English

• The flue gas emitted by factories generates a large amount of wastewater containing nitrite, after undergoing denitrification process [1,2]. Nitrite poses significant risks to human health, as it can not only induce methemoglobinemia but also form potentially carcinogenic N-nitrosamines in the gastrointestinal tract [3,4]. In response to this situation, the World Health Organization (WHO) has set a limit of 3 mg/L for nitrite concentration in drinking water [5,6]. Therefore, to ensure the safety of drinking water, it is necessary to propose efficient methods to treat wastewater containing nitrite.

  Currently, various methods are employed for the treatment of nitrite, including adsorption [7], membrane separation [8], biological approaches [9,10], electrochemical reduction [11,12], and photocatalytic reduction [13,14]. However, adsorption and membrane separation methods only concentrate nitrite without achieving its mineralization, which may result in potential secondary pollution. The biological method requires a relatively long period for cultivating adapted bacteria and mild conditions for microbial communities. The electrochemical method encounters challenges in the preparation of suitable electrocatalysts and the effective modulation of reduction products. In contrast, with the development of advanced reduction processes (ARPs), photocatalytic reduction has gained increasing popularity as an emerging approach for nitrite removal [15].

  Photocatalytic reduction is an innovative and sustainable technology specifically developed for the treatment of oxidative pollutants in wastewater [16,17]. It utilizes electrons and highly reactive radicals generated under illumination to eliminate contaminants [18]. Among the various reductive radicals such as sulphur dioxide radical (SO₂^•−) and bisulphide ion (HS^•−) [19], CO₂^•− is a relatively emerging species [20,21]. It exhibits remarkable reducing capabilities by utilizing its unpaired electrons as electron donors. Hence, it possesses a strong reduction potential (E⁰(CO₂/CO₂^•−) = −1.81 V vs. SHE) and thus could effectively eliminate nitrite [22,23]. Numerous researchers have demonstrated that the reduction of NO₂⁻ by the CO₂^•− radical results in a high selectivity for N₂ in the resulting products [24]. Consequently, the primary products of the NO₂⁻ reduction in this process are N₂ and CO₂, mitigating the risk of secondary pollution compared to sulfur-containing radicals. However, this photocatalytic reduction process is limited by the CO₂^•− generated from the activation of carboxylic acid precursors [21,22]. Consequently, enhancing the efficiency and sustainability of CO₂^•− generation from these acids plays a significant role in modulating nitrite reduction [25]. As shown in Fig. 1, the group of common small molecular carboxylic acids comprises formic acid (FA), oxalic acid (OA), and TA. Considering the structural and property aspects of carboxylic acids, TA presents distinct advantages. Primarily, TA demonstrates heightened reducibility and possesses two carboxyl groups, potentially yielding a greater number of CO₂^•− during activation. Furthermore, with lower toxicity (Fig. 1), TA exhibits an acute oral toxicity LC₅₀ value ranging from 2000 mg/kg to 5000 mg/kg (rat), significantly exceeding the values of formic acid (730 mg/kg) and oxalic acid (375 mg/kg), thereby signifying its reduced potential hazard [26]. Overall, TA serves as an effective and eco-friendly precursor for CO₂^•−, highlighting its promising potential for application in the treatment of nitrite-containing wastewater.

  Figure 1

  

  Figure 1.  Acute oral toxicity of carboxylic acids.

  DownLoad: Full-Size Img PowerPoint

  To confirm the excellent characteristics of TA, a UV/Fe³⁺/TA system was fabricated to explore the activation of TA in the production of CO₂^•− and the reduction of nitrite under various factors (Fe³⁺ dosage, TA dosage and pH). The details are shown in Supporting information.

  In the UV/Fe³⁺/TA system, Fe³⁺ primarily promotes the CO₂^•− activation from carboxylic acids through light-induced electron transfer and ^•OH activation originating from its own photoactivity [25]. To examine the influence of Fe³⁺, the impact of different Fe³⁺ dosage on the reduction of nitrite ions was investigated. As the Fe³⁺ dosage increases, the removal rate of NO₂⁻-N continuously rises, with pseudo-first-order rate constant increasing from 4.48 h^–1 to 6.67 h^–1, suggesting that elevating the Fe³⁺ dosage can accelerate the reduction of NO₂⁻-N (Fig. 2a). As shown in Eqs. 1 and 2, upon exposure to ultraviolet light, Fe³⁺ engages in photoinduced electron transfer, facilitating the activation of TA to generate CO₂^•− and subsequently reducing the NO₂⁻ to form N₂ (Fig. 2b and Eq. 3). Consequently, increasing the Fe³⁺ dosage stimulates the production of CO₂^•−, thereby boosting the reduction rate of NO₂⁻-N.

  Figure 2

  

  Figure 2.  Effect of Fe³⁺ dosage on (a) NO₂⁻-N degradation; (b) Concentrations of NH₄⁺-N, NO₃⁻-N and NO_x at the reaction endpoint; (c) Selectivity of N₂. TA dosage: 10 mmol/L, pH 2.5.

  DownLoad: Full-Size Img PowerPoint

  Except for the main product, N₂, there are three byproducts, namely NH₄⁺-N, NO₃⁻-N, and NO_x, at the reaction endpoint under varying Fe³⁺ dosages, as shown in Fig. 2b. When the Fe³⁺ dosage is less than 4 mmol/L, the NH₄⁺-N concentration remains negligible until the dosage reaches 5 mmol/L, producing around 1 mg N/L of NH₄⁺-N. This effect arises due to the substantial generation of CO₂^•− at the 5 mmol/L dosage, causing an excessive reduction of NO₂⁻-N to NH₄⁺-N. On the other hand, NO₃⁻-N demonstrates an initial decline followed by an increase on the graph. The minimum production of NO₃⁻-N is observed at a Fe³⁺ dosage of 3 mmol/L. This trend prompts speculation regarding the occurrence of reactions outlined in Eqs. 3 and 4. When subjected to ultraviolet light, Fe³⁺ undergoes self-photoreduction and generate ^•OH, which can activate TA to produce CO₂^•−. Nevertheless, ^•OH also exhibits potent oxidative characteristics, leading to the oxidation of NO₂⁻-N and the generation of NO₃⁻-N. Therefore, there is a trade-off effect between the accelerating generation of CO₂^•− and enhanced production of ^•OH with the elevated Fe³⁺ undergoes, which induces the trend of NO₃⁻-N concentration exhibits as Fig. 2b. Specifically, with increasing Fe³⁺ dosage, the concentration of CO₂^•− rises. Simultaneously, the production of ^•OH remains relatively low. Consequently, NO₂⁻-N mainly reacts with CO₂^•− to form N₂, and concurrently, oxidatively generated NO₃⁻-N can also undergo a reduction reaction with CO₂^•−, resulting in a gradual decrease in the accumulation of NO₃⁻-N during the process. However, upon increasing the Fe³⁺ dosage to 4 mmol/L, the extent of Fe³⁺ photoreduction significantly increases, leading to a gradual rise in ^•OH production. This phenomenon, in turn, leads to the accumulation of oxidatively formed NO₃⁻-N [24]. Consequently, NO₃⁻-N exhibits an ascending trend in the system. Finally, concerning NO_x, its presence was not detected in the gas phase, indicating the absence of NO_x generation within this specific dosage range.

  (1)

  (2)

  (3)

  (4)

  Based on the above data, the variation in N₂ selectivity with Fe³⁺ dosages is depicted in Fig. 2c. With the Fe³⁺ dosage increasing from 1 mmol/L to 3 mmol/L, the N₂ selectivity rises from 85.5% to 91.2%. Conversely, as the Fe³⁺ dosage further increases from 3 mmol/L to 5 mmol/L, the N₂ selectivity decreases from 91.2% to 82.4%. This trend parallels the variation tendencies of the nitrogen byproducts in Fig. 2b. It particularly aligns with the trend of NO₃⁻-N production, influenced by the trade-off effect. Through comprehensive analysis, it is evident that at a Fe³⁺ dosage of 3 mmol/L, the production of nitrogen byproducts is minimized, and the N₂ selectivity reaches its peak. Consequently, the Fe³⁺ dosage of 3 mmol/L is chosen as the optimal condition.

  Except for the Fe³⁺ dosage, TA functions as a precursor for CO₂^•− generation in UV/Fe³⁺/TA system and was considered as a key factor. As illustrated in Fig. 3a, an increase in TA dosage results in a significant improvement in the removal rate of NO₂⁻-N. This observation suggests that higher TA dosages can significantly enhance the generation rate of CO₂^•−, consequently enhancing the reduction of NO₂⁻-N. Fig. 3b illustrates the concentrations of three byproducts with changing TA dosages. As shown in Fig. 3b, NO₃⁻-N exhibits the highest production as a byproduct at a TA dosage of 5 mmol/L. With an increase in TA dosage, the elevated production of CO₂^•− promotes the reduction of NO₃⁻-N, resulting in a reduction of its concentration. However, the photocatalytic Fe³⁺-generated ^•OH also facilitates the transformation of NO₂⁻-N into NO₃⁻-N, thus maintaining NO₃⁻-N at approximately 3–3.5 mg N/L at TA dosage ranging from 10 mmol/L to 40 mmol/L. Additionally, as the TA dosage increases to 30 mmol/L or higher, the system generates additional NH₄⁺-N and NO_x, with their concentrations correspondingly increasing as TA dosage rises. This phenomenon could be attributed to the excessive reducing agents TA and CO₂^•− over-reducing NO₂⁻-N into NH₄⁺-N or NO_x. Finally, according to Fig. 3c, at a TA dosage of 5 mmol/L, the N₂ selectivity is 79.9%. As the dosage increases to 10 mmol/L, the N₂ selectivity significantly rises to 91.2%. However, with continued dosage escalation, the N₂ selectivity gradually decreases. Consequently, a TA dosage of 10 mmol/L is selected as the optimal condition to investigate the impact of pH on the performance.

  Figure 3

  

  Figure 3.  Effect of TA dosage on (a) NO₂⁻-N degradation; (b) Concentrations of NH₄⁺-N, NO₃⁻-N and NO_x at the reaction endpoint; (c) Selectivity of N₂. Fe³⁺ dosage: 3 mmol/L, pH 2.5.

  DownLoad: Full-Size Img PowerPoint

  As illustrated in Fig. 4a, the reduction rate of NO₂⁻-N exhibits a decreasing trend as pH increases, especially, this value significantly diminishes when pH ≥ 4.5. This is primarily because (1) the reduction of NO₂⁻-N requires lower pH levels to provide sufficient H⁺ as proton to participate in the reaction, as shown in Eq. 2; (2) with increasing pH, the increasing OH^– precipitates Fe³⁺ into Fe(OH)₃ with lower photocatalytic activity. After the reduction reaction of NO₂⁻-N, the trends of three dominant byproducts are shown in Fig. 4b. It can be observed that when pH > 3.5, the byproduct concentrations are low. However, as the pH gradually decreases to 3.5 and below, the endpoint concentrations of NO₃⁻-N and NO_x significantly increase, while NH₄⁺-N, although slightly elevated, remains at a relatively low concentration level. The occurrence of these phenomena may be attributed to the instability of trivalent NO₂⁻-N to the formation of NO₃⁻-N and NO by dimerization reactions under highly acidic conditions. Based on the above data, the N₂ selectivity was calculated, as illustrated in Fig. 4c. At a pH of 1.5, the prevalence of dimerization reactions results in the generation of a significant amount of NO₃⁻-N and NO_x, leading to lower N₂ selectivity. When the pH is increased, the photochemical reduction of NO₂⁻-N gains higher advantage over its inherent dimerization reaction, resulting in the higher N₂ selectivity. Considering the goal of efficiently treating harmful NO₂⁻-N, it was determined that under conditions of Fe³⁺ dosage at 3 mmol/L, TA dosage at 10 mmol/L, and pH at 2.5, it was possible to completely remove 30 mg N/L of NO₂⁻-N, with 91.2% of NO₂⁻-N being converted into N₂. However, the UV/TA/Fe³⁺ system also exhibits the ability to eliminate NO₂⁻-N at higher pH levels with satisfactory nitrogen selectivity. If the elimination rate of NO₂⁻-N is less favorable than the selectivity, a higher pH may also represent an optimal condition.

  Figure 4

  

  Figure 4.  Effect of pH on (a) NO₂⁻-N degradation; (b) Concentrations of NH₄⁺-N, NO₃⁻-N and NO_x at the reaction endpoint; (c) Selectivity of N₂. Fe³⁺ dosage: 3 mmol/L, TA dosage: 10 mmol/L.

  DownLoad: Full-Size Img PowerPoint

  According to Section S3 in Supporting information, it can be concluded that only when UV, TA, and Fe³⁺ coexist, TA can be promptly activated to form CO₂^•−, thus facilitating the efficient conversion of NO₂⁻-N to N₂. To further investigate the roles of the radicals in the mechanism, quenching experiments were conducted to confirm the presence and contributions of the related radicals. In these experiments, methyl violet (MV) and isopropanol (IPA) were selected as quenchers for CO₂^•− and ^•OH, respectively [24]. Fig. 5a reveals a notable reduction in the efficiency of NO₂⁻-N degradation upon the addition of MV, indicating that CO₂^•− plays a dominant role as the key radical in NO₂⁻-N reduction. In contrast, the impact of IPA addition on NO₂⁻-N removal, as evident in Fig. 5b, appears comparatively limited, signifying that the activation of TA through the ^•OH pathway is less potent compared to its direct activation by UV and Fe³⁺. To provide additional clarity on the presence of these radicals, the electron paramagnetic resonance (EPR) technique was employed, utilizing DMPO as the spin-trapping agent. As depicted in Fig. 5c, the quadruplicate signal can be attributed to the DMPO-^•OH adducts, and the additional faint signals correspond to the DMPO-CO₂^•− [27]. Moreover, as illustrated in Fig. 5d, the influence of these two quenchers on N₂ selectivity becomes more evident. The addition of IPA results in a reduction of N₂ selectivity to approximately 87%, while the introduction of MV sharply decreases N₂ selectivity to around 11%. This observation underscores the predominant role of CO₂^•− in facilitating the successful reduction of NO₂⁻-N to N₂. Therefore, CO₂^•− stands as a critical radical determining whether NO₂⁻-N can be effectively and selectively reduced to N₂.

  Figure 5

  

  Figure 5.  Free radical quenching experiments: (a) NO₂⁻-N concentration in CO₂^•− quenching experiment; (b) NO₂⁻-N concentration in ^•OH quenching experiment; (c) EPR free radical detection; (d) selectivity of N₂ in the presence of quenching agent (A: No quencher, B: 0.5 mmol/L IPA, C: 1.0 mmol/L IPA, D: 0.5 mmol/L MV, E: 1.0 mmol/L MV).

  DownLoad: Full-Size Img PowerPoint

  On the other hand, in the reduction mechanism, UV irradiation accompanied with Fe³⁺ would result in the decarboxylation of TA and the generation of CO₂^•− (Eq. 1). Accordingly, both Fe³⁺ and TA undergo transformations with the proceeding of NO₂⁻-N reduction. Fig. 6a illustrates the relationship between Fe³⁺, Fe²⁺, and the total iron ion concentration. Before the reduction of NO₂⁻-N, the total iron ion concentration matches that of Fe³⁺, with no presence of Fe²⁺ in the solution system. As the reduction reaction progresses, the total iron ion content remains relatively constant, while the Fe³⁺ content gradually decreases, and the Fe²⁺ content increases. This indicates the conversion of Fe³⁺ to Fe²⁺ in the system, indicating that Fe³⁺ undergoes photoinduced electron transfer and photoreduction reactions, as described in Eqs. 1 and 3. Fig. 6b displays the variation in TOC concentration over time. As the reaction proceeds, TOC gradually decreases, signifying a reduction in the organic carbon content within TA. In conjunction with Fig. 6c, after UV irradiation, the carboxyl group intensity at 1264 cm⁻¹ in the system is significantly lower compared to the TA and TA/Fe³⁺ groups, which further demonstrates that TA undergoes continuous decarboxylation and activation catalyzed by Fe³⁺ and UV [24]. Based on the HR-MS spectrum (Fig. S3 in Supporting information), the decarboxylation of TA may involve carboxylic intermediates such as oxalic acid and 2-oxoacetic acid, ultimately leading to the formation of CO₂^•−. While HR-MS captured some of these intermediates, it is important to note that their concentrations are relatively low and most of the tartaric acid undergoes conversion to CO₂^•−. These findings align with those presented in Fig. 6b, showing a decrease in TOC concentration in the UV/TA/Fe³⁺ system post-activation.

  Figure 6

  

  Figure 6.  (a) Concentration of iron species during reduction. (b) TOC concentration during reduction. (c) FTIR spectra of different systems.

  DownLoad: Full-Size Img PowerPoint

  In summary, the reduction mechanism can be deduced as follows: Under UV irradiation, Fe³⁺ activates the carboxyl groups in TA through photoinduced electron transfer, generating CO₂^•− [24,28,29]. Additionally, Fe³⁺ undergoes photodegradation reactions in the presence of UV and produces ^•OH, which can also activate the carboxyl groups in TA to form CO₂^•− radicals [25]. However, it may also oxidize NO₂⁻-N into NO₃⁻-N [24]. Furthermore, TA can undergo self-photolysis under UV irradiation, albeit to a lesser extent. Once CO₂^•− is generated in the solution, it primarily reduces NO₂⁻-N to N₂, however, minor byproducts like NH₄⁺-N, NO₃⁻-N, NO_x could also be undesirably generated due to excessive NO₂⁻-N reduction, dimerization reactions, and ^•OH oxidation during the process [24,29].

  According to the above discussions, the fabricated UV/TA/Fe³⁺ system efficiently reduces NO₂⁻-N to N₂. However, many real water bodies containing NO₂⁻-N are not simple single systems and their compositions are rather complex. Considering seawater as a potential flue gas absorption medium, it generates nitrified seawater containing NO₂⁻-N [30]. Thus, nitrified seawater serves as a typical real NO₂⁻-N pollution system. Therefore, we employ the UV/TA/Fe³⁺ system to assess its performance in treating real nitrified seawater. Table S1 (Supporting information) indicates the concentrations of ions other than NO₂⁻ in this water body.

  As shown in Figs. 7a and b, at an initial NO₂⁻-N concentration of 30 mg N/L, the UV/TA/Fe³⁺ system rapidly removed NO₂⁻-N from nitrified seawater, achieving a N₂ selectivity of 90.7%, which is nearly identical to the optimal performance in aqueous solutions. To further confirm the efficiency of the UV/TA/Fe³⁺ system in treating real water samples, the influence of various ions present in seawater on the selectivity of NO₂⁻-N reduction was investigated. By introducing ions at concentrations consistent with those in nitrified seawater (Table S1 in Supporting information) in the aqueous system, as shown in Figs. 7c and d, it was observed that a range of common anionic and cationic species did not inhibit the reduction of NO₂⁻-N. This further substantiates the potential of the UV/TA/Fe³⁺ system in treating real wastewater. Furthermore, as the initial NO₂⁻-N concentration increased, the rate of NO₂⁻-N reduction in nitrified seawater gradually decreased from 5.61 h^–1 at 30 mg N/L NO₂⁻-N to 3.14 h^–1 at 110 mg N/L NO₂⁻-N. Nevertheless, as depicted in Fig. 7a, within the concentration range of 30–110 mg N/L NO₂⁻-N, the UV/TA/Fe³⁺ system effectively achieved complete removal of nitrites. Finally, Fig. 7b illustrates a slight decrease in N₂ selectivity with increasing initial NO₂⁻-N concentration. This primarily results from the reduction in CO₂^•− available per unit concentration of NO₂⁻-N as the NO₂⁻-N concentration rises, leading to reduced reduction efficiency.

  Figure 7

  

  Figure 7.  (a) Effect of initial concentrations on NO₂⁻-N degradation in nitrified seawater. (b) Effect of initial concentrations on N₂ selectivity. (c) Effect of cations on N₂ selectivity. (d) Effect of anions on N₂ selectivity.

  DownLoad: Full-Size Img PowerPoint

  In conclusion, this study proposes an eco-friendly approach employing TA as a precursor for nitrite reduction. Under the activation of Fe³⁺ and ultraviolet light, this system generates highly reducing CO₂^•− for selective reduction of nitrites to N₂. The impact of TA dosage, Fe³⁺ dosage and pH on the reduction process is systematically examined. The findings reveal that the UV/TA/Fe³⁺ system can completely eliminate nitrites from the wastewater within 1 h, achieving an impressive N₂ selectivity of 91.2%. Finally, this system is applied to the reduction of nitrites in nitrified seawater, demonstrating that even in complex seawater systems, nitrite removal is feasible, with N₂ selectivity consistently exceeding 90%.

  Declaration of competing interest

  The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

  Acknowledgments

  This research was financially supported by National Natural Science Foundation of China (No. 22208081), Central Guidance on Local Science and Technology Development Fund of Hebei Province (No. 226Z3102G) and Fundamental Research Funds of Hebei University of Technology (No. JBKYTD2001).

  Supplementary materials

  Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.cclet.2024.109639.

  
  
• 1. [1]

     X. Chen, X. Tong, J. Gao, et al., Environ. Sci. Technol. 56 (2022) 4542–4552. doi: 10.1021/acs.est.2c00786

  2. [2]

     D.B. Gingerich, M.S. Mauter, Environ. Sci. Technol. 54 (2020) 3783–3792. doi: 10.1021/acs.est.9b07433

  3. [3]

     C.P. Bondonno, L. Zhong, N.P. Bondonno, et al., Trends Food Sci. Technol. 135 (2023) 57–73. doi: 10.1016/j.tifs.2023.03.014

  4. [4]

     M. Carlström, C.H. Moretti, E. Weitzberg, J.O. Lundberg, Free Radic, Bio. Med. 161 (2020) 321–325.

  5. [5]

     O. El hani, A. Karrat, K. Digua, A. Amine, Spectrochim. Acta A 267 (2022) 120574. doi: 10.1016/j.saa.2021.120574

  6. [6]

     H. Min, Z. Han, M. Wang, et al., Inorg. Chem. Front. 7 (2020) 3379–3385. doi: 10.1039/D0QI00780C

  7. [7]

     M.R. Awual, M.M. Hasan, A. Islam, et al., J. Clean. Prod. 228 (2019) 778–785. doi: 10.1016/j.jclepro.2019.04.280

  8. [8]

     X. Xiang, J. Wang, Q. Liu, et al., Sep. Purif. Technol. 275 (2021) 119195. doi: 10.1016/j.seppur.2021.119195

  9. [9]

     M. Jiang, X. Zheng, Y. Chen, Water. Res. 169 (2020) 115242. doi: 10.1016/j.watres.2019.115242

  10. [10]

      B. Zhang, Y. Jiang, K. Zuo, et al., J. Hazard. Mater. 382 (2020) 121228. doi: 10.1016/j.jhazmat.2019.121228

  11. [11]

      X. Zhang, Y. Wang, Y. Wang, et al., Chem. Commun. 58 (2022) 2777–2787. doi: 10.1039/D1CC06690K

  12. [12]

      Y. Zhang, Y. Wang, L. Han, et al., Angew. Chem. Int. Ed. 62 (2023) e202213711. doi: 10.1002/anie.202213711

  13. [13]

      S.R. Huang, P.J. Huang, J. Environ. Chem. Eng. 10 (2022) 106902. doi: 10.1016/j.jece.2021.106902

  14. [14]

      S. Lee, Y. Lee, W. Choi, Appl. Catal. B 327 (2023) 122432. doi: 10.1016/j.apcatb.2023.122432

  15. [15]

      N. Hoinkis, M.I. Litter, Ind. Eng. Chem. Res. 61 (2022) 16408–16417. doi: 10.1021/acs.iecr.2c02296

  16. [16]

      Q. Wang, L. Wang, S. Zheng, et al., J. Hazard. Mater. 451 (2023) 131149. doi: 10.1016/j.jhazmat.2023.131149

  17. [17]

      J. Bi, X. Huang, J. Wang, et al., Chem. Eng. J. 366 (2019) 50–61. doi: 10.1016/j.cej.2019.02.017

  18. [18]

      S. Lu, L. Shen, X. Li, et al., J. Clean. Prod. 378 (2022) 134589. doi: 10.1016/j.jclepro.2022.134589

  19. [19]

      X. Yu, D. Cabooter, R. Dewil, J. Hazard. Mater. 357 (2018) 81–88. doi: 10.1016/j.jhazmat.2018.05.049

  20. [20]

      C.M. Hendy, G.C. Smith, Z. Xu, T. Lian, N.T. Jui, J. Am. Chem. Soc. 143 (2021) 8987–8992. doi: 10.1021/jacs.1c04427

  21. [21]

      J. Guo, J. Deng, B. An, et al., Chemosphere 295 (2022) 133785. doi: 10.1016/j.chemosphere.2022.133785

  22. [22]

      H. Shi, C. Li, L. Wang, et al., Sep. Purif. Technol. 300 (2022) 121854. doi: 10.1016/j.seppur.2022.121854

  23. [23]

      S. Wang, Y. Zhao, J. Feng, et al., Chem. Eng. J. 455 (2023) 140787. doi: 10.1016/j.cej.2022.140787

  24. [24]

      B. An, H. He, B. Duan, J. Deng, Y. Liu, Chemosphere 278 (2021) 130388. doi: 10.1016/j.chemosphere.2021.130388

  25. [25]

      B. Wang, B. An, Y. Liu, J. Chen, J. Zhou, Sep. Purif. Technol. 248 (2020) 117061. doi: 10.1016/j.seppur.2020.117061

  26. [26]

      Merck Millipore, Safety data sheet.

  27. [27]

      Q. Zhou, W. Niu, Y. Li, X. Li, J. Clean. Prod. 258 (2020) 120790. doi: 10.1016/j.jclepro.2020.120790

  28. [28]

      J. Chen, R. Zhang, D. Chen, J. Liu, S. Chen, J. Water Process Eng. 40 (2021) 101934. doi: 10.1016/j.jwpe.2021.101934

  29. [29]

      Z. Shi, F. Wang, Q. Xiao, S. Yu, X. Ji, Catalysts 12 (2022) 348. doi: 10.3390/catal12030348

  30. [30]

      Y. Zhao, Y. Cheng, T. Zhang, et al., Fuel 340 (2023) 127436. doi: 10.1016/j.fuel.2023.127436

• 

  Figure 1  Acute oral toxicity of carboxylic acids.

  下载: 全尺寸图片 幻灯片
  

  Figure 2  Effect of Fe³⁺ dosage on (a) NO₂⁻-N degradation; (b) Concentrations of NH₄⁺-N, NO₃⁻-N and NO_x at the reaction endpoint; (c) Selectivity of N₂. TA dosage: 10 mmol/L, pH 2.5.

  下载: 全尺寸图片 幻灯片
  

  Figure 3  Effect of TA dosage on (a) NO₂⁻-N degradation; (b) Concentrations of NH₄⁺-N, NO₃⁻-N and NO_x at the reaction endpoint; (c) Selectivity of N₂. Fe³⁺ dosage: 3 mmol/L, pH 2.5.

  下载: 全尺寸图片 幻灯片
  

  Figure 4  Effect of pH on (a) NO₂⁻-N degradation; (b) Concentrations of NH₄⁺-N, NO₃⁻-N and NO_x at the reaction endpoint; (c) Selectivity of N₂. Fe³⁺ dosage: 3 mmol/L, TA dosage: 10 mmol/L.

  下载: 全尺寸图片 幻灯片
  

  Figure 5  Free radical quenching experiments: (a) NO₂⁻-N concentration in CO₂^•− quenching experiment; (b) NO₂⁻-N concentration in ^•OH quenching experiment; (c) EPR free radical detection; (d) selectivity of N₂ in the presence of quenching agent (A: No quencher, B: 0.5 mmol/L IPA, C: 1.0 mmol/L IPA, D: 0.5 mmol/L MV, E: 1.0 mmol/L MV).

  下载: 全尺寸图片 幻灯片
  

  Figure 6  (a) Concentration of iron species during reduction. (b) TOC concentration during reduction. (c) FTIR spectra of different systems.

  下载: 全尺寸图片 幻灯片
  

  Figure 7  (a) Effect of initial concentrations on NO₂⁻-N degradation in nitrified seawater. (b) Effect of initial concentrations on N₂ selectivity. (c) Effect of cations on N₂ selectivity. (d) Effect of anions on N₂ selectivity.

  下载: 全尺寸图片 幻灯片
•