Emerging investigator series: contributions of reactive nitrogen species to transformations of organic compounds in water: a critical review

Rachel C. Scholes *
Department of Civil Engineering, University of British Columbia, Vancouver, British Columbia V6T 1Z4, Canada. E-mail: rachel.scholes@ubc.ca

Received 8th March 2022 , Accepted 4th May 2022

First published on 5th May 2022


Reactive nitrogen species (RNS) pose a potential risk to drinking water quality because they react with organic compounds to form toxic byproducts. Since the discovery of RNS formation in sunlit surface waters, these reactive intermediates have been detected in numerous sunlit natural waters and engineered water treatment systems. This critical review summarizes what is known regarding RNS, including their formation, contributions to contaminant transformation, and products resulting from RNS reactions. Reaction mechanisms and rate constants have been described for nitrogen dioxide (˙NO2) reacting with phenolic compounds. However, significant knowledge gaps remain regarding reactions of RNS with other types of organic compounds. Promising methods to quantify RNS concentrations and reaction rates include the use of selective quenchers and probe compounds as well as electron paramagnetic resonance spectroscopy. Additionally, high resolution mass spectrometry methods have enabled the identification of nitr(os)ated byproducts that form via RNS reactions in sunlit surface waters, UV-based treatment systems, treatment systems that employ chemical oxidants such as chlorine and ozone, and certain types of biological treatment processes. Recommendations are provided for future research to increase understanding of RNS reactions and products, and the implications for drinking water toxicity.

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Rachel C. Scholes

Rachel Scholes is an Assistant Professor in the Department of Civil Engineering at the University of British Columbia, where she leads a research group in environmental engineering. Her research focuses on transformations of trace organic contaminants and optimizing trace contaminant removal in engineered and nature-based water treatment systems. She earned an M.S. and Ph.D. in Environmental Engineering from the University of California, Berkeley, and a B.S. in Chemical Engineering from Northwestern University.

Environmental significance

Reactive nitrogen species (RNS) react with organic compounds to form nitrated and nitrosated byproducts that are often more toxic than the parent compounds. RNS formation occurs in sunlit surface waters and in water treatment systems, where RNS reactions contribute to the transformation of trace organic contaminants and the formation of byproducts from dissolved organic matter. This critical review summarizes RNS reactions and assesses their relevance to organic compound transformations in environmental systems and water treatment processes, and provides recommendations for research to further our understanding of RNS reactions and impacts.


The supply of nitrogen in fixed forms, such as ammonia and nitrate, has approximately doubled since pre-industrial levels due to human activity,1 resulting in increased nitrogen pollution in aqueous environments.2–4 Inputs from agricultural fertilizer application and municipal wastewater effluent discharges have increased nitrate concentrations in surface and groundwaters over the past several decades.5 For instance, in the UK, nitrate concentrations in major rivers have increased to levels often exceeding the maximum nitrate concentration of the 1991 Nitrates Directive (i.e., 11.3 mg NO3-N/L) due to nonpoint source inputs from agriculture.6 Global estimates of nitrate pollution indicate similar concentrations may occur in agriculturally-productive regions in the Southern U.S., China, and Europe.7 Nitrogen emissions to surface waters also occur via discharge of municipal wastewater effluent, which often leads to elevated nitrate concentrations in effluent-dominated streams downstream of cities. During seasonal low-flow conditions, approximately 60% of regulated surface water discharges in the U.S. discharge to streams providing less than 10-fold dilution.8 When nitrate is released from wastewater treatment plants employing nitrification, discharges can result in nitrate concentrations of several mg-N/L,9 and concentrations sometimes exceed the World Health Organization's drinking water nitrate limit of 10 mg N L−1 (e.g., in Saskatchewan, Canada10 and Colorado, U.S.11). These elevated nitrogen concentrations can cause human health effects, such as methemoglobinemia (i.e., blue baby syndrome),12 and contribute to eutrophication of surface waters.13,14 Recent studies indicate that nitrate in drinking water may be associated with additional health impacts (e.g., birth defects and cancer) at concentrations below 10 mg N L−1, although further research on the link between drinking water nitrate concentrations and these health outcomes is needed.15

An impact of increasing nitrogen concentrations in surface waters that has recently attracted more attention is the effect on photochemistry in both natural and engineered systems. In irradiated aqueous solutions, nitrate (NO3) and nitrite (NO2) are sources of reactive intermediates that contribute to indirect phototransformation, in which organic contaminants are oxidized by photoproduced reactive intermediates.16,17 The reactive intermediates generated by irradiation of inorganic nitrogen include reactive oxygen species (e.g., hydroxyl radical, ˙OH) and reactive nitrogen species (RNS), such as nitrogen dioxide (˙NO2), nitric oxide (˙NO), and peroxynitrite (ONOO). These reactive intermediates contribute to the disappearance of parent organic compounds and the formation of transformation products from trace organic contaminants and dissolved organic matter (DOM).

RNS are of particular interest relative to other reactive intermediates because their reactions with organic compounds often result in the formation of toxic transformation products.18–20 RNS reactions with organic compounds result in nitrated products, such as nitroaromatic compounds, and compounds with aromatic nitroso or N-nitroso functional groups.18,21–23 N-nitroso compounds are associated with genotoxicity, mutagenicity, and carcinogenicity.24–27 N-nitrosamine toxicity has recently been of particular interest to the drinking water community because these compounds have been identified as an emerging class of nitrogenous disinfection byproducts.28–30 Meanwhile, nitrated compounds commonly undergo biological activation to form reactive oxygen species in vivo, resulting in oxidative stress, lipid peroxidation, and DNA oxidation.31 Nitroaromatic compounds such as nitrobenzenes and nitronaphthalenes, among others, are therefore also considered carcinogenic and mutagenic.31

This critical review provides an overview of what is known about the contribution of RNS to the transformation of organic compounds in surface waters and water treatment systems. RNS formation pathways, steady-state concentrations, and reactions with trace organic contaminants and DOM are discussed in the context of sunlit surface waters, UV-based water treatment, and other oxidative water treatment processes. Calculations using rate constants reported in the literature are used to provide insights into conditions conducive to RNS-based transformations, to provide estimates of the NO3 and NO2 concentrations required for these reactions to be important relative to other phototransformation pathways. Nitrated and nitrosated transformation products that have been identified experimentally are reviewed in the context of identifying common functional groups in organic compounds that are susceptible to reactions with RNS. Finally, critical knowledge gaps requiring further research and recommendations for future studies of RNS reactions are provided.

RNS in sunlit waters

RNS formation mechanisms

The initial photochemical formation of RNS in surface waters occurs via two pathways: direct photolysis of NO3 or NO2, and reactions of NO2 with ˙OH (Fig. 1). When irradiated by sunlight, both NO3 and NO2 absorb light at wavelengths greater than 280 nm, and undergo photolysis reactions forming oxide radical anion (˙O) and RNS (i.e., ˙NO and ˙NO2 from NO2 and NO3, respectively).17 ˙O is rapidly protonated to ˙OH at circumneutral pH (pKa = 11.8).32 Direct photolysis of NO3 also proceeds via a secondary pathway, which produces NO2 and atomic oxygen (O(3P)) at similar yields to the formation of ˙NO2 and ˙OH (i.e., the quantum yields of ˙NO2 and NO2 under sunlight irradiation of nitrate are approximately 1.35% and 1.1%, respectively).33 The resulting NO2 likely contributes to further RNS formation in irradiated NO3 solutions.34,35 A third NO3 photolysis pathway, resulting in formation of ONOO, occurs only under irradiation with wavelengths below 280 nm and is therefore not relevant to sunlight photolysis (Fig. 2).33 Under acidic conditions in the presence of NO2, ˙NO2 also forms in the absence of light via the thermal dismutation of HNO2, which can be a significant source of ˙NO2 in solutions with pH <5.18,20
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Fig. 1 Primary formation reactions of RNS from nitrate and nitrite photolysis in aqueous solutions. Photolysis pathways without a designated wavelength cutoff occur under sunlight irradiation and at wavelengths used in UV-based water treatment systems.

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Fig. 2 NO3 and NO2 molar absorptivity44 and sunlight irradiance spectrum from 280–480 nm.45

RNS formed by direct photolysis react further to produce secondary RNS. For example, ˙NO is a weak oxidant (E0 = 0.39 V)36 that accumulates in solution20,37 and reacts with ˙NO2 to form the nitrosating agent dinitrogen trioxide (N2O3).18 Additionally, ˙OH formed by NO2 photolysis reacts with residual NO2 with a bimolecular rate constant of approximately 1010 M−1 s−1 to form ˙NO2.38 When NO2 is the primary source of RNS, these reactions result in low steady-state concentrations of ˙OH and the formation of ˙NO2, ˙NO, and N2O3. Solutions containing both NO3 and NO2 result in a higher formation rate of ˙NO2 than solutions containing only one ˙OH source because NO2 reacts with the ˙OH produced by NO3 and NO2 photolysis.39

While photolysis of both NO3 and NO2 forms ˙NO2, the rate of RNS production is much higher for NO2 due to the overlap of its absorption spectrum with sunlight and the high quantum yield of RNS in sunlit systems (Fig. 2). NO3 absorbs less strongly than NO2 in the sunlight spectrum (i.e., at wavelengths >280 nm): the maximum molar extinction coefficient (ε) of NO3 in the sunlight spectrum is 7.0 M−1 cm−1 at approximately 300 nm and absorption is negligible above 330 nm. NO3 also exhibits low photolysis quantum yields (e.g., the quantum yield of ˙OH is 1.35%).33 In contrast, NO2 absorbs sunlight with an absorption maximum of 22.1 M−1 cm−1 at approximately 350 nm and absorbs at wavelengths up to 500 nm. NO2 also has a higher quantum yield of ˙OH of up to approximately 7% at wavelengths greater than 280 nm.40,41 As a result, NO2 may be as important to sensitized sunlight photolysis as NO3 even when it is present at 1–2 orders of magnitude lower concentrations.42,43

DOM in surface waters may be an additional source of NO2 and RNS. Photolysis of aqueous DOM solutions has been found to result in the formation of NO2 due to the photochemical release of DOM nitrogen,46 which may be attributable to the photolysis of nitroaromatic and N-nitroso moieties in DOM. For example, nitrophenols are a source of NO2 and HONO in sunlight-irradiated aerosols,47 and chloronitrobenzene released ˙NO2 when irradiated by 254 nm light in the presence of H2O2.48 The formation of NO2 has also been demonstrated to occur when N-nitroso precursors, such as the pesticide imidacloprid, undergo direct photolysis.49 Pesticide concentrations are typically too low for their photolysis to contribute significantly to inorganic nitrogen concentrations in surface waters, but this pathway may be relevant for N-nitroso precursors in DOM. These findings are consistent with the observation that yields of NO2 in irradiated NO3 solutions were higher in the presence of DOM from the Suwannee River,50 which contains approximately 1% nitrogen.51

DOM also affects the direct and indirect photolysis of NO2. When irradiated by sunlight, chromophoric moieties in DOM form excited triplet states (3CDOM*), which can oxidize NO2 to ˙NO2.52 With respect to direct photolysis, aliphatic DOM increased the yield of NO2 and decreased ˙NO2 formation in irradiated NO3 solutions.53 DOM also competes with NO3 and NO2 for UV light, potentially decreasing RNS formation via NO3 and NO2 photolysis. While DOM is potentially an important mediator of photochemical RNS reactions, including acting as both a source and sink of reactive oxygen species and RNS, its complex roles in RNS formation and scavenging are not yet fully understood.

RNS scavenging and steady-state concentrations

The concentrations of RNS in sunlit waters depend on the rates of formation and the rates of scavenging reactions that quench RNS, limiting their accumulation in solution. The predominant scavenging reactions include recombination with other RNS and hydrolysis. For instance, quenching of ˙NO2 proceeds via formation of N2O4 and its subsequent hydrolysis,54 N2O3 hydrolyzes to re-form NO2,20 and peroxynitrous acid rapidly self-isomerizes to form NO3:
2˙NO2 ↔ N2O4k1 = 4.5 × 108 M−1 s−1; k−1 = 6.9 × 103 s−1(1)
N2O4 + H2O → NO3 + NO2 + 2H+k2 = 1 × 103 s−1(2)
N2O3 + H2O → 2NO2 + 2H+k3 = 5.3 × 102 s−1(3)
ONOOH → NO3 + H+k4 = 0.7 s−1(4)

Dissolved inorganic carbon and DOM affect ˙NO2 formation and quenching as well: NO2 reacts with ˙CO3 to form ˙NO2 (kNO2,CO3 = 6.6 × 105 M−1 s−1), and ˙NO2 can be both formed and scavenged by DOM, as discussed above. More importantly, inorganic carbon and DOM scavenge ˙OH, which in turn decreases ˙NO2 formation (the implications of this phenomenon are discussed further below). Treating reactions (1–4) as the primary RNS quenching reactions and accounting for ˙OH scavenging produces estimates of steady-state concentrations of ˙NO2 ([˙NO2]ss) consistent with measurements using probe compounds.54 In this approach, [˙NO2]ss is calculated by rearranging the equations for the rates of formation from photolysis of NO3 and the reaction between NO2 and ˙OH, and scavenging via reactions (1) and (2):

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where Rform,NO3 → ˙OH is the formation rate of ˙OH from direct photolysis of NO3.

Using this approach, [˙NO2]ss has been estimated to be between 10−11 to 10−9 M in sunlit surface waters.17 For example, Minero et al. estimated an [˙NO2]ss of 3.5 × 10−10 M in the surface layer of sunlit lakes containing approximately 70 μM NO3 and 1 μM NO2,55 and several freshwater lakes with micromolar concentrations of NO3 and NO2 had [˙NO2]ss between 5 × 10−11 and 3.6 × 10−10 M.54 These calculations indicate that RNS may be present in surface waters at concentrations orders of magnitude higher than ˙OH and at concentrations similar to other less reactive species such carbonate radical (˙CO3).

Contribution of RNS to reactions in sunlit waters

Despite the well-known chemistry involved in the formation of RNS, isolating the contribution of RNS to organic compound phototransformation in sunlit waters remains a challenge, in part because rate constants for reactions between most organic compounds and RNS are unknown. Experimental data often indicate that the concentration of a target contaminant in irradiated solution decreases more quickly in the presence of NO3 or NO2, but this finding does not necessarily imply that reactions with RNS are the predominant removal mechanism since NO3 and NO2 are also sources of ˙OH. In the following, methods for quantifying the role of RNS in sunlight phototransformation are overviewed, followed by a discussion of phototransformation of phenols and other substituted aromatics that react with RNS in sunlit waters.
Quantification methods. There are at least three established approaches for determining the contribution of RNS reactions to contaminant degradation: (1) using radical scavengers to calculate the transformation rate attributable to ˙NO2, ˙NO, and ˙OH;36,43 (2) using probe compounds to assess [˙NO2]ss;54 (3) and quantifying concentrations of ˙NO with electron paramagnetic resonance (EPR) spectroscopy.56

Radical scavengers have been used to estimate the rate of trace organic contaminant removal by reaction with ˙OH, which is often calculated as the difference between removal rates in the absence and presence of an ˙OH scavenger (e.g., tert-butanol or isopropanol). In the presence of NO2, ˙OH scavengers inhibit formation of ˙NO2, such that the rate of contaminant removal via reaction with ˙OH cannot be estimated directly from ˙OH scavenging experiments. Similarly, there are no known scavengers for ˙NO2 or ˙NO that do not react with other radicals. However, ferulic acid, which reacts with ˙NO2 with a rate constant of 7.4 × 108 M−1 s−1 and with ˙OH with a rate constant of approximately 1010 M−1 s−1,57 has been used as a ˙OH and ˙NO2 scavenger in a handful of studies probing the role of ˙NO2 in contaminant transformation.58–60 Flavanone, which acts as a radical scavenger in the human body, has been used as a scavenger in experiments assessing the role of ˙NO,61 although flavonoids also react with ˙OH.62 While scavengers can provide evidence for a role of RNS by inhibiting RNS reactions, quantifying the roles of ˙OH and RNS in organic compound transformation requires estimates of steady-state reactive intermediate concentrations.

The steady-state concentrations of ˙OH and ˙NO2 can be determined using probe compounds that have known reaction rates with reactive intermediates. For example, the rate of nitrophenol formation in experiments spiked with excess phenol can be used to estimate [˙NO2]ss using the known rate constant for the reaction between ˙NO2 and phenol.54 Tyrosine has also been used as a probe compound for ˙NO2 by measuring the formation rate of nitrotyrosine.63 Similarly, the steady-state concentration of ˙OH can be estimated by tracking formation of para-hydroxybenzoic acid from benzoic acid64 or the degradation of nitrobenzene.36,65 Importantly, the use of nitrophenol formation to estimate [˙NO2]ss is most applicable in the absence of other reactive intermediates that react with phenol via H-abstraction to form phenoxyl radical (e.g., sulfate radical) since phenoxyl radical formation is the rate-limiting step in phenol nitration.18,66 Nitrophenol formation can also occur via an alternate pathway involving nitrosation of phenol (i.e., via reaction with N2O3) followed by oxidation of nitrosophenol to nitrophenol, indicating that nitrophenol formation may not exclusively indicate reaction with ˙NO2.18 As a result, researchers using nitrophenol formation rates to determine [˙NO2]ss should verify that formation is not due to the nitrosation mechanism, for instance by monitoring 4-nitrosophenol formation.

Although selective probe compounds have not been identified for ˙NO, its concentration can be measured by EPR using multiple spin trapping agents, which has enabled the direct detection of ˙NO in sunlit systems, and in the presence of monochloramine and UV light or biochar.37,56,61,67 While EPR can be used to quantify the steady-state concentrations of ˙NO, reactions between ˙NO and organic compounds are likely slow. For instance, nitrosation of phenol occurs via the reaction between phenol and N2O3, rather than ˙NO, at pH <10.8.18 The current lack of measured rate constants for ˙NO and N2O3 reactions with organic compounds hinders estimates of their contributions to contaminant transformation.

Reactions of phenols with RNS in sunlit waters. The predominant RNS-induced transformation pathway for phenol in sunlit waters occurs via two subsequent reactions with ˙NO2.18,20 In the first step, the reaction between phenol and ˙NO2 results in the abstraction of a hydrogen from phenol, leading to the formation of a phenoxyl radical. This step is followed by a reaction between phenoxyl radical and a second ˙NO2, forming a nitrocyclohexadienone intermediate, which reacts rapidly with water via proton transfer to yield 2- or 4-nitrophenol (Fig. 3).20,68 The first step is typically rate limiting and occurs with a rate constant of 3.2 × 103 M−1 s−1. The second step, which is a radical coupling reaction and occurs with a negligible energy barrier, occurs with a rate constant of approximately 3 × 109 M−1 s−1, approaching a diffusion-controlled rate in water.69 Importantly, other processes may contribute to phenoxyl radical formation and thereby increase the rate of nitrophenol formation. Specifically, photoproduced reactive intermediates (e.g., 3CDOM*) react with phenol to form phenoxyl radical,70,71 and some substituted phenolic compounds (e.g., nitrophenol and benzophenone) form phenoxyl radical via direct photolysis.47,72 The contributions of these pathways to nitrophenol formation are not yet fully understood.
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Fig. 3 Formation of nitrophenol via reaction with ˙NO2.

Since ˙NO2 is a relatively weak and selective oxidant (E0 = 1.03),36 it must be present at higher steady-state concentrations than other radicals to contribute significantly to phenol phototransformation kinetics. The reaction rate constant for ˙NO2 with phenol is 3.2 × 103 M−1 s−1, approximately a factor of 106 lower than the reaction rate constant for the reaction between phenol and ˙OH (6.6 × 109 M−1 s−1) and 103 lower than the reaction rate constant for phenol and ˙CO3 (4.9 × 106 M−1 s−1).66 Therefore, the steady-state concentration of ˙NO2 must be orders of magnitude higher than the concentrations of ˙OH and ˙CO3 for its contribution to transformation to be kinetically relevant.

Given [˙NO2]ss in natural waters in the range of 10−11–10−9 M,17,54 the half-life for phenol reacting with these levels of ˙NO2 would be approximately 2–250 days. In the same surface waters, [˙OH]ss was approximately 10−16 M, indicating that the half-life for phenol reacting with ˙OH would be approximately 12 days. The contribution of ˙NO2 relative to ˙OH is likely to be lower in surface waters where photosensitizers other than NO3 are responsible for most of the ˙OH produced. For instance, Vione et al. observed that NO3 concentrations in surface water were too low to contribute significantly to ˙OH production compared to the formation pathway via DOM-sensitized reactions when NO3 concentrations were approximately 0.1 mM in the presence of approximately 5 mg C L−1 DOM.73 Under such conditions, the contribution of ˙NO2 to phenol phototransformation is also likely small compared to the contribution of ˙OH.

To assess the contribution of RNS to the transformation of phenol under a range of inorganic nitrogen concentrations in surface waters, [˙NO2]ss and [˙OH]ss were estimated using Kintecus software (Fig. 4).74 In water containing only NO3, model results indicate that [˙NO2]ss would approach 10−9 M in the presence of up to 2 mM NO3 due to sunlight photolysis, while [˙OH]ss reaches 9 × 10−13 M in the absence of scavengers. Under these conditions, ˙NO2 has a small contribution to phenol phototransformation because ˙OH generated from NO3 photolysis outcompetes ˙NO2 reactions. In the presence of 1 mM inorganic carbon, scavenging of ˙OH by bicarbonate forms ˙CO3, resulting in [˙OH]ss and [˙CO3]ss of approximately 8 × 10−15 M and 10−10 M, respectively, in the presence of 2 mM NO3. Thus, in the presence of typical concentrations of inorganic carbon (e.g., 1 mEq L−1 alkalinity), ˙CO3 reacts directly with phenol and outcompetes both ˙OH and ˙NO2. These modeling results indicate that ˙NO2 generated from NO3 photolysis in sunlit systems is unlikely to be important to phenol disappearance even at elevated NO3 concentrations.

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Fig. 4 Modeled indirect phototransformation rates of phenol during sunlight irradiation via reactions with ˙CO3, ˙OH, and ˙NO2 from nitrate and nitrite.

In the presence of NO2, ˙OH is scavenged to form ˙NO2, resulting in a much higher [˙NO2]ss compared to [˙OH]ss. For instance, model results indicate that [˙NO2]ss is approximately 8 × 10−10 M and [˙OH]ss is approximately 8 × 10−16 M at a NO2 concentration of 100 μM in the absence of inorganic carbon (Fig. 4). In the presence of 1 mM inorganic carbon, [˙NO2]ss, [˙OH]ss, and [˙CO3]ss are approximately 5 × 10−10 M, 8 × 10−16 M, and 10−13 M, respectively. In this case, ˙NO2 contributes approximately 22% of phenol phototransformation. In the presence of 10 μM NO2 and 1 mM inorganic carbon, ˙NO2 accounts for approximately 5% of the phenol phototransformation rate.

Together, these findings indicate that RNS are rarely relevant to phenol phototransformation rates in sunlit waters but may be important in the presence of high concentrations of NO2. NO2 concentrations have been shown to reach 20–40 nM in coastal waters,46 and up to 9 μM in surface waters.34,35 Elevated NO2 concentrations may also occur in surface waters affected by agricultural runoff or wastewater discharges. For example, partial nitritation of wastewater can result in NO2 concentrations approaching 1 mM,75 indicating that wastewater treatment plants that use biological treatment technologies that release high concentrations of NO2 may result in higher contributions of RNS to organic compound transformations downstream.

Phenol transformation products. Despite the limited contribution of RNS to phenol phototransformation rates under most surface water scenarios, the formation of nitrophenols in irradiated surface water and groundwater samples has been observed. For instance, in surface waters containing 10−4 M NO3, nitrophenol reached concentrations up to 5 × 10−8 M when irradiated in a solar simulator.73 Formation of nitrated products has also been observed during sunlight photolysis of phenolic trace organic contaminants in the presence of NO3 or NO2. Nitration of phenolic moieties has been observed for bisphenol A,76 triclosan,76 orthophenyl phenol,76 benzophenones,22,77 and parabens (Fig. 5),76 although rate constants for the reactions between these compounds and ˙NO2 are unknown. DOM is likely to be a more significant precursor to nitrophenolic byproducts than trace organic contaminants, but the formation of nitrophenols from DOM has not been documented in sunlit systems.78 Photolysis of solutions containing NO2 (0.5 mM) and Suwannee River natural organic matter (SRNOM, 16 mg C L−1) by a UV lamp with a Pyrex cutoff filter that blocked irradiation at wavelengths <290 nm resulted in incorporation of NO2-N into SRNOM. However, evidence for nitration or nitrosation was not observed, while amide and lactam functional group formation was observed using 15N-NMR.51 Further research is needed to investigate whether DOM is a precursor to nitrophenolic compounds in sunlit waters.
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Fig. 5 Transformation products observed in the presence of RNS: (A) nitration products, (B) nitrosation of aromatic rings, (C) N-nitrosation. (S) and (UV) indicate formation in solutions irradiated by sunlight and UV lamps, respectively.
Reactions of substituted aromatics with RNS in sunlit waters. The kinetics of reactions between RNS and nonphenolic compounds have largely not been investigated in sunlit waters, but the formation of nitrated and nitrosated products has been observed for aromatic compounds with chlorine functional groups and for compounds that form phenolic intermediates under sunlight irradiation in the presence of NO2 or NO3. Several phenylurea herbicides with chlorinated aromatic functional groups, including chloroteluron, diuron, monuron, and linuron, undergo nitration in sunlight-irradiated NO3 or NO2 solutions.79–82 For other substituted aromatics, phenol intermediates may form first, and then undergo nitration. For example, sulfamethoxazole has been shown to undergo hydroxylation by ˙OH, forming phenolic intermediates, followed by nitration.43,83,84 Similarly, formation of 10−6 M nitrophenol was observed in solar simulator-irradiated groundwater containing approximately 5 mM NO3 and 2 × 10−5 M benzene, likely via phenol nitration.73

Although few examples have been observed to date, aromatic amines are likely susceptible to nitration by RNS as well. Primary aromatic amines undergo hydrogen abstraction more readily than phenols, including by ˙OH, producing anilino radicals (PhN˙H) that can react with ˙NO2 to form nitrated products.85 Reactions between ˙NO2 and several aromatic amine-containing pharmaceuticals (i.e., sulfamethoxazole, lamivudine, emtricitabine, and trimethoprim) occurred with rate constants ranging from 5.9 × 102–6.2 × 103 M−1 s−1 in experiments that used a solar simulator to irradiate solutions containing NO3 or NO2, and nitrated products of sulfamethoxazole without hydroxyl groups were identified.43 However, transformation products were not assessed for the other compounds and further studies are needed to determine whether this reaction mechanism occurs for additional aromatic amine-containing compounds.

RNS in UV-based water treatment processes

RNS formation mechanisms

RNS form readily in water treatment processes that use UV lamps for disinfection and advanced oxidation. Until recently, UV disinfection has been carried out with monochromatic, low-pressure UV (LPUV) lamps that emit light mainly at a wavelength of 254 nm. Medium-pressure UV (MPUV) lamps are also common and emit polychromatic spectra at wavelengths ranging from approximately 210–600 nm. When irradiated with the lower wavelengths of light emitted by UV lamps (i.e., wavelengths <280 nm), NO3 produces RNS and ˙OH at higher quantum yields compared to sunlit systems.33,86 The quantum yield of ˙NO2 and ˙OH from NO3 photolysis increases from 1.35% in the sunlight spectrum to approximately 9% at 254 nm,33,87 and over 20% at wavelengths <240 nm (Table 1).88 The quantum yield of NO2 is also higher at wavelengths <240 nm, which are emitted by MPUV lamps (estimated at 2.8–5.3%), compared to sunlight (1.1%), and NO2 has been shown to accumulate during UV treatment of NO3-containing waters.50,89 The resulting co-occurrence of NO3 and NO2 may contribute to higher rates of ˙NO2 formation, as described for sunlit systems.39 Together, these reactions result in potentially significant ˙NO2 formation in UV treatment systems. For instance, according to kinetic models, LPUV treatment of water containing 100 μM NO2 resulted in a steady-state ˙NO2 concentration of approximately 10−9 M,66 which is at the higher end of estimates calculated for sunlit surface waters.
Table 1 Molar absorptivity (ε) and quantum yields of reactive nitrogen species from nitrate photolysis during sunlight or UV irradiation at circumneutral pH
Species Reactive species quantum yields References
a Note: MPUV lamps emit at wavelengths from approximately 210–600 nm. Molar absorptivity and quantum yields are provided for MPUV wavelengths <240 nm.
>280 nm (Sunlight) ε ≤ 7 M−1 cm−1 254 nm (LPUV) ε ∼3 M−1 cm−1 <240 nm (MPUVa) ε >1000 M−1 cm−1 87
˙NO2 1.35% ±0.3% 9% ∼21% 33, 87 and 88
˙OH 1.35% ± 0.3% 9% ∼21%
ONOO <0.26% 6.5–10% 28% 33 and 86
NO2 1.1% ± 0.2% ∼0.1–3% 2.8–5.3% 33, 50 and 90

Two additional differences between sunlight and UV photolysis of NO3 are the molar absorptivity and formation of ONOO. The molar absorptivity of NO3 is low at the 254 nm irradiation wavelength of LPUV lamps (ε ∼3 M−1 cm−1) and in the sunlight spectrum (ε ≤7 M−1 cm−1), but increases drastically at wavelengths <240 nm (ε >1000 M−1 cm−1),88 resulting in higher RNS formation rates from NO3 in systems using MPUV. UV irradiation of NO3 also results in an additional photolysis pathway: conversion to ONOO. The formation of ONOO, which reacts with organic compounds via nitration, is negligible at wavelengths >280 nm but occurs with quantum yields of approximately 6.5–10% and 28% when NO3 is irradiated by wavelengths of 254 nm and <240 nm, respectively. ONOO also contributes to secondary RNS formation via its reaction with ˙OH, producing ONOO˙. The resulting ONOO˙ decomposes to form O2 and ˙NO, which may further react with ˙NO2 to yield N2O3.87 These additional RNS formation mechanisms, in combination with the high quantum yields and molar absorptivity of NO3 photolysis during UV irradiation, indicate the potential for nitr(os)ation reactions to occur during UV-based treatment processes. These reactions are most likely to occur in advanced oxidation processes (AOPs), which use much higher fluences of UV light than UV disinfection, although steady-state concentrations of RNS have not been reported in these systems.

Similar to the case of sunlit systems, RNS formation is likely affected by DOM in UV-based treatment systems. For instance, irradiation with a LPUV lamp induced the release of ˙NO2 from an organic precursor, chloronitrobenzene, and the resulting ˙NO2 further reacted with the parent compound to form dinitrochlorobenzene.48 The contribution of ˙NO2 from trace organic contaminants is likely minor, but nitroaromatic precursors in DOM may contribute significantly to RNS formation in UV-based treatment systems.

Contribution to trace organic contaminant transformation in UV disinfection systems. The fluence of UV light used for disinfection is relatively low (e.g., to receive credit for 4-log removal of viruses, the U.S. EPA requires a fluence of 186 mJ cm−2), typically resulting in limited removal of trace organic contaminants during UV disinfection. However, reactive intermediates produced from nitrate photolysis, including ˙OH and RNS, may result in trace organic contaminant transformation. A recent investigation using LPUV lamps with a fluence of 172 mJ cm−2 to irradiate water containing 10 mM NO3 resulted in over 80% removal of several trace organic contaminants. RNS reactions were the predominant transformation mechanism for three phenolic compounds (estrone, bisphenol A, and ethinylestradiol), were responsible for 20–70% of the transformation of triclosan and diclofenac, and contributed <50% of the transformation rates of atrazine, carbamazepine, and ibuprofen.36 The NO3concentrations amended in these experiments are much higher than those found in drinking water, such that further research is needed to assess whether significant trace organic contaminant transformation occurs during UV disinfection of NO3-containing water.

In addition to RNS from NO3 photolysis, low concentrations of NO2 (e.g., up to approximately 10 μM) formed during UV treatment may enhance RNS reaction rates. For example, in a MPUV system, concentrations of NO2 of up to 5–10 μM were detected when solutions containing 0.7 mM NO3 were irradiated with a UV fluence of 400 mJ cm−2.50 At a NO2 concentration of approximately 7 μM, RNS and ˙OH were responsible for over 40% transformation of several trace organic contaminants (e.g., BPA, BPS, carbamazepine) that were not susceptible to direct photolysis with a UV fluence of 3660 mJ cm−2.42 The role of NO2 as an intermediate to RNS formation during UV disinfection of NO3-containing waters requires further research.

Contribution to trace organic contaminant transformation in UV/H2O2 systems. The UV/H2O2 process was designed to oxidize trace organic contaminants via reactions with ˙OH produced by H2O2 photolysis. This process requires UV fluences an order of magnitude higher than those used for UV disinfection (e.g., >1000 mJ cm−2) and likely results in RNS reactions with organic compounds, but few studies have investigated the role of RNS in AOPs. During UV/H2O2 treatment of NO3-containing waters, NO3 can become the major source of ˙OH because of its high molar absorptivity, particularly when using MPUV lamps.88,91 ˙OH is typically scavenged primarily by DOM in UV/H2O2 treatment systems. However, in the presence of NO3, ˙OH is also scavenged by photo-produced NO2, resulting in lower [˙OH]ss in the presence of NO3, as evidenced by slower than expected degradation of ˙OH probe compounds.91 These findings are consistent with the conversion of ˙OH to RNS such as ˙NO2 in UV/H2O2 systems treating NO3-containing waters. Experiments investigating NO2 accumulation in UV/H2O2 systems indicated that hydrogen peroxide increases the overall yield of NO2 during UV photolysis of NO3, which may also increase ˙NO2 formation via ˙OH scavenging.89 However, the extent to which RNS contribute to trace organic contaminant transformation during UV/H2O2 treatment of NO3-containing waters is not yet known.
Nitr(os)ated product formation in UV treatment systems. The primary evidence that RNS reactions occur in UV disinfection and UV/H2O2 treatment systems is the formation of nitr(os)ated byproducts. In most studies to date, researchers have employed UV fluences higher than are typical in disinfection systems, with or without H2O2, and have investigated the formation of nitrated transformation products of trace organic contaminants. For example, in river water and wastewater samples spiked with 0.4–0.8 mM NO3 (i.e., 5–10 mg N L−1) and irradiated with LPUV lamps (approximately 2000 mJ cm−2), nitrated transformation products were observed for several phenolic trace organic contaminants, including ortho-phenylphenol, methyl paraben, propyl paraben, triclosan, and bisphenol A (Fig. 5).76 The nitrated products were also observed in unspiked nitrified wastewater effluent and river water containing approximately 0.08 and 0.1 mM NO3, respectively, but the yields of these products were not quantified due to a lack of available analytical standards. Nitrated transformation products of triclosan, bisphenol A, and naproxen also formed during LPUV irradiation in experiments containing 10 mM added NO3,36,83 and nitrated products of chlorophenol and salbutamol have been observed in UV-irradiated solutions.92,93 Transformation products of other aromatic compounds (e.g., carbamazepine and monuron) containing nitro groups have been observed following UV irradiation in the presence of NO2.42,81 As in sunlit systems, hydroxylation of substituted aromatics (e.g., naproxen and sulfamethoxazole) followed by nitration has also been observed in solutions irradiated by UV lamps.83,84

Nitrophenolic compound formation from DOM precursors is likely more prevalent than formation from trace organic contaminants and has been observed in studies investigating MPUV treatment of water containing NO3 at concentrations relevant to drinking water. For example, formation of nitrophenols was observed following LPUV/H2O2 treatment in the presence of approximately 0.8 mM NO3 and 2 mg C L−1 Pony Lake natural organic matter,94 or following MPUV irradiation of solutions containing approximately 0.2 mM NO3 and 2.2 mg C L−1 Pony Lake natural organic matter. Eight-three products that incorporated nitrogen from spiked NO3 into the molecular structure, including several nitrophenolic compounds, were observed by using 15NO3 in irradiation experiments to enable identification of transformation products containing 15N using high resolution mass spectrometry. Among the products, 4-nitrophenol, 4-nitrocatechol, and 2-methoxy-4,6-dinitrophenol were confirmed with analytical standards. 2-hydroxy-5-nitrobenzoic acid, 2,4-dinitrophenol, 5-nitrovanillin, 4-nitrobenzenesulfonic acid, and 4-nitrophthalic acid were later confirmed, and six additional nitrated products were tentatively identified.95 These nitrogen-containing products were also formed following treatment in a full-scale drinking water treatment plant employing a MPUV/H2O2 process.78

Nitrosation reactions have been observed for a handful of trace organic contaminants, but N-nitroso transformation products are susceptible to further transformation due to the photolability of the N-nitroso bond at wavelengths below approximately 275 nm.96–98 An N-nitroso product of mefenamic acid was observed in experiments using LPUV lamps to irradiate solutions containing 0.5 mM NO2,99 but the same product was not observed during LPUV treatment of nitrified wastewater treatment plant effluent containing approximately 0.3 mM NO3 and 10 μM NO2.23 N-nitrosation of diphenylamine has also been observed in a pilot-scale (5 m3 h−1) LPUV treatment system.23 In addition to N-nitrosation, nitrosation of aromatic compounds can occur, and the resulting C-nitrosation products are less susceptible than N-nitroso compounds to direct phototransformation.23 However, nitrosophenols undergo oxidation to the corresponding nitrophenol via reaction with dissolved oxygen,20 indicating that C-nitrosation products may further transform to more stable nitrated products. Nitrosated aromatic byproducts of naproxen and benzophenones have been observed following UV treatment,22,83 and the formation of nitrosophenol has been documented from DOM precursors.51 To date, although nitrosating RNS are expected to form during UV treatment of NO3-containing waters, few researchers have observed nitrosated product formation.

The formation of nitrogenous byproducts during UV treatment coincides with increases in toxicity, specifically mutagenicity, as measured by the Ames assay. Several nitroaromatic compounds and nitrosamines are known to produce a response in the Ames test,25,30,100 and multiple studies have shown that RNS reactions with DOM during UV irradiation result in the formation of nitrophenols and a simultaneous increase in the Ames response.51,78,94,95,101 Similarly, genotoxicity increased and nitrated products formed in a full-scale MPUV/H2O2 treatment process.78 The contribution of specific nitr(os)ated products to the increased genotoxicity is unknown. An effect-directed analysis study identified five potentially-genotoxic nitrated products that were present in relatively high concentrations in MPUV-treated samples in which mutagenicity was observed.95 Studies of nitrated compound toxicity in the human body have indicated that nitrated products generally contribute to genotoxicity, carcinogenesis and cytotoxicity,102 but further research into the toxicity of prevalent nitr(os)ated transformation products in UV-treated water is needed to assess the implications of RNS reactions in UV treatment systems.

Alternative advanced oxidation processes


Recently, RNS reactions were found to contribute to phenol transformation during UV treatment in the presence of peroxydisulfate (i.e., the UV/PDS process). In this AOP, sulfate radicals activate phenol towards nitration by reacting with phenol to form phenoxyl radical.66 This mechanism is more important in UV/PDS than UV/H2O2 because sulfate radicals preferentially react with phenol to form phenoxyl radical, whereas ˙OH reacts with phenol primarily via addition.103 Sulfate radical also contributes to nitration because it reacts with NO2 to form ˙NO2 (Fig. 6). As a result, nitrated product formation was observed when UV/PDS treatment was used in the presence of NO2 or NO3.66,103 RNS reactions could occur in other treatment systems where sulfate radicals are produced. For instance, nitrophenol formation from DOM precursors was observed during treatment with heat-activated PDS in the presence of 10 or 500 μM NO2,104 and conversion of chlorophenol to chloronitrophenol was observed in experiments with heat-activated persulfate in the presence of 2 mM persulfate and 100 μM NO2.105
image file: d2em00102k-f6.tif
Fig. 6 Formation pathways of RNS via reaction with radical oxidants (a), during UV/monochloramine treatment (b), and during chlorination (c).

UV/HOCl and UV/NH2Cl

AOPs based on the UV photolysis of chlorine and chloramine have been proposed as a strategy to achieve trace contaminant removal while providing residual disinfectant in a distribution system. These AOPs result in formation of disinfection byproducts (DBPs) due to the presence of reactive halogen species. NO2 may contribute to DBP formation during UV/chlorine treatment, which results in nitrogenous DBPs (N-DBPs) including trichloronitromethane.106,107 The formation of RNS has also been observed when chloramine-containing water is irradiated by UV light, which may occur in the UV/chloramine advanced oxidation process, when the UV/chlorine AOP is performed in the presence of ammonia, or during UV-AOP treatment of chloramine-treated water in water reuse treatment trains.56

Upon irradiation, chloramine dissociates into an amidogen radical (˙NH2) and a chloride radical (˙Cl) (Fig. 6). ˙NH2 is rapidly oxidized by dissolved oxygen to form aminylperoxyl radical (NH2OO˙), which reacts with water to form ˙NO.108 ˙NH2 has a low oxidation potential and is not a known nitrosating agent, whereas the reaction product of ˙NO with ˙NO2, N2O3, reacts with phenol via nitrosation.18

During UV/NH2Cl treatment, ˙NO formation co-occurred with the formation of nitrosated transformation products of naproxen and ibuprofen.56 Nitrosating RNS were also implicated in the transformation of estradiol and ethinylestradiol during LPUV/NH2Cl treatment, as indicated by a lack of effect of tert-butanol, which scavenges ˙OH and ˙Cl but not ˙NO or N2O3.109 Finally, estradiol and ethinylestradiol were transformed by reactions involving ˙NO in a treatment system containing biochar and NH2Cl, in which ˙NH2 was oxidized to form ˙NO.61 Together, these findings indicate that NH2Cl can be a source of RNS that produce nitrosated products, particularly in UV-irradiated treatment systems.

Emerging AOPs

RNS may play a role in other emerging AOPs that are not yet implemented at full scale.58 For example, sonolysis has been proposed as a desirable AOP because it does not require chemical inputs. In this process, reactive oxygen species are generated via cavitation of bubbles in solution, which also generates RNS from N2 at the air–water interface.58,110,111 Nitrated products of para-aminosalicylic acid and diphenylamine have been observed in experiments without dosed NO2 or NO3.110,112 Electrochemical treatment with boron-doped diamond electrodes in the presence of NO2 also resulted in nitrophenol formation, likely due to NO2 scavenging of ˙OH. However, experiments were performed with 1–100 mM NO2 and the relevance of RNS in electrochemical treatment systems with more realistic NO2 concentrations has not been demonstrated.113 Recently, peracetic acid has been studied as an alternative oxidant for use in AOPs. In the presence of peracetic acid and NO2, the transformation of sulfonamide antibiotics resulted in nitrated and nitrosated products. The reaction pathway did not involve reactions with ˙NO or ˙NO2 based on scavenging experiments using flavonoids and tyrosine as scavengers, indicating that reactions involving ONOO may have been responsible for sulfonamide nitr(os)ation in this system.114

RNS formation via reactions with chemical oxidants

In the absence of UV light, NO2 and ammonia can be activated by chemical oxidants to produce RNS. The formation of N-DBPs in chemical disinfection systems has gained increasing attention because of their toxicity and observed formation in drinking water treatment processes.29,115,116 However, the reactions responsible for N-DBP formation are typically not dominated by RNS, but rather oxidation of N-containing precursors by disinfectants.117,118 NO2 and ammonia can react with oxidants such as chlorine and ozone to form RNS, but their concentrations are generally too low in drinking water treatment systems to produce significant RNS. In the following, examples are discussed of conditions in which RNS are thought to play a role in the formation of N-DBPs. For a comprehensive discussion of N-DBP formation via other pathways, the reader is referred to recent reviews on the topic.29,115,116

Chlorination and chloramination

When NO2 is present during chlorination, it reacts with free chlorine to produce nitryl chloride (ClNO2), which may directly nitrate organic compounds or further react with NO2 to form N2O4 and ˙NO2 (Fig. 6).29,119 For instance, NO2 can be introduced to chlorinated pool water via urine, sweat, or NO3 photolysis.120 In chlorinated pools containing up to approximately 100 nM NO2, formation of nitrated N-DBPs has been attributed to reactions between amine precursors (e.g., from sweat and urine) and ClNO2-derived RNS.120 The proposed RNS reaction pathway is also consistent with the observed formation of other nitrosated byproducts in pool water.97,121 This pathway would only be significant during drinking water disinfection if the source water contains significant concentrations of NO2.

Recently, RNS from dichloramine decomposition were implicated in the formation of N-nitrosodimethylamine (NDMA), a highly toxic N-DBP.122 Under typical drinking water chloramination conditions, NDMA formation is thought to occur via the reaction of dichloramine with amine precursors.123,124 In a parallel reaction pathway, dichloramine hydrolysis produces nitroxyl (HNO), which reacts with dissolved oxygen to form peroxynitrous acid (ONOOH). Based on the results of experiments using uric acid as a ONOOH/ONOO scavenger, it was shown that NDMA formation occurred through a ONOOH/ONOO-mediated pathway at pH 7 (i.e., NDMA formation was not observed in the presence of uric acid). The NDMA yield also decreased by approximately 50% in the presence of uric acid at pH 9.122 These findings indicate that RNS generated by dichloramine decomposition may play a role in N-DBP formation, although the relevance compared to other reaction pathways likely depends on conditions including pH and the concentration of dissolved oxygen.


When NO2 is present during ozonation, it is primarily oxidized to NO3 by reaction with molecular ozone (kNO2,O3 = 3.7 × 105 M−1 s−1). A small fraction of the initial NO2 reacts with ˙OH generated by ozone decomposition. For organic contaminants with bimolecular reaction rate constants with ozone >105 M−1 s−1, and rate constants with ˙OH of 5–10 × 109 M−1 s−1, over 80% of compound removal was estimated to occur via reactions with ozone, and approximately 20% via reactions with ˙OH.125 NO2 has similar reaction kinetics with ozone and ˙OH, indicating that a similar fraction of NO2 (e.g., up to 20%) may react with ˙OH to form ˙NO2 during ozonation of NO2-containing waters. Where ozone and ˙NO2 co-occur, results obtained from atmospheric aerosol studies have indicated that ozone is capable of activating phenolic compounds toward nitration by ˙NO2. For instance, ozone reacts with tyrosine via H abstraction to form a tyrosyl radical, which then reacts with ˙NO2 to form nitrotyrosine.126,127 However, at NO2 concentrations encountered in drinking water treatment, NO2-derived RNS reactions are unlikely to contribute significantly to organic contaminant transformation compared to reactions with ozone and ˙OH.

Ozone can also activate ammonia to form RNS. Ammonia reacts with ozone to form hydroxylamine (NH2OH).116,128 However, the reaction between ammonia and ozone is slow (kNH3,O3 = 20.4 M−1 s−1), such that high levels of ammonia must be present for significant concentrations of NH2OH to form. When ammonia is present during ozonation, the reaction of dimethylamine with NH2OH can contribute to NDMA formation.116,128 This pathway is potentially relevant during ozonation of municipal wastewater effluent, and is consistent with the observed correlation between NDMA formation and ammonia concentrations in wastewater.128

RNS in biological treatment systems

RNS form as intermediates during biological wastewater treatment, and their formation coincides with the formation of nitr(os)ated transformation products. In particular, during ammonia oxidation, NO2 forms at concentrations approaching 20 μM,129 while NH2OH and ˙NO form at concentrations an order of magnitude lower than NO2 and can contribute to nitrosation.130 Researchers have observed nitr(os)ated products of trace organic contaminants in the presence of ammonia oxidation (e.g., nitration of acetaminophen and sulfamethoxazole, and nitration and nitrosation of diclofenac in nitrifying activated sludge,131–133 and nitration of phenols in ammonia-oxidizing batch reactors134). The same RNS present during ammonia oxidation also form during denitrification135 and nitrated transformation products have been observed under denitrifying conditions as well (e.g., nitration and nitrosation of diclofenac in soil aquifer treatment under denitrifying conditions).135

While RNS can contribute to nitration and nitrosation of organic compounds in biological treatment systems, their contribution to organic compound transformation is still a matter of debate because nitr(os)ated products also form via enzymatic pathways (e.g., oxidation of amine precursors by ammonia monooxygenase).129–131,136 The evidence for significant contributions from abiotic nitr(os)ation reactions in ammonia-oxidizing treatment systems is discussed at the molecular, cellular, and community level in a recent comprehensive review by Su et al.130 The authors conclude that RNS are unlikely to be important in typical nitrification/denitrification activated sludge treatment of municipal wastewater, where biotransformation by ammonia oxidizers is more important. Other microbial metabolisms also contribute to transformation in these systems, including heterotrophs with diverse metabolic capabilities.137 In contrast, RNS reactions appear to contribute significantly to transformation of some organic contaminants with primary aromatic amine or phenol functional groups in treatment systems with high NO2 concentrations, such as nitritation/denitritation systems treating high-strength wastewaters. The relevance of biologically-derived RNS in these systems depends heavily on reactor conditions affecting RNS concentrations, including microbial activity and pH.130,138

Outlook and future research needs

Further research is required to assess and mitigate the risk posed by RNS reactions in surface waters and water treatment systems. From a kinetic standpoint, understanding the reactivity of RNS with organic compounds is an ongoing challenge because isolating the role of RNS relative to other transformation pathways is not straightforward. Reaction rate constants between ˙NO2 and several phenolic compounds have been determined on the order of 104 M−1 s−1,43,93 but rates of reaction for additional phenolic compounds, including precursors in DOM, and for other classes of compounds, are unknown. Further attention should also be given to the wavelength dependence of RNS formation. RNS quantum yields vary significantly with wavelength, as demonstrated by the differences between RNS formation in sunlit and UV-based treatment systems. As new UV sources that emit at different wavelengths from traditional LPUV and MPUV lamps, such as UV-LEDs, are considered for implementation in water treatment systems, further research into the effect of irradiation wavelength on RNS formation is needed.

Regarding transformation products, further studies are needed to identify the products of RNS reactions with nonphenolic organic compounds and assess their toxicity. Research from the biochemistry literature can provide insights into potentially relevant precursors to nitr(os)ated products. ˙NO2, ˙NO, and ONOO occur in the human body and react with biomolecules and xenobiotics containing diverse functional groups.102,139,140 For instance, RNS react via nitration with the phenolic biomolecule tyrosine,126,141,142 with catecholamines such as noradrenaline,102 and with the bicyclic aromatic biomolecule tryptophan (Fig. 7).139,142 In some cases, heterocycles react with RNS to form N-nitroso products, as is the case for morpholine143 and 2-amino-3-methylimidzol[4,5-f]quinoline (IQ).102 N-nitrosation of other secondary aliphatic and aromatic amines has been observed, as has nitrosation of furans (e.g., cafestol).102 Aromatic, heterocyclic, and amine functional groups are abundant in DOM and in trace organic contaminants such as pesticides and pharmaceuticals, indicating that RNS reactions with these structures warrant further investigation in aqueous systems.

image file: d2em00102k-f7.tif
Fig. 7 Precursors and products of RNS reactions with biomolecules.

Two other classes of biomolecules that react with RNS and for which analogous precursors exist in DOM are polyunsaturated fatty acids and thiols.102 Polyunsaturated fatty acids react with RNS to form nitrated products,144–146 and related compounds with unsaturated alkyl chains, such as beta-carotene and retinoic acid, undergo nitration reactions in vivo.102,147 Fatty acids occur in aquatic DOM, including due to their release from biofilms as biogenic surfactants, indicating potential relevance in surface waters.148 Finally, many thiol-containing biomolecules react with RNS with rate constants ranging from 104–108 M−1 s−1 at circumneutral pH.139 In one example, glutathione reacts with ˙NO2 and N2O3 with rate constants on the order of 107 M−1 s−1, forming an S-nitrosothiol (S-nitrosoglutathione).102 Similar reactions could occur for organosulfur compounds in water treatment processes, including reactions with reduced sulfur moieties that occur in aquatic DOM.139,149

In addition to highlighting potential precursors, biochemistry research may also provide methods of differentiating the roles of various RNS. In particular, ONOOH and ONOO are potentially relevant to nitr(os)ation reactions in aquatic systems, but their reactions are relatively poorly understood, in part because these species co-occur with other RNS and because ONOOH/ONOO can react directly with organic compounds or via decomposition to ˙NO2.144 Biochemistry research has indicated that ONOOH reacts with dissolved carbon dioxide and rapidly decomposes to form ˙NO2,150 which is the primary reactant resulting in nitration of some biomolecules.141 For example, tyrosine nitration by ONOO occurs via ˙NO2, but tryptophan is susceptible to both ˙NO2 and direct nucleophilic reactions of ONOOH.139 One method to study the role of ONOOH/ONOO involves adding 3-morpholinosydnonimine (Sin-1) to experiments to continuously generate ONOO. Sin-1 decomposes in solution to generate ˙NO and superoxide, which react and result in continuous ONOO formation.151,152 Similarly, ˙NO donor molecules, such as diethylamine NONOate (DEA-NONO) have been used to generate ˙NO and N2O3 in the absence of ˙NO2.152 In combination with the use of scavengers, such as flavonoids,114 these techniques may enable a more complete understanding of RNS relevant to reactions with DOM and trace organic contaminants.

New analytical approaches used by environmental engineering researchers to characterize DOM may also enable advancements in the study of RNS. In particular, ultrahigh resolution mass spectrometry, such as Fourier transform ion cyclotron resonance tandem mass spectrometry (FT-ICR-MS/MS), has recently been used to characterize the molecular composition of DOM, and can indicate the prevalence of structures with heteroatoms such as nitrogen.153 Workflows have also been developed to link precursors and products of DOM reactions using differences in exact mass. For instance, FT-ICR-MS/MS has been used to measure the formation of oxidized products during ozonation and identify relevant precursors.154 Researchers have also employed high resolution Orbitrap mass spectrometry to identify ozonation transformation products based on known reaction mechanisms (e.g., addition of oxygen atoms, loss of alkyl groups, etc.).155 A similar approach could be used to identify nitr(os)ated products and precursors in DOM. Isotope labeling (e.g., the use of 15N-labeled reagents) can enhance this approach by enabling the identification of products in which inorganic N is incorporated into molecular structures, increasing confidence in product identification.78

Together with assessments of product toxicity, further research into RNS reactions may indicate a need for approaches to mitigate nitr(os)ated byproduct formation, for instance by reducing nitrate concentrations in drinking water sources. By taking advantage of probe compounds, RNS sources used by biochemistry researchers, and analytical techniques for DOM analysis, future studies can further our understanding of RNS reactions in surface waters and water treatment systems, and can inform approaches to minimize nitr(os)ated byproduct formation.

Conflicts of interest

There are no conflicts of interest to declare.


I would like to thank Carsten Prasse and David Sedlak for valuable comments on the draft manuscript. I would also like to acknowledge funding support from the University of British Columbia.


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