1. Introduction
Nitrophenols are well known environmental trace compounds and pollutants [
1,
2], which have been detected in various environmental matrices including air [
3,
4,
5,
6], rainwater [
3,
7,
8,
9], cloud water [
10], fog [
10,
11], snow [
1], atmospheric aerosol [
4,
5,
6,
12,
13,
14,
15,
16,
17,
18,
19,
20,
21,
22,
23,
24,
25], soils [
26,
27], and surface waters [
10,
28,
29,
30,
31,
32]. They originate from many anthropogenic and natural sources including: the incineration of wastes [
33], industrial chemical processes [
34], combustion of coal and biomass as well as vehicle and aviation fuels [
35,
36,
37], degradation of pesticides [
34,
38,
39], release of wood preservatives [
34], and atmospheric chemical reactions. 4-nitrophenol (4-NP) and 2,4-dinitrophenol (2,4-DNP), along with sugar anhydrides such as levoglucosan, serve as markers of biomass burning in ambient aerosol [
40,
41,
42,
43]. These compounds and 2-nitrophenol are recognized components of atmospheric brown carbon, i.e., a collection of light absorbing organic compounds in the atmosphere [
22,
44,
45]
Atmospheric reactions that yield nitrophenols take place both in the gas phase and in the aqueous phase. For instance, the gas-phase nitration of phenol involves hydroxyl radicals
•OH and NO
2 in the daytime or nitrate radical
•NO
3 and NO
2 in the night to respectively produce 2-nitrophenol (2-NP) or 2-NP and 4-NP. The chemical mechanisms of both processes were thoroughly reviewed [
1]. Formation of nitrophenols in atmospheric waters of all kinds is at least equally important but less understood. The possible pathways include oxidation of phenols with NO
2 and OH or NO
3 radicals [
46], electrophilic nitration initiated by N
2O
5 and ClNO
2 [
1,
47], photolytic and dark reactions involving nitrate radicals, inorganic nitrates, nitrites and nitrous acid HONO [
48,
49], and photolytic reactions with nitrogen dioxide in the presence of iron oxide and oxygen [
49].
Atmospheric sinks for nitrophenols include photolysis [
50] and the gas-phase reactions with OH radicals and NO
3 radicals [
51,
52] which are characterized by the estimated residence time of several days. More efficient sinks may include partitioning to atmospheric aqueous phases followed by reactions with various radicals and/or photolysis [
1,
46]. More recently, Barsotti, et al. [
53] demonstrated that the irradiation of aqueous solutions or viscous films containing several nitrophenols (2-NP, 4-NP, 2,4-DNP, and 2,6-DNP, i.e., 2,6-dinitrophenol) was an efficient source of HONO and NO
2− ions. Vione, et al. [
54] showed that OH radicals reacted faster than NO
3 radicals with 2-NP and 4-NP in aqueous solutions to lower the atmospheric levels of 2-NP below those of 4-NP. Hems and Abbatt [
55] studied the aqueous-phase photo-oxidation of 2,4-DNP by OH radicals, identified numerous intermediate products thereof and showed the corresponding evolution of UV absorbance of the reacting solutions. In addition, many laboratories studied the aqueous-phase reactions of other substituted phenols of atmospheric interest like guaiacol, nitro-guaiacol, vanillin, or syringol [
55,
56,
57,
58,
59].
Much of the nitrophenol chemistry has been studied for the sake of advanced oxidation processes aimed at mitigation of nitrophenols in aquatic and industrial environments [
60]. The technologies considered include: Fenton and photo-Fenton reactions based on H
2O
2 [
34,
61], TiO
2 based photocatalysis [
62,
63,
64], electrocatalysis [
65], photo-electrocatalysis [
66,
67], and wet catalysis [
68,
69]. Among the latter, a promising process was proposed which utilized reactions of nitrophenols with sulfate radical-anions generated by the cobalt-mediated decomposition of peroxymonosulfate anions [
70].
For years, some nitrophenols (2-NP, 4-NP, 2,4-DNP) have been listed as priority or hazardous pollutants [
71,
72,
73]. Generally, mono- and di-nitrophenols are considered toxic in plants and mammals [
74], while 4-nitrophenol is highly toxic in humans [
75]. Although EPA USA has not considered 4-nitrophenol carcinogenic [
38], a laboratory experiment showed the compound can destroy DNA in vitro [
76].
This work was aimed at elucidating how fast nitrophenols are removed from the atmospheric waters by reaction with sulfate radical-anions, which are important atmospheric oxidants known to react fast with numerous atmospheric pollutants [
77,
78,
79,
80,
81,
82,
83].
3. Results
In this section, we present the experimental results obtained for 4-NP. The results for other nitrophenols were similar so we present them in the
SI.
Figure 1 shows consumption of oxygen during the autoxidation of S(IV) in the presence of 4-NP. The higher was the initial concentration of nitrophenol, the slower was the consumption of O
2.
In each experiment, the autoxidation attained a quasi-stationary rate, as shown in
Figure 2 for the run with [4-NP] = 0.28 mM.
Figure 3 shows the plots of reciprocal stationary rates for autoxidation of S(IV) in the presence of 4-NP or a reference compound ethanol versus initial concentrations of each inhibitor. The plots were linear, so their slopes were used in Equation (2) to calculate the relative rate constant for the reaction of 4-NP with sulfate radical-anions
Plots for other nitrophenols, all of them linear, were placed in the SI. The results of all experiments are collected in
Table 2 and include the slopes of linear plots and the rate constants for reactions of nitrophenols with sulfate radical-anions, both observed and corrected for diffusional limitations. The uncertainties of the observed rate constants were estimated using the total differential method applied to Equation (6) with individual errors equal to the standard errors of the linear slopes and k
EtOH (
Table 2). The uncertainties of the corrected rate constants were estimated in a similar way from Equation (4), assuming arbitrarily the uncertainty of k
difffusion was 10%.
5. Conclusions
Nitrophenols (2-NP, 3-NP, 4-NP, and 2,4-NP) react fast with SO4•− radical-anions in aqueous solutions. Rate constants for these reactions, along with rate constants of several chlorophenols and phenol, correlate linearly with Brown substituent coefficients and with the relative strength of the O–H bonds in the molecules. The correlation allows estimation of rate constants for reactions of other substituted phenols with sulfate radical-anions.
The aqueous-phase reaction of 2-NP with sulfate radical-anions dominates over the aqueous-phase conversion of 2-NP by OH radicals only when SO4•‒ radicals are at least 10 times more abundant than the OH radicals. Similar domination over NO3 radical requires the concentration of sulfate radicals is at least a quarter of the concentration of nitrate radicals.
The comparison of gas-phase conversion of 2-NP by OH or NO3 radicals against the aqueous-phase conversion by sulfate radical-anions depends on the liquid water contents of a particular atmospheric system considered. In deliquescent aerosol and haze water (ω < 10−10 m3 m−3), gas-phase reactions always prevail over the aqueous-phase reactions. In cloud, rain and fog water (10−8 < ω < 10−6 m3 m−3), the aqueous-phase reaction of 2-NP dominates over the gas-phase conversion of 2-NP by hydroxyl or nitrate radicals provided the aqueous-phase concentration of sulfate radical-anions is not smaller than the aqueous-phase concentration of hydroxyl or nitrate radicals. These conclusions are based on the assumption that the gas-phase and aqueous-phase concentrations of OH, NO3, and 2-NP are bound by Henry’s equilibria.
The gas-phase and aqueous-phase conversions of other nitrophenols are expected to follow similar patterns. However, this expectation should be confirmed by calculations when constants of the gas-phase reactions of the nitrophenols with hydroxyl and nitrate radicals are available.
Last not least, we hope that the rate constants determined in the present work for atmospheric purposes may appear useful for designers of advanced oxidation processes aimed at removal of nitrophenols from various waste effluents utilizing sulfate radical-anions.