1. Introduction

The ever-increasing nitrogen pollution in the environment is getting attention in recent years, particularly regarding the high warming-potential nitrous oxide (N₂O) and the discharge of the concerned ammonium (NH₄⁺-N). First, the total emissions of N₂O reached 336.33 ppb (parts per billion) [1], accounting for 6.2% of GHGs (greenhouse gases), and its heat-capturing capacity is approximately 298 times higher than that of carbon dioxide (CO₂), calculated over a 100-year period. The escalating rate of GHG emissions will accelerate global warming, leading to a projected 1.5°C temperature increase before 2030. N₂O is also a primary contributor to ozone depletion, along with chlorofluorocarbons (CFCs), which amplifies the impact on the extreme climate [2]. Furthermore, nitrogen oxides (NOx, including NO and NO₂) readily dissolve in water vapor, leading to the formation of acid rain, which releases heavy metals from soil, indirectly poisoning various organisms and causing ocean acidification. Recent findings indicate that N₂O accumulation results from the quantitative release of marine and terrestrial environments, influenced by the three major human activities of agriculture, chemical factories and wastewater treatment plants (WWTP) [3].

Ammonium (NH₄⁺-N) is another concerned compound, primarily originating from sources, such as animal husbandry, industrial and domestic sewage. Its impact is multifaceted, affecting gas, liquid and solid phases. The Environmental Protection Agency of Taiwan (Taiwan EPA) has set the regulatory limits of NH₄⁺-N and total nitrogen (TN) in the discharge, with the limits being lower than 30 and 35 milligrams nitrogen per liter (mgN/L), respectively. These regulations are set to be enforced in 2024. However, many industries and sewage treatment plants face challenges in treating the wastewater containing nitrogenous compounds to meet the EPA regulation. Moreover, NH₄⁺-N in the systematic environment can convert to pungent ammonia (NH₃) (pKa of NH₄⁺/NH₃ = 9.25) under high pH conditions. NH₃ is a controlled component under the Convention on Long-Range Transboundary Air Pollution (CLRTAP), as outlined in the Gothenburg Protocol [3].

Various methods have been employed to eliminate nitrogen pollutants, such as NH₄⁺-N and N₂O, including ion exchange resins, physical/biological adsorption and biological filtration [4], as well as thermal catalytic cracking and photocatalytic decomposition [5]. However, these treatment technologies often require significant initial investments and ongoing costs for consumable replacements. In recent years, biological treatment technologies with lower costs have been widely applied for the transformation of NH₄⁺-N in diverse environmental settings, encompassing processes such as nitrification, denitrification and other nitrogen-removal procedures [6]. Among them, the reduction of N₂O in the last stage of the denitrification process has become the prevailing method for N₂O elimination in the aquatic system. Achieving a N₂O conversion rate of over 70% is possible when there is an ample carbon source available in the system. However, in low-carbon environments, the release of N₂O becomes unfavorable for denitrifiers relying on high-carbon-to-nitrogen (C/N)-ratio food sources. Hence, autotrophic nitrogen oxidizers, including ammonia-oxidizing microorganism (AOM) and complete nitrifying bacteria (complete ammonia oxidizer, comammox), are potentially valuable contributors to reducing the residual N₂O level in the atmosphere. These autotrophic nitrogen oxidizers offer the advantages of low cost and energy consumption.

Numerous chemical reactions based on the nitrogen cycle [3] have been identified for NH₄⁺-N removal, encompassing a total of 14 reactions. However, most of these reactions have primarily been investigated in laboratory-scale systems. In Taiwan, the aerobic activated sludge tank has emerged as a popular NH₄⁺-N removal system in the water treatment plants. Nevertheless, meeting the regulations set by the Environmental Protection Agency to reduce the total nitrogen requirement to 35 mgN/L by the year 2024 poses a significant challenge. Furthermore, the complete removal of NH₄⁺-N and nitrate (NO₃⁻-N) in urban sewage with low NH₄⁺-N concentration (<50 mgN/L) and wastewater with a low C/N ratio is problematic due to slow-growth rate of microorganisms and insufficient carbon sources. Consequently, treatment processes based on the mechanisms of autotrophic microbes have been proposed as valuable tools for the elimination of nitrogenous compounds in the wastewater.

In this chapter, we will explore the chemical substances and functional microorganisms that play pivotal roles in the nitrogen cycle. Additionally, we will investigate the utilization of two sponge-based biological systems to enhance the growth rate of the slow-growth functional microbes. Specifically, we will focus on the application of autotrophic nitrogen oxidizers for mitigating residual nitrogen pollutants in the environment.

2. Impact of nitrogen pollution on the environment

Water pollution in Taiwan is primarily attributed to the discharge of pollutants from factories, livestock excretion and domestic sewage, accounting for 34.08, 2.75 and 63.17% of the total annual discharge of 3.2 billion cubic meters, respectively. Nitrogenous compounds originate from petrochemical industry (45%), high-tech industry (23%), pig manure, urine wastewater (16%) and domestic sewage (14%). A continuous release of nitrogen pollutants into the atmosphere and water bodies can lead to eutrophication and hypoxia in aquatic ecosystems (resulting from liquidous NH₄⁺-N), detrimentally affecting native plant species (associated with liquidous NO₃⁻-N), causing acid rain (resulting from gaseous NO₂), contributing to global warming (related to gaseous N₂O) and posing various other environmental challenges. The concentration of NH₄⁺-N in the petrochemical industrial wastewater ranges from 10 to 300 mgN/L (with a C/N ratio of approximately 3) [6], while domestic sewage typically contains 18.8 ± 5.71 mgN/L of NH₄⁺-N and 20.4 ± 6.38 mgN/L of total nitrogen (TN). The discharge of such sewage into natural water bodies accounts for 41.6% of national river pollution, with 7.2% classified as severe pollution (NH₄⁺-N > 3.0 mgN/L) (National Environmental Water Quality Monitoring Annual Report in 2022). These findings have further implications for groundwater systems, where 42.0% of regions exceed regulatory limits. Particularly, the Taipei Basin (<0.01 ~ 8.89 mgN/L) and the Jianan Plain (<0.01 ~ 8.76 mgN/L) exhibit the most severe contamination levels (statistical data obtained from the National Environmental Water Quality Monitoring System in 2022).

On the other hand, N₂O, a well-known greenhouse gas, has garnered significant attention due to in contribution to global warming, reaching 336.33 ppb in December 2022 [1]. Approximately 4 teragrams (Tg) per year of nitrogen is released into the atmosphere from oceanic sources, while terrestrial source contributes around 12 Tg per year of nitrogen [7]. Human activities account for 40% of total greenhouse gas emission, with specific sectors making varying contributions [8]. Agricultural soil management is responsible for 74% of emissions, wastewater treatment for 6%, stationary combustion for 5%, chemical production and other product uses for 5%, manure management for 5%, transportation for 4% and other activities for 1% [9]. In terms of industrial processes, the largest amount of N₂O is produced from nitric acid (HNO₃), with an annual emission of about 400 metric tons [10]. Biological nitrogen-removal systems, including the activated sludge system (0.06%), nitrification (2.7–9%) [11], partial nitrification (nitritation), anaerobic ammonia oxidation (anammox), nitritation-anammox procedure (1.3–2.2%) [12], nitrifier denitrification, denitrification (0.6–1.9%) [13] and nitrification-denitrification process (1.9–8.5%) [14], have been identified as potential sources of N₂O emission [15]. Nearly 70% of these emissions are attributed to the NH₄⁺-N oxidation process, resulting in a nitrogen conversion of 27% with equivalent to 600 parts per million by volume (ppmv) [16]. In the family of microbes involved in the nitrogen cycle, aerobic ammonia-oxidizing bacteria (aerAOB) have been found to release higher amounts of N₂O compared to aerobic ammonia-oxidizing archaea (aerAOA) and comammox bacteria [17].

3. Elimination of nitrogen pollutants from the surroundings

Nitrogen compounds, characterized by low molecular weight and high reactivity, have the ability to rapidly disperse into gas phase (atmosphere), liquid phase (various water bodies) and solid phase (soil or sediment), posing biological hazards and contributing to global warming. Many countries or organization, including the United States, Japan and the European Union, have implemented regulations to control the nitrogen concentrations in wastewater discharge, with limits set at less than 60 mgN/L. In Taiwan, the regulations will further restrict the total nitrogen content in public sewer systems to below 35 mgN/L by 2024. Of particular concern is N₂O, which possesses a high greenhouse potential (GHP) and is approximately 298-fold more potent than CO₂ in terms of its heat-trapping capacity. The significant impact of N₂O on global temperature rise cannot be ignored. The 2015 Paris Agreement, signed by 200 countries, aims to mitigate the rate of global warming and limit the temperature increase to within 2°C by the end of the twenty-first century. This collective effort reflects the global commitment to combat the effects of N₂O and other greenhouse gases on climate change.

The commonly physical and chemical methods are employed for the removal of NH₄⁺-N from wastewater. These methods include air stripping, ion exchange, reverse osmosis, electrodialysis and breakpoint chlorination, among others. They enable the efficient conversion or recovery of different forms of ammonium. However, these techniques are often associated with high operational costs and the challenge of disposing of secondary compounds, which limits their economic viability. In the case of N₂O reduction, thermocatalytic methods [18] and photocatalytic methods [5] have been utilized for N₂O decomposition. The use of thermocatalysis dates back to the 1950s [19], and it involves the utilization of various media such as metals, reducing oxides and zeolites [18]. Photocatalytic methods commonly employ zerovalent zinc [20]. Despite their effectiveness, physicochemical techniques are rarely used for N₂O reduction in wastewater treatment plants. This is primarily attributed to the high levels of dissolved oxygen and the low concentration of N₂O typically found in the water field [3].

Considering the aforementioned challenges, the environment-friendly and cost-effective biological treatment technologies present a promising approach for addressing the residual amounts of NH₄⁺-N or N₂O in the environment. In the case of wastewater treatment, the selection of the appropriate biological treatment system depends on the prevailing C/N ratio. A nitrification-denitrification system is suitable for high C/N ratios, whereas a nitritation-anammox process is more effective for low C/N ratios [21]. In terms of N₂O reduction, two main reaction pathways are commonly considered. The first pathway describes denitrification, where N₂O is reduced to N₂ as the final stage of the process [22]. The second pathway involves NO₃⁻-N ammonification [23]. Our research team has also been exploring an alternative elimination pathway involving oxidation [24]; however, the precise mechanism underlying NH₂OH generation in this process is still not fully understood. Further investigations are needed to unravel this key aspect.

4. Pathway of nitrogen transformation

Nitrogen is a vital element in the biosphere, playing a crucial role in atmospheric composition and the metabolic processes of living organisms. The nitrogen cycle encompasses various catabolic and anabolic reactions that drive the transformation of nitrogen compounds [25]. Five catabolic reactions include nitritification, nitratification, denitrification, dissimilatory nitrate reduction and anaerobic ammonium oxidation. These processes involve the conversion of nitrogenous compounds to different forms, facilitating their cycling within ecosystems. On the other hand, three anabolic reactions encompass ammonium uptake, assimilatory nitrate reduction and nitrogen fixation, which are responsible for the incorporation of nitrogen into organic molecules and the production of nitrogenous compounds essential for life processes. Additionally, ammonification is a crucial step in the nitrogen cycle, where organic nitrogen is converted back to ammonium. Understanding the intricate pathways of nitrogen transformation is essential for comprehending the dynamics of nitrogen cycling in various environmental systems.

The fundamental mechanisms of nitrogenous oxidation and reduction are nitrification and denitrification. Nitrification can be categorized into autotrophic nitrification and heterotrophic nitrification by different microbial communities. The former is carried out by aerobic autotrophic ammonia and nitrite oxidizers, and the latter is catalyzed by fungi as well as heterotrophic bacteria [26]. Autotrophic nitrification is a chemolithotrophic oxidation of ammonia to nitrate under strict aerobic conditions and conducted in two sequential oxidative stages: ammonia oxidation and nitrite oxidation. The yield of cells per unit of ammonia oxidizer (AOB) as the genus Nitrosomonas is approximately 0.15 mg cells/mg NH₄⁺-N_oxidized, while for nitrite-oxidizing bacteria (NOB) such as the genus Nitrobacter, it is around 0.02 mg cells/mg NO₂⁻-N_oxidized. Oxygen consumption during these reactions is estimated to be 3.16 mg O₂/mg NH₄⁺-N_oxidized and 1.11 mg O₂/mg NO₂⁻-N_oxidized, respectively. Additionally, alkalinity in the form of 7.07 mg CaCO₃/mg NH₄⁺-N is required for ammonium oxidation to maintain pH stability in the system.

Denitrification is the reduction of the oxidized nitrogen compounds (NO₂⁻-N and NO₃⁻-N) to dinitrogen (N₂) through the intermediate production of nitrogen oxide (NO) and N₂O. This transformation occurs via three distinct pathways, including respiratory denitrification, aerobic denitrification and lithoautotrophic denitrification. In the respiratory denitrification, heterotrophic microorganisms use nitrite and/or nitrate as electron acceptors, while organic matter serves as the carbon and energy source in the absence of oxygen [27]. In environmental biotechnology applications, a variety of electron donors and carbon sources, such as methanol, acetate, glucose, ethanol and others, can be used to facilitate respiratory denitrification. Next, aerobic denitrification is complete denitrification occurring at high dissolved oxygen (DO) concentration, and heterotrophic organisms are responsible for corespiration of nitrate and oxygen and they are widespread in the environment [28]. Third, autotrophic denitrifiers catalyzed the lithoautotrophic denitrification using inorganic sulfur compounds, hydrogen or ammonia as electron donors [28]. These specialized microorganisms play a crucial role in the removal of nitrogen compounds in specific ecological niches.

Nitrous oxide (N₂O) serves as a common intermediate in various nitrogen treatment systems. Within the nitrogen cycle, N₂O is primarily produced through three metabolic mechanisms, namely (1) oxidation of hydroxylamine (NH₂OH) [30], (2) nitrifier denitrification [31] and (3) anoxic nitrite reduction. First of all, NH₂OH oxidation plays a crucial role in ammonium oxidation and is a significant reaction leading to N₂O production. This process is catalyzed by hydroxylamine dehydrogenase (HAO) enzymes [32]. Two pathways have been identified for NH₂OH oxidation: (a) NH₂OH is first oxidized to NOH, which is subsequently chemically converted to N₂O [33]; (b) NH₂OH is first oxidized to NO, followed by enzymatic reduction to N₂O mediated by cytochrome c554 (cyt c554) [34].

The second is nitrifier denitrification, and the main player as Nitrosomonas europaea and other AOBs can reduce NO₂⁻ to NO, N₂O or N₂ in the absence of oxygen [35]. Two enzymes involved in this reaction are nitrite reductase (NIR) [36] and nitric oxide reductase (NOR) [37]. NIR enzymes catalyze the reduction of NO₂⁻-N to NO and subsequently, NOR enzymes facilitate the reduction of NO to N₂O [38]. Studies have shown that N. europaea lacking NIR enzyme produced four fold higher amounts of N₂O compared to the wild type with NIR enzyme, indicating the role of NIR in supporting HAO enzymes to enhance nitrification activity [39]. Notably, strains lacking NOR enzymes did not exhibit a significant effect on N₂O production [37]. These research findings suggest that NIR enzymes can support the function of HAO enzymes to raise the nitrification activity under the sufficient electron source. Therefore, nitrifier denitrification is not the main source of N₂O emission under normal situation of microbial growth. Overall, the metabolic pathways involving NH₂OH oxidation and nitrifier denitrification contribute to the production of N₂O within the nitrogen cycle. Further investigation into these mechanisms is necessary to better understand N₂O emissions and develop effective strategies for its mitigation.

On the other hand, the release of N₂O can occur under four different environmental conditions during the anoxic nitrite reduction. The first condition is N₂O accumulation due to the inhibition of nitrous oxide reductase (N₂OR) under DO concentration attaining 0.2–0.5 mg/L [40]. The second condition arises when N₂OR becomes inactive, interrupting the reduction of N₂O to N₂ at low pH levels [23]. The third condition occurs when there is an insufficient electron source that resulted from a low biodegradable organic load [41]. Lastly, nitrite (NO₂⁻-N) as the electron acceptor is more prone to induce the N₂O accumulation catalyzed by NIR/NOR enzymes, with a conversion rate of 55% per N transformation, compared to nitrate (NO₃⁻-N) at 0.8%/N transformation [42]. It is important to note that anammox bacteria and nitrite oxidizers are unlikely contributors to N₂O production, as the pathways for their potential generation of N₂O have been elucidated. Instead, four factors, including microaerobic environment, insufficient electron source, NO₂⁻-N accumulation and acidification, likely stimulate ammonia oxidizers and denitrifiers to produce N₂O in wastewater treatment systems. Furthermore, NO₂⁻-N accumulation and acidification also promote abiotic decomposition processes that contribute to N₂O emission.

To further eliminate the presence of high-GHP-N₂O, two biological reduction mechanisms have been identified: (1) N₂O reduction during denitrification and (2) NO₃⁻-N ammoniation. During denitrification, nitrous oxide reductase (N₂OR) catalyzes the reduction of N₂O to N₂ [43]. In addition, studies have revealed the growth and activity of Rhodobacter capsulatus and Wolinella succinogens in the presence of high N₂O concentration [44, 45]. However, N₂O becomes the final product of denitrification at low C/N ratio in an influent, making it challenging to further initiate the reduction of N₂O to N₂. In terms of NO₃⁻-N ammoniation, Bacillus vireti utilize N₂O oxidized to NO_x by activating the NOS operon under anaerobic condition, while simultaneously synthesizing microbial cells.

5. Biological technologies based on the nitrogen cycle

The biological nitrogen cycle encompasses 14 currently known biochemical conversion mechanisms, which can be broadly categorized into nitrification, comammox, denitrification, anammox, nitrate assimilation, respiratory ammonification (dissimilatory nitrate reduction to ammonia (DNRA)) and nitrogen fixation. Nitrification is a well-established process employed in wastewater treatment plants, where NH₄⁺-N is sequentially oxidized to NO₃⁻-N via three steps with the intermediates of NH₂OH and NO₂⁻-N. In contrast, comammox performs the direct one-step oxidation of NH₄⁺-N to NO₃⁻-N [46]. Furthermore, NO₃⁻-N serves as the inducer for initiating denitrification, which involves a four-stage reduction of NO₃⁻-N to N₂ with the intermediates of NO₂⁻-N, NO and N₂O.

In contrast to high-carbon-demanding denitrification, autotrophically anaerobic ammonia oxidation (anammox) has attracted significant attention due to its low energy consumption and minimal sludge production. Anammox is an energetically favorable reaction that utilizes NH₄⁺-N and NO₂⁻-N/nitrogen oxide (NO) as electron donors and acceptors to yield gaseous nitrogen. Furthermore, anammox organisms utilize CO₂ as the sole carbon source for cellular material synthesis [47]. Notably, hydrazine (N₂H₄) plays a crucial role as an electron donor in the conversion of NO₂⁻-N to NH₂OH in the anammox process, distinguishing it from other nitrogen removal processes.

To achieve the comprehensive removal of nitrogen compounds, various biochemical reactions and their combinations have been applied. For instance, nitrification-denitrification process has been recommended for the wastewater containing a high C/N ratio, whereas nitritation-anammox system is suitable for the influent with a low C/N ratio. In the two-stage nitrification-denitrification process, organic matter in wastewater is initially degraded to lighten the inhibition of autotrophic nitrifiers. The resulting NH₄⁺-N is then further oxidized to NO₃⁻-N by nitrification. NO₃⁻-N can be circularly used as electron acceptor for denitrification, leading to 70–90% nitrogen removal after the long-term operation. In comparison to two-stage nitrification-denitrification, a single reactor that combines the advantages of both reactions has been developed, known as the SHARON process (the acronym for Single reactor High activity Ammonium Removal Over Nitrite) [48].

In the case of nitritation-anammox, NH₄⁺-N undergoes partial oxidation to NO₂⁻-N by supplying 75% of the required oxygen, as opposed to the complete oxidization of NO₃₋-N. Subsequently, NH₄⁺-N and NO₂⁻-N are reduced to N₂. The partial oxidation of NH₄⁺-N is also known as partial nitrification [49], as it directly provides the necessary substrates for the anammox family without extra energy consumption. Throughout this process, two groups of autotrophic microbes work together to convert NH₄⁺-N into N₂, making it well suited for the wastewater with low organic content. In comparison to the two-stage system, a single reactor is employed to facilitate the growth of both autotrophic aerobically and anaerobically ammonia oxidizers, which are responsible for the transformation of NH₄⁺-N to N₂. This process is commonly referred to as CANON (the acronym for Completely Autotrophic Nitrogen removal Over Nitrite) [50].

To simplify the understanding of the system’s functionality, we focus on the characteristics of an ammonia oxidizer such as the genus Nitrosomonas. One system that controls the activity of nitrification and denitrification through the regulation of oxygen is known as the OLAND process (Oxygen-Limited Autotrophic Nitrification and Denitrification) [51]. Another system, referred to as the NO_x process [52], operates by regulating the levels of NO_x (NO/NO₂) to facilitate nitrification and denitrification. Additionally, the archaeal family can anaerobically oxidize methane (CH₄) coupled with NO₃⁻-N reduction, known as N-damo [51]. In comparison to anammox, N-damo achieved a further reduction of 0.19 mM CH₄ while utilizing 1 mM NH₄⁺-N [53]. On the other hand, aerobic deammonification directly converts NH₄⁺-N to N₂ and NO₂⁻-N via NH₂OH, although the detailed mechanism of this process is not yet well understood [54].

Recently, significant attention has been given to the production of GHP-N₂O through four reactions involved in the nitrogen cycle, including NH₂OH oxidization [30], nitrifier denitrification [31], comammox [46] and NO₂⁻-N reduction [37]. During NH₂OH oxidation, it is believed that hydroxylamine oxidase (NH₂OH oxidase, HAO) or nitric oxide reductase (NO reductase, NOR) present in microbial cells catalyzes the oxidation or reduction pathways for N₂O formation [30]. Genus Nitrosomonas, in the process of nitrifier denitrification, utilizes NH₄⁺-N or nitrogen oxides to produce N₂O under anoxic condition [55]. The comammox reaction, facilitated by the genus Nitrospira, also leads to the release of N₂O [56]. The final pathway is NO₂⁻-N reduction, which occurs under conditions of high dissolved oxygen [56], low pH [23], insufficient organic loading [43] and NO₂⁻-N in replacement of NO₃⁻-N as electron acceptor [42]. The main mechanism of N₂O reduction is the reduction of N₂O to N₂ catalyzed by N₂OR enzyme (clade II nosZ) in denitrifiers [57]. In addition, there are still unclear mechanisms of N₂O elimination, including the co-metabolism of NO₃⁻-N ammonification [23] and N₂O nitrification [24].

6. Key microbes involved in the nitrogen cycle

The nitrogen cycle involves the participation of six prominent groups of microorganisms responsible for 14 biological reactions. These groups are aerobic ammonia-oxidizing bacteria (aerAOB), aerobic ammonia-oxidizing archaea (aerAOA), anaerobic ammonia-oxidizing bacteria (anAOB or anammox bacteria (AMX)), nitrite-oxidizing bacteria (NOB), denitrifying microbes (DENer) and nitrogen-fixing bacteria (NFB).

The first group is aerobic ammonia-oxidizing bacteria (aerAOB), and it uses ammonia monooxygenase (AMO) and hydroxylamine dehydrogenase (HAO) to catalyze the oxidation of NH₄⁺-N to NO₂⁻-N via NH₂OH. This family comprises six genera of Nitrosomonas, Nitrosolobus, Nitrosovibrio, Nitrosospira, Nitrosococcus and Candidatus Nitrosoglobus [58] within two bacterial phyla of β- and γ-proteobacteria [25]. Among them, the genus Nitrosomonas is not only an obligate autotrophic nitrifier but can also act as a denitrifier, reducing NO₂⁻-N using hydrogen (H₂) as an electron donor [31]. Furthermore, the oxidation of NH₂OH to NO is initially catalyzed by HAO enzyme, and then NO is converted to NO₂⁻-N by nitric oxide oxidoreductase (NOO) [56]. However, the oxidation of NH₂OH to N₂O occurs with NO₂⁻-N as the electron acceptor in the absence of oxygen [59].

The second group is aerobic ammonia-oxidizing archaea (aerAOA), and it catalyzes the similar mechanism of NH₄⁺-N oxidation as aerobic ammonia-oxidizing bacteria (aerAOB). However, there is a distinction in the process: the intermediate of NO in the NH₂OH oxidation is rapidly consumed, and no free NO is released to the atmosphere. This family encompasses 10 genera of Nitrosoarchaeum, Nitrosopumilus, Cenarchaeum, Nitrososphaera, Candidatus Nitrosocaldus, Candidatus Nitrosotalea [60], Candidatus Nitrosotenuis [61], Candidatus Nitrosopelagicus, Candidatus Nitrosocosmicus [62] and Candidatus Nitrosomarinus [63] within the phylum Thaumarchaeota. These archaea are prevalent in the ocean and interact with other team players in the system, contributing to one third of total N₂O emission.

The third group is anaerobic ammonia-oxidizing bacteria (anAOB or AMX), and it performs the reduction of NH₄⁺-N and NO/NO₂⁻-N to N₂. Three families of anammox microbes with distinct biogeographical distributions have been identified: freshwater Candidatus Brocadiaceae (including the genera Brocadia, Kuenenia, Anammoxoglobus and Jettenia) [64], marine Candidatus Scalinduaceae (represented by the genus Scalindua) [65] and marine Candidatus Bathyanammoxibiaceae [66] in the order Candidatus Brocadiales within the phylum Planctomycetes [67]. Notably, Anammoxoglobus propionicus exhibits the ability to reduce NO₂⁻-N by simultaneously utilizing NH₄⁺-N and propionate as electron donors [68]. Furthermore, Kuenenia stuttgartiensis is capable of performing dissimilatory NO₃⁻-N reduction to NH₄⁺-N (DNRA) [69]. Additionally, this group demonstrates the capacity for carbon fixation under anaerobic condition [26], making it advantageous for GHP-CO₂ elimination applications.

The fourth group is nitrite-oxidizing bacteria (NOB) and it conducts the oxidation of NO₂⁻-N to NO₃⁻-N. This family includes seven genera: Nitrobacter in the phylum α-proteobacteria, Candidatus Nitrotoga in the phylum β-proteobacteria, Nitrococcus in the phylum γ-proteobacteria, Nitrospira in the phylum Nitrospirota, both of Nitrospina and Candidatus Nitromaritima in the phylum Nitrosponota and Nitrolancea in the phylum Thermomicrobiota [70, 71]. Notably, the widely distributed Nitrospira are further divided into canonical nitrite-oxidizing Nitrospira (canonical-Nitrospira) and comammox-Nitrospira [72]. In comparison of canonical-Nitrospira, comammox-Nitrospira catalyzes the complete oxidization of NH₄⁺-N to NO₃⁻-N. For example, Candidatus Nitrospira inopinata exhibits higher affinility of NH₄⁺-N than ammonia oxidizer under the limited NH₄⁺-N condition [73]. Moreover, Candidatus Nitrologa has been discovered in marine environments and demonstrates tolerance to high salinity [72].

The fifth group is denitrifying microbes (DENer), and they possess the ability to reduce nitrogen oxides (such as NO₃⁻-N, NO₂⁻-N, NO and N₂O) to N₂ under anaerobic, micro-aerophilic and occasionally aerobic conditions. This diverse family can be categorized into heterotrophs, autotrophs and mixotrophs based on their energy source. While heterotrophic denitrifier commonly utilizes organics as electron donor, autotrophic denitrifier (AuDen) primarily uses H₂, reduced inorganic sulfur compounds (RISCs, such as S⁰, S²⁻ and S₂O₃²⁻), sulfite (SO₃²⁻), thiocyanate (SCN⁻), iron oxides (e.g., iron disulfide (FeS₂), Fe²⁺ and Fe⁰ or zerovalent iron (ZVI)) and trivalent arsenic (As³⁺). The autotrophic families belong to various phyla including α -, β-, γ- and ε-proteobacteria [74]. It is noteworthy that certain nondenitrifying microbes possess the N₂OR enzyme that can directly utilize the residual N₂O in the environment as a source of energy and nutrients, including the genera Anaeromyxobacter, Dyadobacter, Gemmatimonas, Ignavibacterium, Melioribacter and Pedobacter. They potentially play a role in the elimination of GHP-N₂O [75]. Lastly, nitrogen-fixing bacteria (NFB) catalyze the reduction of N₂ to NH₃ through the process of nitrogen fixation. They are widely distributed in various environemts, including the phyla Proteobacteria, Chlorobi, Firmicutes and Cyanobacteria, and three methanogenic archaea of the genera Methanosarcina, Methanococcus and Methanothermobacter [76].

7. The fixation of carbon dioxide by autotrophic nitrogen families

Carbon dioxide (CO₂) is an essential element for living organisms; however, its concentration was calculated to be 421.00 ppm (parts per million) in March 2023 [1], contributing to the global warming. Microbes are the key contributors for biological CO₂ elimination by utilizing six different pathways for cell or carbohydrate syntheses. Six known pathways of CO₂ fixation are Wood-Ljungdahl pathway (W-L), 3-Hydroxypropionate 4-hydroxybutyrate cycle (3HP-4HB), Calvin-Benson-Bassham cycle (CBB), 3-Hydroxypropionate cycle (3-HP), Reductive tricarboxylic acid cycle (rTCA) and Dicarboxylate 4-hydroxybutyrate cycle (DC-4HB) [77]. In the context of autotrophic CO₂ assimilation, five distinct groups of nitrogen-related microbes are involved in the four above-mentioned pathways. First, aerobic ammonia-oxidizing archaea including mesophilic Crenarchaeota and thermophilic Thaumarchaeota prefer to use their respective modified versions of 3-Hydroxypropionate 4-hydroxybutyrate cycle (3HP-4HB) known as the Crenarchaeal HP/HB cycle and Thaumarchaeal HP/HB cycle [78]. These pathways enable them to assimilate CO₂ and carry out NH₄⁺-N oxidation simultaneously. Second, anaerobic ammonia oxidizers (anammox bacteria) employ the Wood-Ljungdahl pathway for CO₂ assimilation during their unique anaerobic ammonia oxidation process [64]. Third, the Calvin Benson Bassham cycle (CBB) is present in ammonia-oxidizing bacteria [79] as well as in the four genera of nitrite oxidizers, namely Nitrobacter, Nitrococcus, Nitrotoga and Nitrolancea [80]. These organisms utilize the CBB cycle to fix CO₂ and perform their respective oxidation processes. Fourth, nitrite-oxidizing Nitrospira and autotrophic-denitrifying Thiobacillus denitrificans have been found to involve in Reductive tricarboxylic acid (rTCA) cycle [74]. This pathway allows these organisms to fix CO₂ while carrying out NO₂⁻-N oxidation or denitrification. It is worth noting that nitrite-oxidizing Nitrococcus and Nitrospina have higher potential for CO₂ utilization in marine environments compared to ammonia-oxidizing archaea and bacteria, particularly during the early exponential phase of microbial growth. Conversely, ammonia-oxidizing Nitrosomonas demonstrate rapid rate of cell synthesis in both late exponential and stationary phases [81]. In summary, autotrophic nitrogen-functional microbes possess the remarkable ability to utilize nitrogen compounds and CO₂ as energy source and cell synthesis. This capability not only contributes to the reduction in global warming but also aids in the removal of nitrogenous pollutants from the environment.

8. Cultivation systems of nitrogen-functional microbes

The cultivation of nitrogen-functional microbes relies on providing suitable energy sources for their growth. This includes nitrogen sources, alkalinity (typically NaHCO₃), buffer (KH₂PO₄ and Na₂HPO₄), nutrients and trace elements [82]. Different nitrogen sources are utilized depending on the specific group of microbes being cultivated. For instance, NH₄⁺-N, NO₂⁻-N, NH₄⁺-N/NO₂⁻-N, NO₂⁻-N/NO₃⁻-N and N₂O were, respectively, used for ammonia oxidizers (AOA and AOB), nitrite oxidizer, anammox, denitrifier and N₂O-utilizing microbes. In addition to nitrogen sources, essential nutrients, such as calcium, magnesium and iron, are provided through CaCl₂, MgSO₄ and FeSO₄. In the case of ammonia oxidizers carrying out partial nitrification, Na₂SO₄ is added as a supplement. Trace elements, which are crucial for microbial growth, consist of CuSO₄, ZnSO₄, MnCl₂, NiCl₂, CoCl₂, NaMoO₄, NaSeO₄, NaWO₄, Na₂-EDTA and H₃BO₄. To avoid Na₂-EDTA and H₃BO₄ from serving as carbon source for the growth of heterotrophic bacteria, they are excluded from the trace elements for cultivation of N₂O-utilizing microbes. These cultivation systems aim to provide the necessary nutrients and conditions for the successful growth of nitrogen-functional microbes, enabling their study and potential application in various nitrogen cycling processes.

The cultivation system’s design plays a crucial role in the successful enlargement of nitrogen-functional microbes. An important factor to consider is the choice of a suitable habitat for their growth. In this regard, the downflow hanging sponge (DHS) system utilizes a polyurethane sponge as a supporting material, as depicted in Figure 1. This sponge provides a three-dimensional (3D) space that facilitates the growth of microorganisms. When wastewater, containing microbes and foods, flows into the sponge, microbial cells are retained both inside and outside the sponge media. The unique microenvironment allows for the coexistence of aerobic and anaerobic nitrogen-functional microbes. Specifically, the surface of the sponge supports the growth of aerobic autotrophs responsible for nitritation, while deeper within the media, anaerobic microbes catalyze the anammox process [82]. Since its initial development in 1995, the DHS system has undergone several modifications, resulting in six different configurations of sponge setups [29]. The superiorities of the DHS are high biomass retention, long sludge retention, minimal sludge production and less energy consumption, particularly benefiting to cultivate the slow-growing autotrophic nitrogen-functional microbes. The combination of the polyurethane sponge as a support material and the unique microenvironment provided by the DHS system contributes to the successful cultivation of nitrogen-functional microbes, enabling their study and application in various nitrogen cycling processes.

Figure 1.
Conceptual cross-sectional view of the cube-type downflow hanging sponge (DHS sponge) (modified from [29]).

The G1-type DHS reactor is a nonsubmerged fixed-bed reactor, illustrated in Figure 2. It consists of a closed rectangular column with a total volume of 2.5 L, while the working volume is 0.596 L, considering the 98.4% void ratio of the sponge material. Inside the reactor, 19 strips of triangular sponge (sized 2.8 × 2.8 × 4 cm) are arranged on two opposite inner walls, with a gap of 0.5 cm between each consecutive sponge. The height of the reactor column was 1 m, but the effective height was 2 m, as the sponge strips adhered on opposite walls were connected in series during the operation of the reactor. This design allows for enhanced contact between the wastewater and the sponge media, promoting efficient microbial growth and nutrient removal. Another cultivated system is the upflow submerged sponge (USS) reactor, shown in Figure 3. This reactor configuration provides an effective volume of 1.5 L within a 3-L column. The USS reactor employs a 1-L sponge volume as the attached media, creating a coexisted environment for suspended and biofilm-type microorganisms. The temperature control in the USS reactor is achieved through a cycled water system, ensuring optimal conditions for microbial activity and growth. These cultivation systems, namely the G1-type DHS reactor and the USS reactor, provide suitable environments for the growth of nitrogen-functional microbes. The well-designed arrangement of sponge media in these reactors allows for efficient nutrient utilization and microbial interactions, enabling the study and application of nitrogen cycling processes in wastewater treatment and environmental biotechnology.

Figure 2.
Schematic diagram of G1-type downflow hanging sponge (DHS) system.

Figure 3.
Schematic diagram of upflow submerged sponge (USS) system.

9. Test procedure and analytical methods used in NH₄⁺-N and N₂O oxidation

The G1-type DHS system was used for three processes that enhance NH₄⁺-N transformation. First, a stepwise increase of NH₄⁺-N concentration from 30 to 400 mg N/L was performed at eight different phases in the nitritation system, which corresponded to the nitrogen load of 0.47 to 6.42 kg N/m³-day. Second, the potential of the anammox system was tested using four different parameters of cultivated temperature, inflow rate, substrate concentrations (including NH₄⁺-N and NO₂⁻-N) and effluent recirculation. Third, the operational conditions of the complete nitrogen transformation in a single-type DHS reactor (namely CnDHS) were similar to those of the nitritation system. All systems were placed in the 30–35°C incubator. In addition, the low oxygen concentration inside both nitritation and CnDHS systems was controlled by adjusting the airflow rate from 1.5 to 16 L/day based on the partial pressure of oxygen in the reactor.

Two USS systems were used to improve the efficiency of anammox and N₂O oxidation. In the anammox reactor, the increase of nitrogen load was observed to take place from 9.60 to 38.4 mgN/L-day, along with the concentration increase of chloride from 160 to 1200 mg/L under a fixed HRT of 4.2 days. For N₂O oxidation, the rise of N₂O load was performed by increasing the substrate flow rate from 0.04 to 0.32 L/day (HRT shortened from 4 days to 0.5 days) under the fixed N₂O substrate of 25 mM in the liquid based on the 100% gaseous N₂O dissolved in the medium. The flow of air inside the reactor was controlled to adjust the desired oxygen concentration, and the exhaust gas from the reactor was collected using a gas bag. The microbial activity was further tested in the batch assays with different N₂O concentrations under the satisfactory oxygen conditions.

To monitor the performance of all systems, NH₄⁺-N, NO₂⁻-N and NO₃⁻-N in influent and effluent were regularly measured using a colorimetric method (HACH, USA) and ion chromatograph (SH-120A, SHINE, New Zealand). The composition of off-gas was determined using gas chromatography (Shimadzu GC-8APT for O₂, GC-8AIT for N₂O, CO₂ and N₂). Theoretical DO concentration in the bulk liquid flowing on the sponge surface was computed from oxygen content in the gas phase according to modified Henry’s equation, and the actual DO concentration in effluent was measured directly by a DO meter (YSI/Nanotech Inc., Japan). The concentration of the retained biomass in the sponge material was measured at the end of the operation, and the biomass was stored in a −20°C freezer for microbial clarification. DNA was first isolated using MOBIO PowerSoil DNA extraction kit, and microbial community and functional genes catalyzed nitrogen transformation were further analyzed by TOPO cloning kit and SybrGreen quantitative PCR (QuantStudio 1, ThermoFisher Scientific, USA) with the specific primer pairs.

10. Use of autotrophic N-removal processes for NH₄⁺-N reduction

This section aims to assess the efficiency of NH₄⁺-N conversion by comparing three autotrophic N-removal processes, including nitritation (also called partial nitrification), anammox and complete nitrogen transformation in a single-type reactor. The G1-type DHS is used for microbial enlargement and functional evaluation. The findings of the study are presented and discussed below.

In the operation of nitritation, the G1-type DHS reactor was subjected to a total of seven phases, with NH₄⁺-N inflow rates ranging from 0.47 to 1.60 kgN/m³-day. The hydraulic retention time (HRT) was fixed at 1.5 h, and the temperature was maintained at 30°C. The airflow rate, ranging from 2 to 16 L/day, was adjusted to control the oxygen content in the system. Microbes cultivated in the DHS exhibited high capability for NH₄⁺-N oxidation, achieving rates of up to 1.92 kgN/m³-day even at low oxygen condition. This performance surpasses those of a fixed-film bioreactor (0.58 kgN/m³-day) [83] and submerged membrane bioreactor (1.30 kgN/m³-day) [84] operating under sufficient oxygen supply. Furthermore, partial NH₄⁺-N oxidation of 58.6% was attained at NH₄⁺-N load of 3.24 kgN/m³-day. This resulted in the production of 37.5% NO₂⁻-N and 4.0% NO₃⁻-N with an oxygen concentration of 0.40% O₂ (0.16 mg/L of DO) (see Figure 4) [85]. Similarly, a biofilm system demonstrated that partial nitrification with 50% NH₄⁺-N conversion was attained at oxygen concentrations below 0.2 mg/L [86]. The growth rate of ammonium oxidizers was 2.56-fold faster than that of nitrite oxidizers under the DO concentration below 1.0 mg/L [87]. Similarly, the growth yield of Nitrosomonas sp. under oxygen stress as low as 1% was found to be 5 times higher compared to conditions with saturated DO conditions. The G1-type DHS reactor provided a high biomass concentration of 3.84 g volatile solids (VS)/L, enabling a nitrifying activity of 0.20 kg NH₄⁺-N/kg VS-day. The activity of ammonia oxidizer in the DHS was comparable to those of the suspended growth-type reactors (0.17–0.29 kg NH₄⁺-N/kg VS-day) [88], and higher than that of the biofilm-type system (0.08–0.10 kg NH₄⁺-N/kg VS-day) [89]. However, GHP-N₂O production was detected at a level of 0.5% in the DHS, equivalent to 13% of the oxidized NH₄⁺-N under a gas-phase oxygen content of 0.4% (0.16 mg/L of DO) (see Figure 5). Similarly, N₂O production was observed from 10% of the oxidized NH₄⁺-N under the O₂ concentration of 0.18 mg/L [30]. The analysis of microbial community revealed that 32.4% phylotypes closely related to Nitrosomonas sp. strain ENI-11 dominated in the DHS, while denitrifying genera of Azoarcus and Bradyrhizobium and nitrite-oxidizing Nitrobacter coexisted and participated in the nitrogen cycle [90]. The amoA gene encoded in the enzyme, which catalyzes ammonium oxidation, was used for determining the functional microbes, resulting in the phylotypes within the Nitrosomonas europaea/Nitrosococcus mobilis lineage being the key players in the nitritation in the DHS at low oxygen atmosphere. The images of fluorescence in situhybridization (FISH) demonstrated that 41% of β-proteobacterial ammonia oxidizers coexisted with 5.4% of Nitrobacter spp. within the bacterial community, accounting for 83% of the total population. Based on these findings, it can be concluded that the ammonia oxidizer as Nitrosomonas family was numerically dominant over nitrite oxidizer in the DHS reactor, facilitating the occurrence of nitritation at low oxygen supply. However, the presence of nitrite reductase, involved in N₂O production through nitrifier denitrification, was induced at low oxygen partial pressures [39].

Figure 4.
Effect of O₂ in the downflow hanging sponge (DHS) for nitritation on the ratio of NO₂⁻-N_produced/NH₄⁺-N_removed. Wherein, P1 ~ P7 is the data taken from phase 1 to phase 7 of the operation in the DHS.

Figure 5.
Effect of O₂ in the downflow hanging sponge (DHS) for nitritation on the ratio of N₂O_produced/NH₄⁺-N_removed. Wherein, P6 ~ P7 is the data taken from phase 6 to phase 7 of the operation in the DHS.

The optimal proportion of NH₄⁺-N and NO₂⁻-N for anammox reaction was achieved through first-stage nitritation. The DHS employed for the anammox process operated at a total nitrogen load ranging from 0.48 to 5.96 kgN/m³-day with NH₄⁺-N and NO₂⁻-N maintained at an equal proportion. The HRT was set between 0.7 and 2 h, and the reactor was operated at a temperature range of 30 to 35°C. The highest nitrogen-removal rate achieved in the DHS was 2.27 kg N/m³-day, which surpassed the performance of other biofilm systems with the removal rates of 0.2 to 2.0 kgN/m³-day [91, 92]. However, the nitrogen-removal rate in the DHS was lower than that reported for an upflow fixed-bed column reactor designed for highly enriched anammox [93]. The DHS exhibited a biomass concentration of 5.59 g VS/L within the sponge media, enabling anammox activity of 0.39 kgN/kg VS-day. Remarkably high removal efficiency of 95.4% was achieved at a loading rate of 1.94 kgN/m³-day and HRT of 1.0 h, giving NO₂⁻-N_utilized/NH₄⁺-N_removed of 1.25 ± 0.080 and NO₃⁻-N_production/NH₄⁺-N_removed of 0.25 ± 0.042 [94]. Notably, no N₂O was detected in the DHS, highlighting the physiological capacity of anammox bacteria to suppress N₂O production [95]. Moreover, based on theoretical calculations, approximately 76% of the removed NH₄⁺-N was converted to N₂ through the anammox reaction, while the remaining 24% was suggested to be consumed via NO₃⁻-N reduction processes, including assimilation and dissimilation, as well as denitrification [96]. The anoxic microenvironment within the sponge media of the DHS, as depicted in Figure 1, likely provided a reducing environment and limited carbon sources. Additionally, cell lysis resulting from microbial mortality during resting periods further contributed to the availability of organic matter. As discussed earlier, the co-occurrence of nitrate reduction or denitrification alongside the anammox reaction in the DHS led to higher nitrogen removal (95%) than other systems [93, 96, 97, 98]. The key microbial players in this community included anammox genera Kuenenia and Anammoxoglobus, ammonium-oxidizing genus Nitrosomonas, as well as denitrifying capability of the genera Comamonas [99] and Diaphorobacter [100]. Together, these microbial groups worked synergistically to reduce nitrogenous compounds and facilitate efficient nitrogen removal in the system.

On the contrary, when it comes to the USS system designed for treating high salinity wastewater with a low C/N ratio, a remarkable removal efficiency of 93.3% was attained at a nitrogen load of 38.4 mgN/L-day, even under a chloride (Cl⁻) concentration of 300 mg/L. However, it should be noted that the increase of Cl⁻ concentration to 1200 mg/L resulted in an extended adaptation period of 1 month for the utilization of NH₄⁺-N and NO₂⁻-N. Comparing the USS system to the DHS system used for treating fresh water, the USS system enlarged the main groups of anammox Brocadia, ammonia-oxidizing Nitrosomonas, canonical nitrite-oxidizing or comammox-functional Nitrospira. Additionally, denitrifying genera, such as Denitratisoma, Acinetobacter, Pseudomonas and Comamonas, were also observed in the bacterial community of the USS system. The activities of microbes involved in ammonia oxidation, comammox, anammox and denitrification were assessed using the functional indicators of amoA, crenamoA, Nts-amoA, hszA, nirS and nirK genes. As shown in Figure 6, the abundance of the former four genes notably increased after 693 days of the cultivation in the USS system. In contrast, the presence of denitrifying nirS and nirK genes decreased over time. These findings suggest that anammox bacteria replaced the denitrifying microbes in facilitating the reduction of nitrogen oxides such as NO₂⁻-N and NO₃⁻-N. Additionally, the USS system demonstrated the dominance of slow-growing autotrophic nitrifiers harboring the amoA gene.

Figure 6.
Functional genes of microbes are involved in the conversion of NH₄⁺-N and NO₂⁻-N in the. Upflow submerged sponge (USS) system for treating high-salinity wastewater. Wherein, amoA and crenamoA stand for ammonium oxidation, Nts-amoA for Comamonas, hszA for anaerobic ammonia oxidation (anammox) reaction, and nirS and nirK for nitrite oxidation. In addition, the value below the detected limits of 2.32x10¹ copy/kg VSS was used as 1.61x10¹ copy/kg VSS.

To optimize the synergy between nitritation and anammox, a single reactor capable of completely converting NH₄⁺-N to N₂ was implemented in the DHS. Initially, the slow-growing and environmentally sensitive anammox bacteria were cultivated within the DHS, followed by the colonization of enlarged aerobic ammonia oxidizers coated on the outer surface of the sponge media. The DHS operation involved varying NH₄⁺-N loads from 0.30 to 2.42 kg N/m³-day, while maintaining limited oxygen levels controlled by airflow of 1.5 to 5.4 L/day. Remarkably, the maximum nitrogen-removal rate reached 1.53 kgN/m³-day, surpassing the performance of suspended sludge system (0.2 kgN/m³-day) [101] and biofilm-type reactor (1.5 kg N/m³-day) [89]. Furthermore, this system demonstrated stable autotrophic nitrogen removal, even at an HRT as short as 2 h, in contrast to other processes requiring much longer HRTs (up to 10 h). Notably, an impressive efficiency of 84.8% was attained at NH₄⁺-N load of 1.51 kgN/m³-day, giving 84.5% of N₂ production alongside 8.0% NH₄⁺-N and 7.5% of nitrogen oxides. The precise adjustment of oxygen content in the DHS proved crucial in controlling the proportion of NO₃⁻-N and N₂ production. In Figure 7, it is evident that an oxygen concentration of 1.0% serves as the critical threshold for distinguishing the dominant reaction pathway between anammox and nitrification, as indicated by the ratio of NO₃⁻-N/(NO₂⁻-N + NO₃⁻-N). Anammox bacteria predominantly catalyze the production of NO₃⁻-N when the O₂ content falls below 1%, whereas complete nitrification, driven by the faster growth rate of nitrite oxidizers compared to ammonia oxidizers, occurs at O₂ concentrations above 1% in gas phase. Remarkably, a reactor operating with O₂ levels below 0.5% air saturation efficiently cultivates microbes with varying oxygen requirements [102]. However, the restricted O₂ concentration below 1% stimulates GHP-N₂O production through the activity of NO₂⁻-N and NO reductases, resulting in a loss of 7.2% nitrogen in the DHS. Considering the mass balance, it is observed that 55.7% of NH₄⁺-N is utilized for NO₂⁻-N production, 34.5% is further transformed to gaseous N₂, but 9.5% is diverted toward the formation of N₂O under an O₂ concentration below 0.48% (Figure 8). In the DHS, a similar pattern of N₂O production was observed during nitritation, with N₂O accounting for 13% of the total nitrogen gas at an O₂ content of 0.4%. Additionally, the DHS supported the coexistence of six different nitrogen-functional microbes, namely aerobic ammonia-oxidizing Nitrosomonas, anaerobic ammonia-oxidizing Brocadia, canonical nitrite-oxidizing or comammox-functional Nitrospira, denitrifying Comamonas and nitrogen-fixing Bradyrhizobium. This diverse microbial community, facilitated by the DHS’s excellent biomass retention capacity, created favorable conditions for the complete transformation of NH₄⁺-N to N₂.

Figure 7.
Effect of O₂ in the downflow hanging sponge (DHS) for N removal on the production of NO₃⁻-N/(NO₂⁻-N + NO₃⁻-N).

Figure 8.
Effect of O₂ in the downflow hanging sponge (DHS) for N removal on the ratio of N₂O-N_produced/NH₄⁺-N_removed. Wherein, P1 ~ P7 is the data taken from phase 1 to phase 7 of the operation in the DHS.

11. DHS-type systems applied for N₂O elimination

The emission of GHP-N₂O is commonly observed in various mixed systems involved in nitrogen transformation, particularly during NH₂OH oxidization [30], nitrifier denitrification [31], comammox [46] and NO₂⁻-N reduction [37]. Traditionally, the primary approach for N₂O removal is carried out through its reduction in the denitrification, which requires an adequate carbon source to activate the microbes possessing the N₂OR enzyme. However, this study explores an alternative possibility of N₂O oxidation catalyzed by inhered ammonia oxidizers, utilizing the abundant atmospheric O₂ as electron acceptor. The theoretically thermodynamic equations for these reactions are presented in Eqs. (1) and (2). These equations demonstrate the thermodynamic feasibility of N₂O oxidation to NO₃⁻-N by ammonia oxidizers using O₂ as the electron acceptor.

N2O+O2+H2O→2NO2−+2H+ ΔG0=−5.33 KJ/mole e−E1

N2O+2O2+H2O→2NO3−+2H+ ΔG0=−21.2 KJ/mole e−E2

In order to assess the potential of N₂O oxidation, a series of experiments were conducted, including batch assay and continuous systems such as DHS and USS. The batch assay exhibited an average N₂O-removal rate of 107.1 mg N/g VSS (volatile suspended soilds) over a 60-day incubation period. During this time, NO₃⁻-N and N₂-N production accounted for 6.66 and 1.86% of the total nitrogen, respectively. The microbial community in the batch culture consisted of 93% domain Bacteria and 7% domain Archaea. Notably, the population of ammonia-oxidizing Nitrososphaera gargensis-like group experienced a 3.45-fold increase, representing 68.4% of the total archaeal population. Moving on to the continuous-flow DHS system, a range of N₂O-N loads, equivalent to HRT from 4 days to 12 hours, was applied under fully saturated O₂ conditions. The DHS demonstrated an impressive N₂O removal efficiency of 95% with a rate of 6.99 kg N/kg VSS-day. Concurrently, NO₃⁻-N production reached a maximum of 22.9 μmoles/day, while NH₄⁺-N and NO₂⁻-N were not detected throughout the entire operation. These findings indicate the effective capability of the system in removing N₂O while promoting NO₃⁻-N production.

In the operation of the DHS, the concentration of NO₃⁻-N was accumulated in the effluent up to 7.79 mg N/L at the loadings below 1 mmole N₂O-N/day, whereas gaseous N₂ became the predominant end-product at the loadings over 2 mmoles N₂O-N/day [103]. The mass balance analysis, as illustrated in Figure 9, revealed that the produced N₂ was derived from two main pathways: the reduction of NO₃⁻ formed from the N₂O oxidation and the direct reduction of N₂O itself. The conversion yield of mole N₂-N per mole N₂O-N (0.109 as average of phases II-V) was approximately twice as high as the conversion yield of mole NO₃⁻-N per mole N₂O-N (0.052 in phase I). Furthermore, the anomalous increase in pH was observed in the final phase of the DHS operation, suggesting the potential accumulation of alkaline intermediates, such as NH₂OH. This observation lends support to the hypothesis that N₂O oxidation to NO₃⁻-N via NH₂OH as the intermediate may occur in the system.

Figure 9.
The type of production from the transformed N₂O in the downflow hanging sponge (DHS).

In addition, a diverse community of the nitrogen-functional microbes coexisted in the DHS, including the dominant nitrite-oxidizing bacteria of the genus Nitrospina and ammonia-oxidizing archaea of the genus Nitrosophaera [104]. Furthermore, the family Acidobacteriaceae exhibited both denitrifying and DNRA activities [105]. A smaller population of ammonia-oxidizing bacteria was also detected (3.97 × 10³ copies/μg DNA based on the amoA gene). The presence of the genera Nitrosophaera and Nitrospira, which are typically associated with marine environments, dominated in the aerobic DHS during N₂O transformation, potentially serving as regulators of marine N₂O production. Moreover, the contribution of nitrogen conversion through other redox reactions involving metals was estimated to be less than 10%. Therefore, we hypothesize that the major nitrogen loss observed in the DHS system could be attributed to the accumulation of unidentified intermediates resulting from N₂O transformation such as NH₂OH, nitrogen oxide (NO_x), dinitrogen trioxide (N₂O₃) and dinitrogen tetroxide (N₂O₄) and possibly other compounds that were not analyzed in this study.

Another system utilized for cultivating the N₂O-functional community is the submerged USS system, which received an average N₂O concentration of 10.96 ppm with N₂O load of 188 μmoles/day. Throughout the 240-day operation, the removal efficiency ranged from 13.6 to 37.5%, which resulted in effluent concentrations of NH₄⁺-N and NO₂⁻-N ranging from 0.08 mgN/L to 0.60 mgN/L and 0.12 mgN/L to 0.44 mgN/L, respectively. After a 30-day enrichment, the production of NO₃⁻-N commenced, reaching a maximum concentration of 2.74 mgN/L, equivalent to a production rate of 47.0 μmoles/day, with 70.98% originating from the oxidation of N₂O on day 50, surpassing that of the aerobic DHS.

However, N₂ production began while the transformed NO₃⁻-N concentration was below 10%, and the maximum N₂ accumulation reached 53.6 μmoles/day, representing 5.4% of N₂O conversion. In contrast to the DHS system, where N₂ served as the end-product instead of NO₃⁻-N, the USS reactor exhibited N₂ production in conjunction with the presence of NO₃⁻-N. Additionally, the average of O₂ consumption during the 240-day operation ranged from 50.2 μmoles/day to 140.2 μmoles/day. Conversely, the concentrations of chloride and sulfate present in the cultivated medium experienced a reduction of over 78.2%, indicating their potential utilization by the nitrogen-functional microbes. Additionally, the activity of the N₂O oxidation, as determined by quantitative polymerase chain reaction (qPCR) analysis targeting functional genes, revealed a significant increase in the population of both ammonia oxidizers and nitrite oxidizers, reaching levels of approximately 10⁵ cells/kg VSS. Notably, the growth rate of ammonia-oxidizing archaea outpaced that of ammonia-oxidizing bacteria, while the genus Nitrospira exhibited a higher growth rate compared to the genus Nitrobacter.

On the one hand, the four effect factors tested by the batch assay were discussed as follows. First, the influence of oxygen concentration was evaluated, and it was found that increasing the air input by 3-fold resulted in a 1.36-fold increase in NO₃⁻-N production compared to the aerobic DHS operated with 7% of air supply. This led to a NO₃⁻-N production rate of 10.16 mg/gVSS. Second, the addition of 34.5% methane (CH₄) led to the highest NO₃⁻-N production, reaching a value of 15.19 mg NO₃⁻-N/gVSS. Third, the enrichment of ammonia oxidizers using 7.5 mM NH₄⁺-N in conjunction with N₂O-degrading microbes resulted in a conversion of 9.62 mg NO₃⁻-N/gVSS. Fourth, the addition of 1.4 mM manganese (Mn²⁺) aimed to convert N₂O to NO₃⁻-N, but it achieved a lower conversion rate of 6.58 mg NO₃⁻-N/gVSS compared to the aerobic DHS. The microbial populations of both domains Bacteria and Archaea displayed an increase in the growth rate in response to O₂, CH₄ and NH₄⁺-N, except in the case of the assay with Mn²⁺ addition. Notably, the increase in the population of nitrite oxidizing genera Nitrobacter and Nitrospira was superior than that of ammonia oxidizers.

To further confirm N₂O oxidation pathways involving NH₄⁺-N oxidation and NO₂⁻-N oxidation, pure cultures of the genera Nitrosomonas and Nitrobacter were employed for N₂O elimination using sodium bicarbonate (NaHCO₃) and carbon dioxide (CO₂) as inorganic carbon sources. Ammonia-oxidizing Nitrosomonas demonstrated efficient catalysis of N₂O transformation, resulting in the production of 0.20 to 0.64 mole NO_x/mole N₂O and 0.30 to 0.54 mole NH₄⁺-N/mole N₂O. The addition of a two fold carbon source facilitated an increase in N₂O oxidation. Notably, CO₃²⁻ derived from NaHCO₃ was more readily utilized for cell synthesis compared to gaseous CO₂. In contrast, the nitrite-oxidizing Nitrobacter exhibited lower N₂O transformation rates ranging from 25.94 to 53.84%. The addition of either inorganic carbon source resulted in high NO_x production from the oxidized N₂O, ranging from 0.90 to 0.91%. Based on these findings, it is suggested that the aerobic degradation of N₂O follows a possible route involving the conversion of N₂O to NO₃⁻-N via NH₂OH and NO₂⁻-N as the intermediate. Additionally, aerobic nitrogen-functional microbes were the key contributors for N₂O elimination.

12. Summary

Both NH₄⁺-N and GHP-N₂O are of great concern due to their impact on the globally ecological environment. This chapter introduces the application of the polyurethane sponge as a useful medium, providing a three-dimensional space for microbial growth. Two types of sponge-based systems, namely DHS and USS reactors, were utilized for various reactions of autotrophic nitrogen transformation. In terms of NH₄⁺-N conversion, processes, such as nitritation, anammox and one-stage nitrogen removal, demonstrated satisfactory rates of total nitrogen removal. Regarding N₂O elimination, three potential routes were identified for N₂O transformation, involving the production of NO₃⁻-N through the conversion of NO₂⁻-N, NH₄⁺-N or direct conversion to N₂ as end-product. Five different autotrophic nitrogen-functional microbes cooperated synergistically within the expanded system, contributing to the reduction of nitrogenous compounds.

Acknowledgments

We are grateful to Professor Jer-Horng Wu who works in the Department of Environmental Engineering, National Cheng Kung University in Taiwan, for assisting with microbial clarification and data analysis of the N₂O oxidation system. We also thank 21 Century Center of Excellence (COE) program (Japan), Ministry of Science and Technology (Taiwan) and NCKU Research & Department Foundation (Taiwan) for supporting the funding.

Zamosteanu Nina

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