosystems

4.1. Agronomic–Ecological Trade Offs in N Fertilization: Yield Gains vs. Environmental Costs

Nitrogen fertilization remains a cornerstone of modern agriculture, significantly enhancing crop yields and improving NUE when applied judiciously. However, the agronomic benefits of N fertilization are often accompanied by substantial environmental externalities. Globally, it is estimated that less than 50% of applied N is taken up by crops, with the reminder lost through leaching, volatilization, and gaseous emissions [143,144]. These losses not only reduce input efficiency but also contribute to adverse ecological consequences. Nitrous oxide, a potent greenhouse gas with a global warming potential nearly 300 times greater than CO₂, is largely produced during microbial nitrification and denitrification processes stimulated by excessive N inputs [145,146]. Moreover, NO₃⁻ leaching poses a significant risk to groundwater quality, particularly under conditions of high rainfall or irrigation following fertilizer application [147]. Soil acidification is another chronic outcome of ammonium-based fertilization, resulting from proton release during nitrification and leading to the depletion of base cations and reduced microbial diversity [118,148]. While strategies such as split N applications, nitrification inhibitors, and biological nitrification inhibition can mitigate some of these losses, they must be tailored to site-specific conditions and microbial community responses. Balancing yield optimization with ecological stewardship requires integrative approaches that account for both agronomic performance and environmental thresholds. Future research and policy frameworks should emphasize context-specific fertilization regimes that minimize N losses while sustaining crop productivity under diverse agroecosystems.

4.2. Emphasize on Measuring Gross N Transformation Rates

Quantifying gross N transformation rates, including mineralization, immobilization, nitrification, and denitrification is essential for understanding the fate of applied N and improving NUE in agroecosystems. Unlike net measurements, which reflect the balance of N pools, gross transformation rates provide insights into how rapidly N is cycling, and which pathways dominate under specific management or environmental conditions [149]. This knowledge is critical for guiding practices such as timing and placement of fertilizer, selection of cover crops, or the use of nitrification inhibitors.

From a farmer’s perspective, managing the N cycle more effectively translates directly into higher profitability and environmental sustainability. Inefficient N use not only wastes input costs but also results in leaching, volatilization, or greenhouse gas losses. While synthetic N fertilizers remain the most direct tool for increasing yield, their overuse often leads to diminishing returns, especially when losses are high. Integrating strategies based on a clear understanding of N transformations—such as matching fertilizer types to crop demand and soil microbial processes—can enhance NUE without necessarily increasing N input. In this way, improved knowledge of N cycling contributes to maintaining yields, reducing environmental harm, and enhancing input efficiency, which aligns with long-term agronomic and economic goals.

Traditional studies commonly assess net N turnover rates, which reflect the net outcome of simultaneous production and consumption processes such as the balance between mineralization and immobilization. While useful, these measurements can obscure the underlying dynamics of nitrogen cycling, especially when gross production and consumption occur simultaneously at high rates, resulting in deceptively small net changes. In contrast, gross N transformation rates, which independently quantify individual processes, like ammonification, nitrification, immobilization, and denitrification, offer a mechanistic understanding of how nitrogen moves through the soil system [142,150]. This level of detail is essential for addressing key knowledge gaps, such as the following:

• What proportion of added fertilizer N is quickly immobilized or lost via denitrification?
• How do microbial communities respond to different fertilizer sources or management practices?
• Which pathways dominate under different environmental or cropping conditions?

Current fertilizer recommendations often assume a uniform transformation pattern, but gross N measurements reveal that soil type, microbial activity, and organic matter inputs can drastically shift how N is processed and retained. For example, two soils may show similar net nitrification, but one may have rapid production and equally rapid immobilization, implying vulnerability to loss if microbial balance shifts.

Gross N transformations are typically quantified using ¹⁵N isotope pool dilution techniques, which allow short-term, high-resolution tracking of individual N pools and fluxes [151]. While technically demanding and historically limited to research settings, advances in tracer technology and modeling approaches are making these methods increasingly accessible for applied studies. Wider adoption of these tools could help fine-tune N management in real-world systems, reducing losses, improving fertilizer use efficiency, and guiding more site-specific, biologically informed fertilizer strategies.

4.3. Usage of Nitrification Inhibitors

Nitrification inhibitors (NIs) are widely used agronomic tools aimed at enhancing NUE and reducing N losses through leaching and gaseous emissions. These compounds work by slowing the microbial oxidation of NH₄⁺ to nitrite NO₃⁻, a process that not only reduces nitrate leaching but also curtails N₂O emissions (a potent greenhouse gas). Nitrification inhibitors primarily target the AMO enzyme, catalyzing the first step in nitrification. Compounds such as dicyandiamide (DCD), nitrapyrin, and 3,4-dimethyl-pyrazole phosphate (DMPP) are known to chelate copper at the AMO active site, effectively inhibiting its function [152,153]. Some, like nitrapyrin, also act as alternative substrates, forming inhibitory byproducts during oxidation [51].

The effectiveness of NIs has been well documented. For example, DCD and nitrapyrin have been shown to reduce nitrate leaching by up to 58.5% and N2O emissions by 83.8%, and to increase crop yields by approximately 7.5% in certain agroecosystems [153,154]. However, their success is not uniform across all soil types, and performance can be strongly influenced by soil pH, temperature, texture, and organic matter content. For instance, NIs tend to be more effective in clay-rich soils, where NH₄⁺ ions are retained through adsorption to clay particles or soil organic matter, thereby reducing N losses [155]. Soil pH also plays critical role in modulating the mobility and degradation rate of NIs [156]. A meta-analysis indicated that acidic soils exhibit more pronounced positive response to NI application compared to neutral or alkaline soils, attributable to higher microbial activity, which accelerates the degradation rate of NIs [157]. Moreover, high levels of SOM may enhance microbial activity by serving as an energy source, which can accelerate the biodegradation of NIs and diminish their inhibitory effect on nitrification [158]. Additionally, under certain conditions, such as high pH or arid soils, nitrification suppression may lead to elevated ammonium levels, thereby increasing the risk of ammonia volatilization [152].

Importantly, the microbial targets of NIs, AOA and AOB respond differently to these inhibitors. Most commercial NIs have been developed and tested primarily on AOB, particularly Nitrosomonas europaea, a model copiotrophic bacterium. AOB are generally more sensitive to inhibitors such as nitrapyrin and DCD. In contrast, AOA often exhibit higher resistance or variable sensitivity due to differences in enzyme structure and environmental niche [159]. This means that, in acidic or low-ammonium soils where AOA dominate, NIs may be less effective unless tailored to the archaeal community. Without understanding the dominant ammonia oxidizers and their ecological traits, NI application may be inconsistent or ineffective.

Thus, while nitrification inhibitors (NIs) offer clear benefits for improving nitrogen use efficiency and reducing environmental losses, they are not a one-size-fits-all solution. Their effectiveness depends on a range of biophysical and management factors, reinforcing the importance of having detailed knowledge of gross nitrogen transformations, as discussed earlier. Understanding which nitrogen pathways—nitrification, immobilization, or denitrification—dominate within a given soil–crop system is crucial for determining whether an NI is appropriate, which specific inhibitor should be used, and the optimal timing and method of application to ensure maximum efficiency.

Despite their proven potential, NIs have not been widely adopted in all cropping systems, primarily due to several practical and economic barriers. These include high costs and limited availability, variable performance under different soil and climate conditions, regulatory constraints in certain regions, and a general lack of awareness or technical knowledge among farmers regarding their proper use. Therefore, while NIs are valuable tools in nitrogen management, their use should be strategically integrated into broader nutrient management plans. This includes coupling them with site-specific understanding of soil microbial activity, as well as complementary agronomic practices, such as adjusting fertilizer timing, employing cover crops, and managing soil pH. Taken together, these strategies can help optimize nitrogen efficiency while minimizing environmental losses, leading to more resilient and sustainable agricultural systems.

4.4. Biological Nitrification Inhibition (BNI)

Biological nitrification inhibition (BNI) is a promising natural strategy for reducing soil nitrification and associated nitrogen losses in agroecosystems. This process involves the release of specific compounds from plant roots (via root exudation) that suppress the activity of nitrifying microorganisms, particularly AOB and AOA [160,161]. BNI offers an eco-friendly alternative to synthetic nitrification inhibitors (SNIs), potentially enhancing NUE and reducing environmental impacts, such as N₂O emissions. BNIs are naturally occurring compounds produced by certain plant species, including tropical grasses, like Brachiaria humidicola, and crops such as Sorghum bicolor [162]. These compounds, including terpenoids, phenolic acids, and fatty alcohols, inhibit key enzymes in the nitrification process, such as AMO, and hydroxylamine oxidoreductase (HAO) [160,163]. For instance, methyl 3-(4-hydroxyphenyl) propionate (MHPP) is a BNI that not only inhibits nitrification but also enhances root growth and N uptake [163,164]. By naturally suppressing nitrification, biological nitrification inhibitors (BNIs) can reduce the environmental footprint of agriculture compared to systems without inhibitors or those relying solely on synthetic fertilizers. Evidence shows that BNIs, such as those produced by Brachiaria species, can reduce N₂O emissions by up to 90%, lower nitrate leaching, and increase nitrogen use efficiency and crop yields [165]. In contrast, some synthetic nitrification inhibitors may reduce N₂O but increase NH₃ volatilization by up to 87%, highlighting the potential environmental advantage of BNIs [164]. They help reduce greenhouse gas emissions and prevent nitrate pollution in water bodies [161,166]. Despite their potential, the widespread adoption of BNIs faces several challenges. For instance, efficacy of BHIs can vary with different soli microbial communities and environmental conditions. Despite promising results, key knowledge gaps remain regarding the specificity and agronomic viability of biological nitrification inhibitors (BNIs). For instance, it is still unclear which nitrifying organisms AOB, AOA, or comammox, are most affected by specific BNI compounds under varying soil and climate conditions [164,167]. Additionally, the molecular pathways responsible for BNI production in plants are not yet fully characterized, limiting targeted breeding efforts. More research is also needed to assess how BNIs interact with different cash crops, especially in systems where root exudate profiles may influence or interfere with BNI effectiveness [160].

From an agronomic perspective, BNIs are still in the early stages of practical application. While some crops, like Brachiaria and sorghum, naturally produce BNIs, introducing these traits into major cereals, like wheat and rice, remains a breeding challenge. However, if successful, such efforts could offer farmers a low-input, sustainable strategy to reduce nitrogen losses.

The evolutionary advantage of BNI production appears to be adaptive: by slowing nitrification, plants retain more NH_4, their preferred nitrogen form, in many systems and minimize NO₃ leaching and denitrification losses, especially under low N input conditions. This suggests that BNI production is a selectable trait that co-evolved with nutrient conservation strategies in nutrient-poor environments.

Despite the growing recognition of BNI as a sustainable tool to suppress nitrification and enhance NUE, several practical challenges impede its widespread adoption in conventional agricultural systems. First, crop compatibility remains a significant constraint, as only a few species, such as Brachiaria humidicola and Sorghum bicolor, naturally produce effective BNIs. Integrating BNI traits into major cereals (e.g., wheat, rice, maize) via breeding programs is ongoing but not yet commercially available at scale [160,166]. Second, economic viability is often unclear, particularly when immediate yield benefits are not apparent, making it difficult for farmers to justify changes in cultivar selection or management practices. Additionally, knowledge gaps and limited farmer awareness of BNI mechanisms hinder its adoption, especially in regions where agronomic extension services are under-resourced. To overcome these barriers, multi-pronged strategies are needed, including (i) public breeding programs that prioritize BNI trait integration into high-yielding genotypes, (ii) economic incentives or subsidies to support early adoption in pilot regions, and (iii) a policy framework that recognizes the ecosystem services provided by BNI-capable crops (e.g., reduced N₂O emissions or nitrate leaching) within sustainability certification or C credit schemes. Moreover, on-farm participatory research and extension education should be expanded to demonstrate the co-benefits of BNI adoption under real-world conditions. Addressing these socioeconomic and infrastructural barriers is essential to translate the ecological promise of BNI into tangible, field-scale outcomes.

4.5. Optimizing Irrigation and Fertilizer Inputs for Enhanced Agroecosystem Productivity

Efficient irrigation and fertilizer management are crucial to improving NUE and sustaining crop yields in agroecosystems [168]. This balance is particularly vital in the context of increasing food demand and environmental concerns [144]. The upper soil layer is predominantly dominated by NO₃⁻ except for brief periods following the application of synthetic N fertilizers [169]. Short term applications of synthetic N fertilizers can significantly increase NH₄⁺ concentrations and soil pH, which in turn promote elevated NH₃ volatilization. Nitrate distribution in the soil profile is strongly affected by irrigation practices and precipitation patterns, which in turn influence NUE and loss. Excessive irrigation or heavy rainfall after fertilization can lead to nitrate leaching losses of up to 40–60% of applied N, especially in light-textured soils [147]. For example, nitrate concentrations below 60 cm depth increased by 70% following a 60 mm rainfall event [147]. These dynamics highlight the need for coordinated fertilizer and water management to retain nitrate in the root zone and reduce environmental losses.

Several factors contribute to elevated NO₃⁻ accumulation in the soil profile, including the application of synthetic N fertilizers in quantities exceeding crop requirements, which results in considerable ecological harm [170]. Improving plant N uptake and NUE from various N fertilizers can be achieved by aligning fertilizer applications with the specific requirements of crop growth stages, rather than applying them entirely before planting [171]. To optimize efficiency, fertilizers should be applied in split doses and timed according to the crop’s developmental stages, informed by thorough soil testing at the cultivation site. Side-dressing fertilizers is an effective management strategy, as it enhances N uptake and NUE while mitigating environmental impacts, including a 60% reduction in N₂O emissions [172].

5. Conclusions

The current review paper underscores the complex interplay between plant-driven processes, fertilization regimes, and ammonia oxidizer dynamics in agroecosystems. The review indicates that plant community composition, root exudation, and photosynthetic type are major determinants of the abundance and activity of ammonia oxidizing archaea and bacteria, thus reflecting their separate ecological niches and metabolic strategies. Moreover, fertilization regimes and substrate supply were found to affect these microbial groups differentially, thereby impacting nitrogen cycling and ecosystem sustainability. Optimization of nitrogen use in agricultural systems has to be performed within an integrated approach: efficient fertilization methods, such as split application and side-dressing, have to be combined with the use of nitrification inhibitors and biological nitrification suppression. Such measures not only enhance nitrogen use efficiency but also reduce the environmental risk of nitrate leaching and nitrous oxide emissions. In addition, this review paper also shows that plant functional traits and resource use strategies are key drivers of the structure of the soil microbial community and nitrogen cycling. Therefore, plant–microbe interactions should be considered in the sustainable management of agriculture.

Future studies should therefore be carried out to segregate the mechanistic pathways through which plants affect AOA and AOB, together with their functions, under changing environmental conditions. Application of certain techniques, such as 15N tracing and metagenomic analysis, can provide more detailed information on gross nitrogen transformation rates and microbial community dynamics. These studies, therefore, bridge the knowledge gaps in the interactions between plants, soil, and microbe for better inference of agronomic interventions toward improved productivity of agroecosystems, together with environmental stewardship.