# The carbon dioxide radical anion as a reductant in water treatment and atmospheric chemistry: a review

Thomas Schaefer, Jonathan D Raff, Mallikarjuna Nadagouda, Eric Lichtfouse, Dan Meyerstein, Hartmut Herrmann, Virender Sharma

► To cite this version:

Thomas Schaefer, Jonathan D Raff, Mallikarjuna Nadagouda, Eric Lichtfouse, Dan Meyerstein, et al. The carbon dioxide radical anion as a reductant in water treatment and atmospheric chemistry: a review. Environmental Chemistry Letters, In press, 10.1007/s10311-025-01894-9, hal-05481693

HAL Id: hal-05481693

https://hal.science/hal-05481693v1

Submitted on 28 Jan 2026

HAL is a multi-disciplinary open access
archive for the deposit and dissemination of sci-
entific research documents, whether they are pub-
lished or not. The documents may come from
teaching and research institutions in France or
abroad, or from public or private research centers.

L'archive ouverte pluridisciplinaire HAL, est
destinée au dépôt et à la diffusion de documents
scientifiques de niveau recherche, publiés ou non,
émanant des établissements d'enseignement et de
recherche français ou étrangers, des laboratoires
publics ou privés.

<figure><img src="image_1.png" /></figure>
# The carbon dioxide radical anion as a reductant in water treatment and atmospheric chemistry: a review

Thomas Schaefer<fnref n="1" /> ID • Jonathan D. Raff<fnref n="2" /> ID • Mallikarjuna N. Nadagouda<fnref n="3" /> ID • Eric Lichtfouse<fnref n="4" /> ID • Dan Meyerstein<fnref n="5" /> ID • Hartmut Herrmann<fnref n="1" />,<fnref n="6" />,<fnref n="7" /> ID • Virender K. Sharma<fnref n="8" /> ID

## Abstract

The carbon dioxide radical anion, CO₂•⁻, is a highly reactive radical species involved in the reduction of the CO₂ greenhouse gas, organic synthesis, atmospheric aerosol chemistry, and treatment of halogenated compounds. In recent years, CO₂•⁻ has emerged as a strong reductant, or single electron donor. Here we present techniques used to generate CO₂•⁻ and we discuss applications to degrading pollutants such as halogenated alkanes. The potential occurrence of such reductions in water and aqueous aerosols is discussed, notably for the degradation of perfluoroalkyl substances. In the laboratory, CO₂•⁻ is directly generated by either direct electrochemical reduction of CO₂ or hydrogen atom transfer of either formate salts with and without catalysts or ferrioxalate through photochemical or radiolytic processes. The CO₂•⁻ has an ultraviolet spectrum, and CO₂•⁻ vibration modes are characterized by fast kinetics using infrared and Raman spectroscopy. The second-order rate constants of the reactions of CO₂•⁻ with halogenated alkanes, of -1.84±0.22 V, are generally slower than that of the hydrated electron, of -2.87 V, and give a negative linear relationship with energy of lower unoccupied molecular orbital, suggesting single electron transfer mechanism in reducing the halogenated compounds.

**Keywords** Reduction · Electron transfer · Kinetics · Mechanism · Degradation · PFAS · Carbon dioxide radical anion

<figure><img src="image_2.png" /></figure>

<footnote>Virender K. Sharma
vks38@miami.edu</footnote>

<footnote>Jonathan D. Raff
jdraff@iu.edu</footnote>

<footnote>Mallikarjuna N. Nadagouda
nnmalli@gmail.com</footnote>

<footnote>Eric Lichtfouse
eric.lichtfouse@icloud.com</footnote>

<footnote>Hartmut Herrmann
herrmann@tropos.de</footnote>

<footnote n="1">¹ Atmospheric Chemistry Department (ACD), Leibniz Institute for Tropospheric Research (TROPOS), Permoserstraße 15, 04318 Leipzig, Germany</footnote>

<footnote n="2">² O'Neill School of Public and Environmental Affairs, and the Department of Chemistry, Indiana University Bloomington, 702 N. Walnut Grove Ave., Bloomington, IN 47405, USA</footnote>

<footnote n="3">³ Department of Mechanical and Materials Engineering, Wright State University, Dayton, OH 45435, USA</footnote>

<footnote n="4">⁴ State Key Laboratory of Multiphase Flow in Power Engineering, Xi’an Jiaotong University, Xi’an 710049, Shaanxi, China</footnote>

<footnote n="5">⁵ Department of Chemical Sciences, Ariel University, Ariel, Israel</footnote>

<footnote n="6">⁶ School of Environmental Science and Engineering, Shandong University, Qingdao 266237, China</footnote>

<footnote n="7">⁷ Shanghai Key Laboratory of Atmospheric Particle Pollution and Prevention, Department of Environmental Science & Engineering, Institute of Atmospheric Sciences, Fudan University, Shanghai 200433, China</footnote>

<footnote n="8">⁸ Department of Chemical, Environmental, and Materials Engineering, University of Miami, 1251 Memorial Drive, Coral Gables, FL 33146, USA</footnote>
Introduction

Society issues include the continuous generation of pollutants of concerns in the environment and the increase of global warming, which require immediate attention. Among the various pollutants, halogenated compounds have a wide range of applications as pharmaceuticals, polymers, pest control agents, flame retardants, and chemical weapons [1–5]. For example, halogenated alkanes have found common use and even crucial synthetic intermediates of industrial chemicals [6]. Perfluoroalkyl substances (PFAS) are other class of halogenated compound that have become of great concerns in the past decade [7–9]. Many approaches are being examined to decrease the level of halogenated compounds by applying reactive radical species [10–12]. The decrease in global warming is being pursued through carbon dioxide sequestration and converting global warming gases to value added products through involvement of radicals [13–15].

The production of radical species in environment, biological, catalytic, electrochemical, and industrial processes has been known for several decades [16–22]. Since inorganic anions like bicarbonate/carbonate ($HCO_3^- / CO_3^{2-}$) are ubiquitous in most environmental and biological systems, the generation of highly reactive species, either oxidizing or reducing, may likely occur through interactions of radicals with these anions. For example, Fenton and Fenton-like reactions in the presence of $HCO_3^- / CO_3^{2-}$ result in carbonate radical anion ($CO_3^{•-}$), an oxidant [23, 24]. The oxidation reaction of simple organic molecule, formic acid or formate with hydroxyl radical ($HO^{•}$) gives the formation of the carbon dioxide radical anions ($COOH / CO_2^{•-}$), reducingants [25]. There have been several publications in the literature for implicating the different types of radical species in advanced oxidation/reduction processes (AOPs/ARPs) in water treatments, and enzymatic reactions in the human body [26]. The involvement of $HO^{•}$ and $CO_3^{•-}$ in oxidation of a wide range of pollutants and diseases has been widely reported [27–32].

The association of $CO_2^{•-}$ with diseases has been suggested. Examples include the generation of $CO_2^{•-}$ in metabolism of carbon tetrachloride in rat microsomes and in sodium formate poisoning [33]. In the last few decades, the role of $CO_2^{•-}$ in synthesizing organic molecules has been forthcoming [29]. In organic chemistry, $CO_2^{•-}$ acts as a strong single electron reductant to ultimately yield carboxylic acids through Giese radical addition to the unsaturated C–C bonds. Examples include the hydro-carboxylation of activated alkenes (e.g., acrylamides and styrenes) [29]. An elegant review on this topic has been recently written [30]. $CO_2^{•-}$ may be involved as an important intermediate in conversion of $CO_2$ into value added products [16].

The participation of $CO_2^{•-}$ in environmental processes has not been described in-depth. This is surprising considering the literature known on the role of $CO_2^{•-}$ in organic synthesis [28, 30] clearly showing the high potential of $CO_2^{•-}$ in transformation of organic pollutants in the environment [30]. We have therefore aim to write this review giving its generation methods and properties and then summarizing what is currently known in its role in abatements of pollutants in water. One of the main objectives of the review is to encourage young researchers to consider $CO_2^{•-}$ as one of the sources to carry out advanced reduction processes in cleaning polluted water, especially organic molecules like halogenated hydrocarbons and per- and polyfluoroalkyl substances (PFAS) that require preferentially reduction strategies. The researchers will also be benefited by performing research to elucidate the role of reducing species like $CO_2^{•-}$ in aqueous atmospheric chemistry, particularly under high ionic strength aerosols particles. This review presents generation and properties of $CO_2^{•-}$, followed by its applications in treatment of halogenated alkanes and PFAS. The final section presents current status of $CO_2^{•-}$ in liquid atmospheric phase.

Generation of the carbon dioxide radical anion

The transformation of $CO_2$, a highly stable linear molecule, to the $CO_2^{•-}$, a bent molecule, is challenging because this requires change in geometry and orbital hybridization, which demands high activation energy. Different approaches have been applied to generate $CO_2^{•-}$, which included chemical, catalytic, electrochemically, and pulse radiolysis processes, [30, 34] which utilized either hydrogen-atom transfer (HAT) or single-electron transfer (SET) step (Fig. 1). In the hydrogen-atom transfer, a hydrogen atom is transferred from one molecule to another [35].

In the chemical methods, formate salt is relatively low cost and available and about 1 million tons/y of formic acid has been manufactured [36]. Researchers have used hydrogen-atom transfer reagents with and without light irradiation to convert the salts to $CO_2^{•-}$. The $CO_2$ was converted by applying UV and visible irradiation with organics or catalysts [37, 38]. Organic radicals were produced using irradiation, which assisted in reducing $CO_2$ to $CO_2^{•-}$. The use of visible light to form $CO_2^{•-}$ from $CO_2$ has emerged as a favorable area of investigations [37]. Similarly, researchers are using electron in electrochemical methods to yield $CO_2^{•-}$ [34].

The pulse radiolysis/steady-state radiolysis method has also been applied to generate $CO_2^{•-}$ that helps to understand its nature in water and organic solvent. This technique also
<figure><img src="image_3.png" /><figcaption>**Fig. 1** Generation of CO₂•⁻ from CO₂ or formate salts. Generally, there are four types of methods applied to generate CO₂•⁻ from precursors, CO₂ and formate salt, which include electrochemical, radiolysis, catalytic, and chemical methods. The single electron, produced in these methods, usually able to yields CO₂•⁻. Formate salts involved hydrogen atom transfer process to give CO₂•⁻.</figcaption></figure>

<figure><img src="image_4.png" /><figcaption>**Fig. 2** Absorbance spectrum of HOOCO• /CO₂•⁻ in irradiated aqueous solution. The CO₂•⁻ has an absorbance in the ultraviolet wavelength range and has an absorption maximum at 235 nm. Two researchers reported different molar absorptivity in the lower wavelength range, less than 290 nm. The protonation of CO₂•⁻ results in lower molar absorptivity</figcaption></figure>

<table><caption>**Table 1** pH-dependent CO₂•⁻ acid-base equilibrium as a function of pKₐ</caption><thead><tr><th rowspan="2">pH value</th><th colspan="2">pK<sub>a</sub>=2.3 [47]</th><th colspan="2">pK<sub>a</sub>=3.9 [50]</th></tr><tr><th>HOCO• (%)</th><th>CO<sub>2</sub>•⁻ (%)</th><th>HOCO• (%)</th><th>CO<sub>2</sub>•⁻ (%)</th></tr></thead><tbody><tr><td>0</td><td>99.5</td><td>0.5</td><td>100.0</td><td>0.0</td></tr><tr><td>1</td><td>95.2</td><td>4.8</td><td>99.9</td><td>0.1</td></tr><tr><td>2</td><td>66.6</td><td>33.4</td><td>98.8</td><td>1.2</td></tr><tr><td>3</td><td>16.6</td><td>83.4</td><td>88.8</td><td>11.2</td></tr><tr><td>4</td><td>2.0</td><td>98.0</td><td>44.3</td><td>55.7</td></tr><tr><td>5</td><td>0.2</td><td>99.8</td><td>7.4</td><td>92.6</td></tr><tr><td>6</td><td>0.0</td><td>100.0</td><td>0.8</td><td>99.2</td></tr><tr><td>7</td><td>0.0</td><td>100.0</td><td>0.1</td><td>99.9</td></tr><tr><td>8</td><td>0.0</td><td>100.0</td><td>0.0</td><td>100.0</td></tr></tbody></table>

used to characterize CO₂•⁻ spectroscopically. An ionizing radiation of water in aqueous solution using either electron beams or isotope gamma sources reducing radicals (e.g., hydrate electron (e⁻_aq)) or oxidizing radicals (e.g., HO•) radical in water [25]. The e⁻_aq can attach to CO₂ to yield CO₂•⁻ (reaction 1) [39, 40].

$$CO_{2(aq)} + e_{aq}^{-} \rightarrow CO_{2}^{\bullet-} \quad (1)$$

The CO₂•⁻ is a strong reducing species with a standard redox potential of—1.90 V vs. NHE [41]. This value may also be dependent on the solvent used in carrying out the measurements. Comparatively, the standard redox potential of e⁻_aq is -2.88 V vs. NHE [41]. Therefore, the reduction of CO₂ by e⁻_aq is possible, which could be observed in 7 ps in the picosecond (ps) pulse radiolysis technique [42]. Rate constants of 7.7 - 9.0 × 10⁹ M⁻¹ s⁻¹ have been determined [42, 43]. The value of the rate constant for the reduction of CO₂ by e⁻_aq in acetonitrile solvent has been reported to be 3.2 × 10¹⁰ M⁻¹ s⁻¹ to form CO₂•⁻ [44].

The pulse radiolysis technique enabled to collect UV spectra of CO₂•⁻ (Fig. 2). Data were taken from Neta et al., [45] and Buxton and Sellers, [46]. The CO₂•⁻ has an absorption in the UV range with a maximum at 235 nm (ε=3.0 × 10³ M⁻¹ cm⁻¹). It is thus not convenient to perform kinetic measurement at this UV wavelength. The experimental studies and ab initio molecular dynamics (AIMD) simulations showed that only small energy is needed to convert e⁻_aq...CO₂ complex to give CO₂•⁻ [39].

The characteristic of the spectrum of CO₂•⁻ is presented in Fig. 2. The spectral measurements, determined in the pH range from 3.0 – 9.0. However, two researchers obtained different molar absorptivities of CO₂•⁻ [28, 29]. The spectrum of protonated species of CO₂•⁻ (HOCO•) could also be obtained (Fig. 2).

The protonation of CO₂•⁻ occurs in strong acidic solution (reaction 2) [46–49].

$$HOCO\bullet \rightleftharpoons H^+ + CO_2^{\bullet-} \quad pK_a = 1.4 - 3.9 \quad (2)$$

Several researchers investigated the acid-base equilibrium of the CO₂•⁻ and a wide range of value of pKₐ have been reported, with quite large effects on the acid-base balance at certain pH values [49]. Flyunt et al. reported that the pKₐ value of 1.4 specified by Buxton and Sellers is an artifact [46, 47]. The study by Flyunt et al. shows a pKₐ value of 2.3, which is also in the lower range of pKₐ values. Table 1 summarizes reported values of pKₐ in the literature and shows their impact.

A nice discussion on this variation has been done pointing out the possible assumptions made in applying different techniques, e.g., conductivity, electronic absorption, and
spectroscopic and electron paramagnetic resonance (EPR) to evaluate this value. One of the main assumptions were on the formal charge on the carbon or oxygen of the CO₂•⁻. A recent approach of using Raman spectroscopic techniques attempted to resolve this variation by monitoring the Raman band intensities of aqueous CO₂•⁻ at different pHs. The band intensity rapidly decreased below pH 4.0 and the value of pKₐ as 3.4±0.2 was determined [49]. This value is reasonable because similarity of pKₐ between the parent molecules of monocarboxylic acids and their alpha carbon centered radicals. The pKₐ value of formic acid is 3.745, which is close to 3.4, determined experimentally [49, 51].

The formic acid/formate ion can react with HO• to give CO₂•⁻ (reactions 3 and 4) [25].

$$ \mathrm{HCOOH} + \mathrm{HO}^{\bullet} \rightarrow \mathrm{HOCO}^{\bullet} + \mathrm{H}_{2}\mathrm{O} \quad k = 1.4 \times 10^{8} \mathrm{M}^{-1} \mathrm{s}^{-1} \quad (3) $$

$$ \mathrm{HCOO}^{-} + \mathrm{HO}^{\bullet} \rightarrow \mathrm{CO}_{2}^{\bullet-} + \mathrm{H}_{2}\mathrm{O} \quad k = 2.6 \times 10^{9} \mathrm{M}^{-1} \mathrm{s}^{-1} \quad (4) $$

The CO₂•⁻ may go through a recombination reaction (5) and the lifetime would be in the time scale of microsecond, depending on the experimental conditions [46, 47].

The recombination of CO₂•⁻ in neutral and alkaline conditions leads in oxalic acid (reaction 5). In acidic solutions, when the CO₂•⁻ radical is present in its protonated form (HOCO•), the cross-termination between HOCO• and CO₂•⁻ yields CO₂ and formate (reaction 6), similar to the disproportionation reaction of the two HOCO• radicals, which leads to CO₂ and formic acid (reaction 7) [25, 47].

$$ \mathrm{CO}_{2}^{\bullet-} + \mathrm{CO}_{2}^{\bullet-} \rightarrow \mathrm{C}_{2}\mathrm{O}_{4}^{2-} \quad (5) $$

$$ \mathrm{HOCO}^{\bullet} + \mathrm{CO}_{2}^{\bullet-} \rightarrow \mathrm{CO}_{2} + \mathrm{HCO}_{2}^{-} \quad (\text{in acidic solutions}) \quad (6) $$

$$ \mathrm{HOCO}^{\bullet} + \mathrm{HOCO}^{\bullet} \rightarrow \mathrm{CO}_{2} + \mathrm{HCO}_{2}\mathrm{H} \quad (\text{in acidic solutions}) \quad (7) $$

The rate constant of the self-termination of the CO₂•⁻ is considered as pH independent [46, 52]. However, the rate

constant was reported to be in a range of k=3.8 × 10⁸ M⁻¹ s⁻¹ to k=7.5 × 10⁸ M⁻¹ s⁻¹ [45–47].

In addition to the self-termination reaction, the CO₂•⁻ undergoes electron transfer by reducing O₂, transition metal ions or aromatic systems. However, it is also possible for CO₂•⁻ to oxidize through an addition reaction at a double bond, or with more complex molecules such as enzymes [53, 54].

However, a recent study demonstrated the stabilization of CO₂•⁻ on gold and copper nanoparticles surfaces (CO₂)ₐₛₙ, [42] which has a life time of three-orders of magnitude higher than in homogenous solution of the CO₂•⁻. This suggests that (CO₂)ₐₛₙ may be fundamentally different from CO₂•⁻ₐq.

The photochemistry of ferric-oxalic acid system (or iron(III)-oxalto complex) has also been applied in many studies to demonstrate the formation of CO₂•⁻ [55]. Recently, a detailed mechanism on the formation of CO₂•⁻ has been examined using time-resolved spectroscopic technique involving infrared (IR) spectroscopy method, which is known for the sensitivity towards bonding types and functional groups of the molecules [56]. The main steps of the mechanism are represented by the Fig. 3. The mechanism of Fe(III)-oxalato complex involves intramolecular electron transfer in sub-picosecond from oxalate to iron that yields iron(II) complexed in which no change in two oxalate ligands and the third oxalate ligand is oxidized. Interestingly, the oxidized oxalate ligand goes through electron transfer within 40 ps to give free solvated CO₂(aq) and a CO₂•⁻, based on spectroscopic measurements and theoretical calculations [56]. The molecular stimulation also supported the steps given in Fig. 3. It is possible that the similar complexes of organic acids such as malonate, lactate, and pyruvate may undergo through photolysis to generate CO₂•⁻.

<figure><img src="image_5.png" /><figcaption>**Fig. 3** Key species and configurations in the photolysis of ferrioxalate up to 3 ns determined from transient mid-infrared data and molecular simulation. **A)** One photo-excited oxalate ligand reduces iron(III) within 0.1 ps and the resulting oxalate radical anion, C₂O₄•⁻, dissociates rapidly, likely on the time scale of a single C–C vibration. **B)** CO₂ leaves the iron(II) coordination sphere within ~100 ps, while the radical anion, CO₂•⁻, is predicted to remain in the iron(II) coordination sphere for at least 10 ns</figcaption></figure>
<table><caption>**Table 2** Properties of CO₂ compared to CO₂•⁻ [18]</caption><thead><tr><th>Properties</th><th>Carbon dioxide</th><th>Carbon dioxide<br>radical anion</th></tr></thead><tbody><tr><td>Bond angle (OCO)</td><td>180°</td><td>127±8°</td></tr><tr><td>Bond length (CO)</td><td>1.16 Å</td><td>1.25 Å</td></tr><tr><td>Proton affinity</td><td>540.5 kJ mol<sup>-1</sup></td><td>770–780 kJ mol<sup>-1</sup></td></tr><tr><td>Quadrupole moment</td><td>~4.3 DÅ</td><td>~6.5–7.5 DÅ</td></tr><tr><td>Dipole moment</td><td>0 D</td><td>~0.27 D</td></tr><tr><td>Symmetry group</td><td>D<sub>h</sub></td><td>C<sub>2v</sub></td></tr></tbody></table>

<figure><img src="image_6.png" /><figcaption>**Fig. 4 A)** Infrared spectra of hydrated CO₂•⁻ (H₂O)ₙ (n=61). The dashed line represents the IR spectrum of liquid water, scaled to the maximum intensity of the n=61 spectrum. (Adapted from Herburger et al. [57]) **B)** Resonance Raman spectrum of CO₂•⁻ probed at 235 nm in water. (Adapted from Janik and Tripathi [49])</figcaption></figure>

## Physicochemical properties

The formation of CO₂•⁻ results in significant change in physicochemical properties from the CO₂ (Table 2). The 180° bond angle of O-C-O bents to 127±8° in the CO₂•⁻. The asymmetric stretch in CO₂•⁻ is seen in the infrared spectroscopic measurements, which has a band at 1650 cm⁻¹. The C-O bond lengths of 1.17 Å in CO₂ increases to 1.25 Å in CO₂•⁻. The metastability of CO₂•⁻ can be seen in the electron affinity of CO₂, which is negative. In the CO₂•⁻, the C-O bonds are weakened, and the carbon center acquires a

radical character. This causes variation in physicochemical properties from CO₂ to CO₂•⁻ (Table 2).

Initial spectroscopic characterization of CO₂•⁻ was carried out by stabilizing the radical into solid alkali halides matrices, generated by radiating the solid-formate solution with gamma rays [48]. The infrared spectra were determined in matrices (KCl, KBr, and NaBr) that had a band at 1671 cm⁻¹ corresponding to the C-O antisymmetric stretching of the radical anion. This band disappeared when solid samples were dissolved into water. The antisymmetric CO-bond stretching infrared band in the carbon-12 and -13 isotopic frequencies were used to calculate a bond angle of 127°±8°. In later years, Fourier transform-ion cyclotron resonance (FTICR) mass spectrometer and pulse radiolysis-IR detector were applied to obtain spectroscopic information on the hydrated CO₂•⁻(H₂O)ₙ. Figure 4A shows the infrared spectra of CO₂•⁻(H₂O)ₙ at T=80 K where n=61.

The asymmetric band blue shifted in water, which was more with an increase in cluster size (e.g., 1650 cm⁻¹ for n=61). This band is associated with frequencies of normal modes of CO₂•⁻ and various H₂O molecules oscillating. In Fig. 4A, the spectrum at 1150–1350 cm⁻¹ represents the symmetric CO₂•⁻ stretching mode in which band position depends on the hydrogen-bond network surrounding the CO₂•⁻. The symmetric band of the radical could be reproduced using quantum calculations, [57] which suggested the hydrogen-bond network around CO₂•⁻.

Additional information on the vibrational modes was obtained by applying pulse radiolysis time-resolved resonance Raman spectroscopy [49]. The resonance Raman spectrum of CO₂•⁻ given in Fig. 4B, [49] consists a stronger vibration at 1298 cm⁻¹ and a relatively weak vibration at 742 cm⁻¹, which were assigned to OCO bending mode. The combination of 742 cm⁻¹ and 1298 cm⁻¹ vibrational frequencies yields a band at 2029 cm⁻¹. The Raman band at 1298 cm⁻¹ corresponds to the symmetric CO stretch. In the inert matrices, this stretch of the CO₂•⁻ was 45 cm⁻¹ lower than seen in water. Interestingly, detachment of the electron from the moiety of CO₂ in the direction of solvation shell of the hydrated CO₂•⁻ was suggested based on almost absence of overtones for the stretch of the symmetric CO.

The spectral characterization of CO₂•⁻ using electron paramagnetic resonance (EPR) is restricted to the solid matrices [58]. For example, the values of g, line width, and ¹³C splitting for the CO₂•⁻ were determined as 2.0006, 28 G, and 135 G, respectively. A complex spectral pattern of EPR was seen due to two nearly equivalent ¹⁷O nuclei. The ¹⁷O has a nuclear spin 5/2, the principal directions of the g matrix for each ¹⁷O nuclei and the signal was additionally splitted into 2I+1=6 lines for the C¹⁷O¹⁶O isotopomer or into 2nI+1=11 lines for the case of C¹⁷O¹⁷O species [58] (Fig. 5).
<figure><img src="image_7.png" /><figcaption>**Fig. 5** Experimental (a) and simulated (b) electron paramagnetic resonance spectra of C¹⁷O₂•⁻ For the sake of clarity only the Aₓₓ component of species a is shown by the stick diagram. The spectra were taken at 77 K. Adapted from with the permission of Wiley Inc. [59]</figcaption></figure>

<figure><img src="image_8.png" /><figcaption>**Fig. 6** Electron paramagnetic resonance spectrum of the carbon dioxide radical (CO₂•⁻) DMPO adduct adapted Villamena et al.[60] Measured by the UV photolysis of 100 µM H₂O₂ in the presence of the CO₂•⁻ precursor 0.1 M sodium formate (NaHCO₂) and 0.05 M 5,5-dimethyl-1-pyrroline N-oxide (DMPO) at pH 6.8 of the aqueous solution</figcaption></figure>

In aqueous solution, the identification of CO₂•⁻ was carried out by using spin trapping technique by adding an agent like 5,5-Dimethyl-1-Pyrroline N-Oxide (DMPO) [60], a commonly used spin-trap in establishing the involvement of radicals in aqueous reactions. The CO₂•⁻ forms carbon-centered adduct, evidenced by experimental and theoretical calculations. The spin population on the CO₂•⁻ controls the mode of the radical [60].

The electron paramagnetic resonance spectrum of the adduct given in Fig. 6 has 99% signal of the sextet consisting of a$_{N}$ and a$_{β-H}$ as 15.74 and 18.74 G, respectively. It should be pointed out these features are different from

<figure><img src="image_9.png" /><figcaption>**Fig. 7** Relationship of the second-order rate constants (k, M⁻¹ s⁻¹) for CO₂•⁻ and e⁻_aq with the energy of lowest unoccupied molecular orbital (E_LUMO) of chlorinated alkanes. The negative values of E_LUMO have high reactivity of chlorinated alkanes. The higher reactivity of e⁻_aq with chlorinated alkanes than that of CO₂•⁻</figcaption></figure>

the adduct formed by HO•, a commonly interfering radical, with DMPO (i.e., DMPO-¹⁶OH or DMPO-¹⁷OH adducts) [60].

## Water treatment

### Reactivity of CO₂•⁻

### Halogenated compounds

The kinetics of the reactions of CO₂•⁻ with compounds of environmental relevant is limited compared to other reductants like e⁻_aq. The focus of the studied reactions with CO₂•⁻ has been on the organic synthesis utilizing the reactivity of intermediate CO₂•⁻ species.

A recent work on the kinetics of the reactions of CO₂•⁻ with halogenated alkanes have determined the second-order rate constants, k(CO₂•⁻) [50]. The decay of CO₂•⁻ at 340 nm was followed under pseudo-order conditions to obtain k(CO₂•⁻). The values of k(CO₂•⁻) were in the range of 5.8 × 10⁶ M⁻¹ s⁻¹ (for CH₂Cl₂) to 2.5 × 10⁸ M⁻¹ s⁻¹ (for CCl₄). The rate constants increase with the substitution of chlorine atom into alkanes, which is typical for the reductant species. A linear negative relationship between log k(CO₂•⁻) and E_LUMO (energy of lowest unoccupied molecular orbital) of these chlorinated alkanes was found (Fig. 7 and Table 3) [50, 54]. Comparatively, the second-order rate constants for the reactivity of e⁻_aq with halogenated alkanes (k(e⁻_aq)) also had a negative relationship (Fig. 7), however, values of k(e⁻_aq) are two-order of magnitude higher than that of k(CO₂•⁻). Both relationships could be expressed as:
<table><caption>**Table 3** Second-order rate constants ($k$, M⁻¹ s⁻¹) and the energy of lowest unoccupied molecular orbital ($E_{LUMO}$, eV) for CO₂•⁻ and e⁻_aq of chlorinated alkanes [50, 54]</caption><thead><tr><th>Compound</th><th>E<sub>LUMO</sub><sup>*</sup><br>/eV</th><th>Rate constants/k<sub>2nd</sub>, M<sup>-1</sup> s<sup>-1</sup></th></tr></thead><tbody><tr><td>CO<sub>2</sub>•<sup>-</sup></td><td></td><td></td></tr><tr><td>CCl<sub>4</sub></td><td>-0.231</td><td>2.5 × 10<sup>8</sup></td></tr><tr><td>CHCl<sub>3</sub></td><td>-4.32 × 10<sup>-3</sup></td><td>6.2 × 10<sup>7</sup></td></tr><tr><td>CH<sub>2</sub>Cl<sub>2</sub></td><td>1.91 × 10<sup>-2</sup></td><td>5.8 × 10<sup>6</sup></td></tr><tr><td>e<sup>-</sup><sub>aq</sub></td><td></td><td></td></tr><tr><td>CCl<sub>4</sub></td><td>-0.231</td><td>1.87 × 10<sup>10</sup> §</td></tr><tr><td>CHCl<sub>3</sub></td><td>-4.32 × 10<sup>-3</sup></td><td>3 × 10<sup>10</sup> #</td></tr><tr><td>CH<sub>2</sub>Cl<sub>2</sub></td><td>1.91 × 10<sup>-2</sup></td><td>6 × 10<sup>9</sup> #</td></tr><tr><td>CH<sub>3</sub>Cl</td><td>4.88 × 10<sup>-2</sup></td><td>8.06 × 10<sup>8</sup> §</td></tr><tr><td>CH<sub>3</sub>CH<sub>2</sub>Cl</td><td>4.56 × 10<sup>-2</sup></td><td>7 × 10<sup>8</sup> #</td></tr><tr><td>CH<sub>2</sub>ClCH<sub>2</sub>Cl</td><td>1.94 × 10<sup>-2</sup></td><td>1.77 × 10<sup>9</sup> §</td></tr><tr><td>CH<sub>3</sub>CHCl<sub>2</sub></td><td>1.93 × 10<sup>-2</sup></td><td>9 × 10<sup>9</sup> #</td></tr><tr><td>CH<sub>2</sub>ClCHCl<sub>2</sub></td><td>6.82 × 10<sup>-3</sup></td><td>8.4 × 10<sup>9</sup> #</td></tr><tr><td>CH<sub>3</sub>CCl<sub>3</sub></td><td>-2.65 × 10<sup>-3</sup></td><td>1.95 × 10<sup>10</sup> §</td></tr><tr><td>C<sub>2</sub>Cl<sub>6</sub></td><td>-2.07 × 10<sup>-2</sup></td><td>3.9 × 10<sup>10</sup> #</td></tr><tr><td>CH<sub>2</sub>ClCH<sub>2</sub>CH<sub>3</sub></td><td>4.54 × 10<sup>-2</sup></td><td>6.67 × 10<sup>8</sup> §</td></tr><tr><td>CH<sub>2</sub>ClCH<sub>2</sub>CH<sub>2</sub>CH<sub>3</sub></td><td>4.53 × 10<sup>-2</sup></td><td>4.45 × 10<sup>8</sup> §</td></tr><tr><td>CH<sub>2</sub>Br<sub>2</sub></td><td>-2.58 × 10<sup>-2</sup></td><td>2 × 10<sup>10</sup> #</td></tr><tr><td>CH<sub>3</sub>CH<sub>2</sub>Br</td><td>-5.68 × 10<sup>-3</sup></td><td>1.07 × 10<sup>10</sup> §</td></tr><tr><td>CH<sub>3</sub>CH<sub>2</sub>CH<sub>2</sub>Br</td><td>-5.60 × 10<sup>-3</sup></td><td>9.67 × 10<sup>9</sup> §</td></tr><tr><td>CH<sub>3</sub>I</td><td>-1.58 × 10<sup>-2</sup></td><td>1.6 × 10<sup>10</sup> §</td></tr><tr><td>CH<sub>2</sub>I<sub>2</sub></td><td>-3.26 × 10<sup>-2</sup></td><td>3.4 × 10<sup>10</sup> §</td></tr><tr><td>C<sub>2</sub>H<sub>5</sub>I</td><td>-1.64 × 10<sup>-2</sup></td><td>1.5 × 10<sup>10</sup> §</td></tr></tbody></table>

$$ \log k (\text{CO}_2^\bullet^-) = (7.54 \pm 0.06) - (3.89 \pm 0.33) \times 10^1 \quad r^2 = 0.99 \quad (8) $$

$$ \log k (e_{aq}^-) = (1.01 \pm 0.09) \times 10^1 - (2.58 \pm 0.29) \times 10^1 \quad r^2 = 0.89 \quad (9) $$

These relationships, especially with the CO₂•⁻, may be utilized to estimate the values of k(CO₂•⁻) for other halogenated alkanes. For example, k(CO₂•⁻) was calculated using Eq. (6) as 4.4 × 10⁵ M⁻¹ s⁻¹.[50] Importantly, the estimations of k(CO₂•⁻) for PFASs, another class of halogenated alkanes of emerging interest, may be estimated for their reactivity with CO₂•⁻ for which scant information is available. To check this possibility, we have determined E<sub>LUMO</sub> for, for which the rate constants have been determined independently in literature [61, 62].

The relationship shown in Fig. 7, may be used to learn the mechanism of the reactivity of CO₂•⁻ with halogenated alkanes. Halogenated alkanes with low value of E<sub>LUMO</sub> suggests easy acceptance of electrons by these molecules. The lower values of E<sub>LUMO</sub> means high effectiveness of electron acceptance. The rate-determining step in a single electron transfer (SET) mechanism is the transfer of an electron from a donor to an acceptor. A decrease in energy of LUMO suggests a more energetically favorable electron transfer that

results in increasing rate or higher rate constants. The negative trend of k(CO₂•⁻) with E<sub>LUMO</sub> (i.e., low E<sub>LUMO</sub> molecules has high k(CO₂•⁻)) indicates that the CO₂•⁻ may be attacking these compounds by direct-electron transfer. The slopes of the two linear relationship (Eq. (6) and (7)) differ with higher slope with e⁻_aq than that with CO₂•⁻. It is thus possible that the electron-transfer to halogenated electrons by e⁻_aq may be somewhat easier than by CO₂•⁻. If this holds also for the PFASs, future kinetics investigations of these reactions are needed. It is interesting to note that dehalogenation by CO₂•⁻ formed from formate can cause a chain reaction (reaction 10 and 11) [63].

$$ \mathrm{RX} + \mathrm{CO}_2^\bullet^- \rightarrow \mathrm{R}^\bullet + \mathrm{CO}_2 + \mathrm{X}^- \quad (10) $$

$$ \mathrm{R}^\bullet + \mathrm{HCO}_2^- \rightarrow \mathrm{RH} + \mathrm{CO}_2^\bullet^- \quad (11) $$

## Perfluoroalkyl substances (PFAS)

The scant information on the kinetics of the reaction between CO₂•⁻ and PFAS is available. The known rate constant using a pulse radiolysis technique is for the perfluorobutanesulfonate (PFBS, C₄F₉SO₃⁻), which is 4.8 × 10⁷ M⁻¹ s⁻¹.[61] Interestingly, this rate constant is higher than that of e⁻_aq with PFBS (k<1.0 × 10⁷ M⁻¹ s⁻¹). Products of the degradation of PFBS by CO₂•⁻ and e⁻_aq were characterized by ¹⁹F nuclear magnetic resonance and high-performance liquid chromatography-mass spectrometer techniques. Products suggest defluorination of PFBS takes place, however, the cleavage of C-F was rather difficult in the case of degradation by CO₂•⁻ than that of e⁻_aq. Results suggest the mechanisms of the degradation by CO₂•⁻ is different from the degradation mechanism by e⁻_aq. When e⁻_aq reacts with PFBS, e⁻_aq first produces a radical anion (PFBS + e⁻_aq → PFBS•⁻) that subsequently leads to defluorination (PFBS → product(s)+F⁻) [61]. However, CO₂•⁻ may be reacting through a complex mechanism than does not easily facilitate the defluorination of PFBS. This warrants systematic mechanism investigation of PFBS and other PFAS by CO₂•⁻.

## Degradation of organic compounds

The CO₂•⁻ as a reductant species performs both single electron transfer and carboxylation steps, particularly under photochemical processes and has been exploited in synthesis of organic molecules [29]. Comparatively, scant information on CO₂•⁻ as a reductant to reduce organic contaminants is available in literature. The studies on this advanced reduction process (ARP) are mostly limited to degradation of aliphatic halogenated alkanes.

<footnote>*Calculated using Gaussian semi-empirical method PM3</footnote>

<footnote>§Values and #averaged values from NDRL/NIST Solution Kinetics Database Resources</footnote>
Chlorinated alkanes

In the initial study, the degradation of carbon tetrachloride (CCl₄) by UV-peroxydisulfate (PDS)-formic acid system at pH 6.0–8.0 was sought [64, 65]. The generated CO₂•- in the system degrades CCl₄. The process is influenced by the concentrations of PDS and formic acid. The complete degradation of CCl₄ at concentration of 1.5 × 10⁻⁴ M could be achieved within 20 min at high concentrations of PDS (e.g., [PDS]=2.25 × 10⁻³ M and formic acid=2.25 × 10⁻³ M or [PDS]=1.50 × 10⁻³ M and formic acid=3.55 × 10⁻³ M, and photon flux of 2.1 × 10⁻⁵ Einstein cm⁻² s⁻¹). However, no mineralization of CCl₄ to Cl⁻ was seen [66]. The solution pH had a minimal effect on the degradation of CCl₄. The maximum dichlorination of 1.5 × 10⁻⁴ M was 80% in 30 min when the concentrations were [PDS]=1.50 × 10⁻³ M and formic acid=3.55 × 10⁻³ M, and photon flux of 2.1 × 10⁻⁵ Einstein cm⁻² s⁻¹. Anions (Cl⁻, NO₃⁻, and HCO₃⁻) decreased the degradation rate of CCl₄ by UV-PDS-formic acid system [64]. The decreasing performance may be related to potential reactions of these anions with the reactive species, CO₂•-, CH₃Cl₃-n•, (n=0–3), and SO₄²⁻ in the system.

A later study was conducted on the degradation of chloroform (CHCl₃) by the UV-formic acid, UV-acetic acid, and UV-propionic acid at initial pH 2.5 and UV lamp (λ=254 nm, P=5 W) [67]. The CO₂•- was proposed as the reactive species to degrade CHCl₃, the efficiency of which decreased as formic acid>acetic acid>propionic acid and hence the degradation rates. At neutral pH and alkaline medium, the degradation of CHCl₃ decreased in the UV-formic acid system. In the UV-PDS-formic acid, UV-PDS-citric acid, and UV-PDS-oxalic acid system, the degradation of trichloroacetic acid (2.0 × 10⁻⁴ M) was investigated at pH 4.5 [68]. A 20 W mercury lamp at 254 nm was used. The degradation was highest in the UV-PDS-formic acid, similar to another study [67], i.e., increase in chain length of the acid used decreased the performance of degrading CHCl₃. Deoxygenated water had enhanced effect on the degradation of trichloroacetic acid. The samples with oxygen could compete with the reactive species, CO₂•- (k~3.0 × 10⁹ M⁻¹ s⁻¹) [25] to decrease the rate of the degradation. A significant degradation of trichloroacetic acid of 50% was observed even in the presence of 1.0 × 10⁻³ M. Samples of real water matrices also showed 70% degradation of trichloroacetic acid in 120 min. Importantly, degradation of trichloroacetic acid by the UV-PDS-formic acid system gave dichloroacetic acid and monochloroacetic acid as the products, which are considered to be less toxic than the parent molecule, trichloroacetic acid.

A detailed study on the degradation of halogenated alkanes was conducted to understand the mechanism of the reduction process by UV₂₅₄ nm·H₂O₂-formic acid system

[50]. The complete degradation of CCl₄ was observed in 10 min while ~70% degradation of CHCl₃ was found in the same time period ([CCl₄]₀=[CHCl₃]=1.0 × 10⁻⁵ M, [Formic acid]₀=1.0 × 10⁻² M, [H₂O₂]₀=5.0 × 10⁻³ M, pH 7.0, and intensity=1.75 × 10⁻³ W cm⁻²). This is consistent with the rate constants trend of the reactivity of CO₂•- with halogenated alkanes, which decrease with decrease in number of chlorine atoms in the molecules (see Fig. 7). Experiments on degradation of CHCl₃ with nitrogen saturated, air saturated, oxygen saturated conditions were also performed. Nitrogen saturated solution has higher degradation than oxygen saturated solution, which was expected because oxygen scavenges the reactive species, CO₂•-, to reduce CHCl₃. The methyl viologen (MV²⁺) in the experiments also decrease the rate of degradation of CHCl₃ due to scavenging effect through its reactivity with CO₂•- (k=6.4 × 10⁹ M⁻¹ s⁻¹) [50].

The reducing products of the degradation of CHCl₃ were identified using a high-performance liquid chromatography technique. The major product was the release of Cl⁻ from the parent CHCl₃. The minor products were dichloroacetic acid (CHCl₂CO₂⁻) and tetrachloroethane (C₂H₂Cl₄). The sum of the identified products accounted for 75% of Cl loss from the CHCl₃.[50] Formation of these products may be described by reactions (12, 13, 14).

$$ \mathrm{CHCl_3} + \mathrm{CO_2}^{\bullet-} \rightarrow \mathrm{CHCl_2}^{\bullet-} + \mathrm{Cl}^{-} + \mathrm{CO_2} \quad (12) $$

$$ \mathrm{CHCl_2}^{\bullet-} + \mathrm{CO_2}^{\bullet-} \rightarrow \mathrm{CHCl_2COO}^{-} \quad (13) $$

$$ \mathrm{CHCl_2}^{\bullet-} + \mathrm{CHCl_2}^{\bullet-} \rightarrow \mathrm{C_2H_2Cl_4} \quad (14) $$

It is possible that the CO₂•- transfers the electron to CHCl₃ to give unstable chloroform radical anion (CHCl₃•-), which decays immediately to CHCl₂•- and Cl⁻ (reaction 12). The CHCl₂•- may either combine with CO₂•- (reaction 13) or self combines (reaction 14) to yield the observed minor products. The unidentified products could be in the gaseous form like CH₃Cl, which could not be seen by the high-performance liquid chromatography technique method.

## Halogenated compounds

Degradation of tested halogenated contaminants by CO₂•- in 10 min has been investigated [50]. and the results are presented in Fig. 8. Among the studied inorganic oxy anions, the complete elimination of ClO₂⁻ could be achieved by CO₂•- mediated reductant process (Fig. 8A). However, incomplete degradations of ClO₃⁻ and BrO₃⁻ were seen, with efficiency of 7% and 36%, respectively (Fig. 8A). Higher degradation efficiency of ClO₂⁻ than that of ClO₃⁻ is unexpected based on the standard redox potentials of the two halo acids in which ClO₂⁻ is a weaker oxidant than ClO₃⁻ (E⁰(ClO₂⁻/
<figure><img src="image_10.png" /><figcaption>**Fig. 8** Degradation of different kinds of halogenated compounds by $CO_2^•$ in 10 min. (Experimental conditions: [halogenated compounds]=1.0×10⁻⁵ M, [Formic acid]₀=1.0×10⁻² M, [H₂O₂]₀=5.0×10⁻³ M, pH 7.0, and light intensity=1.75×10⁻³ W cm⁻²). **A)** The complete elimination of $ClO_2^-$ while incomplete degradation of $ClO_3^-$ and $BrO_3^-$ by $CO_2^•$ involve complex mechanism.</figcaption></figure>

**B)** Halomethanes containing chlorine and bromine atoms could be easily removed by $CO_2^•$ in which bromo compounds are relatively easy to be eliminated, and **C)** an introduction of nitrogen atom to the chlorinated hydrocarbons assisted in complete elimination by $CO_2^•$. However, having -COOH group in such compounds resisted degradation by $CO_2^•$.

$ClO_2^−$ = +0.954 V and $E°(ClO_3^−/ClO_2^−$ = +1.210 V versus standard hydrogen electrode at 25 °C) [69], likely due to single electron transfer and oxygen transfer steps, respectively. Overall, this suggests a complex mechanism of the degradation of halo-acids by $CO_2^•$.

Brominated compounds tend to degrade with greater efficiency than the chlorinated compounds (Fig. 8B). For example, CHBr₃ has higher degradation efficiency than CHCl₃ (100% versus 75%). Literature suggests that brominated compounds have lower $E_{LUMO}$ that of chlorinated compounds and hence are easier to be attacked by $CO_2^•$. The degradation of nitrogen atom containing halogenated compounds was also studied (Fig. 8C). These chlorinated compounds have lower $E_{LUMO}$, causing complete elimination of such molecules. The halo-acetic acids have -COOH groups, which have higher values of $E_{LUMO}$ than without the -COOH group in the molecules; $CO_2^•$ is thus less prone to react with these compounds (Fig. 8C).

The review of results presented in Sect. "Water Treatment" pertains to aquatic environment with emphasis on water treatment. The reactions presented occur in water (liquid phase). Hence, these reactions are also relevant in the liquid phase of the atmosphere where many redox species are involved included $CO_2^•$. In the next section, we have reviewed potential of $CO_2^•$ in the chemistry of liquid phase of atmosphere.

## Atmospheric aqueous phase

The multiphase chemistry in atmosphere happens between chemicals in different phases, which include gas-liquid phases interactions [70]. This chemistry may play important role in acid rain, stratospheric ozone hole, oxidizing capacity of atmosphere, climate change, and effect of aerosols particles on human health [71]. For example, morphology, composition, and surfaces of atmospheric aerosol particles are complex and fundamental chemistry may be significant to unravel phase separation, mass transfer, and reactivity environment. The complexity of reactivity can be seen in large energy barriers for gas phase reactions compared to condensed phase. The reaction between $SO_2$ and $H_2O_2$ is not efficient while it occurs efficiently in water phase. The mechanisms of different types of interactions are thus far from fully understood. Many reviews of individual interactions in different environment of atmosphere have been written and an elegant opinion discussed the field of multiphase chemistry to assess changes in environment using atmospheric modelling and field measurements concurrently [70–87]. This strategy may assist to evaluate the influence of atmospheric changes to the health of humans. Among the various aspect of multiphase chemistry, the emphasis of reducing capacity of atmospheric of atmosphere is forthcoming in recent years. Below are few examples, particularly the $CO_2^•$ that may be generated with possible involvement in the chemistry of the atmosphere aqueous phase.
<figure><img src="image_11.png" /><figcaption>**Fig. 9** Possible formation and degradation pathways of CO₂•⁻ in aerosols</figcaption></figure>

Aerosol particles consist many kinds of radicals like reactive oxygen species (ROS, e.g., O₂•⁻ and HO•) and carbon centered radicals (CCR, e.g., CH₃• and C(C₆H₅)₃•. These radicals are produced via photochemical reactions, biomass burnings, biogenic aerosols sources, and ozonoly- sis of gases (e.g., Fig. 9). Under oxygenated environments, the lifetime of the radicals is usually short because of easy capture by oxygen. However, examples exist of radicals with lifetimes of days and environmentally persistent free radicals (EPFR) are such examples, which may be stabilized by metal-organic complexes and surfaces. As stated above, the stability of CO₂•⁻ increases orders of magnitude when adsorbed on metal nanoparticles and hence the possibility of its accumulation in aerosols particles to alter the chemis- try of aerosols in atmosphere. This would be the case under anoxic conditions that exist in aerosols particles. This has been shown in a study on the photochemistry of viscous iron containing organic rich particles [88]. Atmospheric particles contain abundant organic matter and can easily interact with trace metals like iron. High viscous nature of secondary organic aerosol particles has been found. These particles under viscous condition produce radicals and oxygen could not penetrate because of the collective effect of slow diffu- sion near the surface of the particle and the fast reaction. The radicals could thus be accumulated under an anoxic organic matrix. CO₂•⁻ is one of possible radical in such aerosol par- ticles. The reactions that could generate CO₂•⁻ in aerosol particles is discussed next.

Many studies have shown the generation of CO₂•⁻ dur-
ing the oxidation of carboxylate ions (e.g., formate and oxa-
late). We have discussed in Sect. 2, the reaction between
HO• and formate that produces CO₂•⁻ and this may be also
occurring in aerosols.

Formic acid is one of the most prevalent organic acids found in the atmosphere [89]. In the condensed organic phase, it can be formed by the photochemical degradation of organic constituents [90]. It accounts for over 60% of the free acid in precipitation in remote areas, and over 30% in polluted areas [89]. In addition, oxalic acid is by far the most common dicarboxylic acid in the atmospheric particle phase and is also formed by the chemical oxidation and degradation of various organic compounds in the environment [91]. Highest concentrations of oxalic acid in aerosols were measured during biomass combustion, where concentrations exceeded 1 µg m⁻³.[92] Values measured in pristine, remote regions are often in the order of tens of ng m⁻³, while values of several hundred ng m⁻³ are reached in urban regions [93, 94].

In cloud water and aerosols, the total iron content is mea-
sured within a concentration range of 0.2–21.8 ng m⁻³ and
1–10 µg m⁻³.[94, 95] Most of the dissolved iron is likely to
be present in the form of Fe(III) [94–96]. Due to the effect
of organic ligands, such as (poly)carboxylic acids, particu-
larly oxalic acid [92, 94], and humic substances [97]. and
the complex formation, it is to be expected that iron is dis-
solving in the delinquent aerosol particle. Iron complexes
comprise over 95% of the total iron found in cloud samples
[98]. While oxalate is generally considered to be the most
important organic ligand of Fe(III) in cloud chemistry due
to its high stability constant [99].

However, it is extremely difficult to estimate that this
photolytic pathway might be the more effective in the
atmospheric aqueous phase, as the radical concentrations
in clouds and deliquescent particles estimated by the mul-
tiphase mechanism CAPRAM 3.0i are [HO•]=1.4×10⁻¹⁶
to 8.0×10⁻¹² M) [73]. A few experimental studies evaluated
the hydroxyl radical steady state concentrations in cloud
waters: 7.2±5.0×10⁻¹⁶ M and 7.2±5.0×10⁻¹⁶ M. [100]
[101] Modeling to pinpoint these paths with CAPRAM is
currently under development [102].

The photolysis of iron(III)-citric/citrate is also consid-
ered to generate CO₂•⁻.[103] The conversion of dissolved
organic matter to inorganic carbon may have CO₂•⁻ as an
intermediate [104].

The production of CO₂•⁻ and its accumulation in aero-
sols may have significant role in reduction of inorganic
and organic compounds of the composition of anoxic par-
ticles. For example, reduction of NO₃⁻ to NO₂⁻. In the
presence of O₂ the CO₂•⁻ is converted to a large fraction to
O₂•⁻ and CO₂ (Fig. 9)[105]. Other possibility of reactions
of carbonate and Fe(III)-organic complexes (e.g., Fe(III)
EDTA + CO₂•⁻ →, k=5.0×10⁷ M⁻¹ s⁻¹). It should be cau-
tion to interpret the rate constants reported in homogeneous
system to the reactivity of CO₂•⁻ on the surfaces of par-
ticles. Currently, there is great interest in toxic halogenated
compounds found in the air or aerosols, particularly PFASs, which have been determined worldwide at remote sites to possible input sources [106]. Aerosols may be assisting in transporting PFASs to the different locations after input to the atmosphere. PFASs in sea spray aerosols (SSA) have been found [107], Interestingly, PFASs were in much higher concentrations (> 100,000 times) in the SSA compared to seawater concentrations. Based on the limited information known from the reactivity of CO₂•⁻ with PFASs under water conditions, the enriched CO₂•⁻ in aerosols anoxic particles may result in defluorination or convert the long-chain PFASs into shorter-chain PFASs, which tend to have higher toxicity than the parent molecule. This is significant considering PFASs return to land through different processes. In summary, the CO₂•⁻ in particles or on surfaces of particles and its implications on the chemistry in aqueous aerosol particles need the attention of the atmospheric researchers.

## Conclusion

CO₂•⁻ is a promising reductant for developing advanced reduction processes (ARPs) to degrade inorganic and organic contaminants. Among the various methods of CO₂•⁻ generation photolysis of formic acid-H₂O₂ system appears simple and straightforward with the need of sophisticated apparatus. The UV spectrum of CO₂•⁻ allows studying the kinetics of its reactions with halogenated organic compounds. However, the possible interference of the inherent oxygen in the system which has fast kinetics may interfere with determining the intended reaction rate constant. The second-order rate constants of correlation of the reactions, k(CO₂•⁻) for the chlorinated alkanes were two-order of magnitude slower than that of same reactions with hydrated electron. Furthermore, k(CO₂•⁻) has a strong negative linear relationship with EₐLUMO and therefore CO₂•⁻ likely carry out single-electron transfer steps mechanisms to reduce the compounds. CO₂•⁻ under high acid conditions is protonated (pKₐ=2.3–3.9) and how HOCO• behaves toward organic compounds remains largely unknown and future investigations over a wide acidic pH range are needed.

Studies of degrading organic contaminants are usually limited to chlorinated compounds, which showed promising potential of CO₂•⁻ for de-chlorination. However, scant information is available on its effective role in degrading PFAS, which are highly toxic and of great concerns of their occurrence in drinking water resources. This will require detailed kinetics and mechanistic investigations on the reactions of CO₂•⁻ with PFAS molecules. Finally, the potential of CO₂•⁻ as a reductant in ARPs-mediated treatment of real-water needs careful evaluation. The solution pH may not be of concern because the pKₐ is much lower than the pH

encountered in water treatment. However, dissolved oxygen may compete for CO₂•⁻ to limit its contribution to degrade the target contaminants. Furthermore, competing ions like carbonate/bicarbonate and natural organic matter also compete for CO₂•⁻. The participation of such secondary radicals produced from competing reactions need to be determined by performing decay of the target contaminants under various experimental condition and in samples of real-water. After extended data on the degradation of organic compounds of environmental relevance, kinetics computer modeling and machine learning may shed light on assessing fully the role of the CO₂•⁻ in ARPs.

Finally, a case has been made for the formation and accumulation of CO₂•⁻ in aqueous aerosols. This is usually determined by laboratory measurements of a few reduction reactions under laboratory conditions and therefore, investigations of reactions more relevant to atmospheric or aerosols conditions are needed to determine if CO₂•⁻ has a role as a reactive species to reduce inorganic molecules like nitrate and organics such as PFASs. These studies in combination with field measurements and kinetic modeling will determine if CO₂•⁻ is of significant relevance to the chemistry of aqueous atmosphere.

**Acknowledgements** V.K. Sharma would like to thank the Alexander von Humboldt Foundation, Germany for granting a Humboldt Research Award under 1241014 to him for staying at the Leibniz-Institute for Tropospheric Research (TROPOS) with its Atmospheric Chemistry Department (ACD) as its base to foster new scientific collaborations. J.D.R. was supported by a grant from the U.S. National Science Foundation (NSF) (CHE-2305078). We sincerely thank you the anonymous reviewers for their comments which improved the paper greatly.

**Funding** Humboldt Research Award J.D.R. was supported by a grant from the U.S. National Science Foundation (NSF) (CHE-2305078).

## Declarations

**Conflict of interests** VKS and EL are Editors of Environmental Chemistry Letters. The authors declare no conflict of interest.
## References

1. Dong X, Roeckl JL, Waldvogel SR et al (2021) Merging shuttle reactions and paired electrolysis for reversible vicinal dihalogenations. Science 371:507-514. https://doi.org/10.1126/science.abf2974

2. Rozsypal T, Pejchal J, Opravil J et al (2025) Destruction of chemical weapons stockpiles in the Russian Federation: a review. Environ Chem Lett 23:393-418. https://doi.org/10.1007/s10311-024-01812-5

3. He K, Feerick A, Jin H et al (2024) Retention of per- and polyfluoroalkyl substances by syringe filters. Environ Chem Lett 22:1569-1579. https://doi.org/10.1007/s10311-024-01718-2

4. Morin-Crini N, Lichtfouse E, Liu G et al (2022) Worldwide cases of water pollution by emerging contaminants: a review. Environ Chem Lett 20:2311-2338. https://doi.org/10.1007/s10311-022-01447-4

5. Kodavanti PRS, Costa LG, Aschner M: Chapter One - Perspective on halogenated organic compounds, in: Adv. Neurotoxicol., edited by: Kodavanti, P. R. S., Aschner, M., and Costa, L. G., Academic Press, 1-25, https://doi.org/10.1016/bs.ant.2023.06.001, 2023.

6. Li X, Chevez T, De Silva AO et al (2021) Which of the (Mixed) Halogenated n-Alkanes Are Likely To Be Persistent Organic Pollutants? Environ Sci Technol 55:15912-15920. https://doi.org/10.1021/acs.est.1c05465

7. Tokranov AK, Ransom KM, Bexfield LM et al (2024) Predictions of groundwater PFAS occurrence at drinking water supply depths in the United States. Science 386:748-755. https://doi.org/10.1126/science ado6638

8. Li Q, Hu Y, Wang P et al (2025) Influence of prenatal exposure to per- and polyfluoroalkyl substances under gestational diabetes mellitus on birth and child growth. Environ Chem Lett 23:765-770. https://doi.org/10.1007/s10311-025-01820-z

9. Sharma VK, Ma X, Shetty D et al (2025) Recent advances in covalent organic frameworks for PFAS removal: a comparative review on 2D and 3D architectures. J Chem Eng 519:164298. https://doi.org/10.1016/j.cej.2025.164298

10. Sharma VK, Ma X, Zboril R (2023) Single atom catalyst-mediated generation of reactive species in water treatment. Chem Soc Rev 52:7673-7686. https://doi.org/10.1039/D3CS00627A

11. Tan S, Wang R, Wang K et al (2025) Unravelling the structure-dependent defluorination mechanisms of per- and polyfluoroalkyl substances by hydrated electrons in UV/sulfite. Nat Water 3:734-745. https://doi.org/10.1038/s44221-025-00449-0

12. Zhu Z-S, Wang Y, Wang P et al (2025) Multidimensional engineering of single-atom cobalt catalysts for ultrafast Fenton-like reactions. Nat Water 3:211-221. https://doi.org/10.1038/s44221-024-00382-8

13. Kumar A, Sharma A, Majumder P et al (2025) Chemical and microalgal conversion of carbon dioxide into fuels and materials: a review. Environ Chem Lett 23:1209-1229. https://doi.org/10.1007/s10311-025-01855-2

14. Sillu D, Achal V (2024) Carbon dioxide sequestration with carbon anhydrase nanobiocatalysts: a review. Environ Chem Lett 22:2213-2239. https://doi.org/10.1007/s10311-024-01755-x

15. Zhang Z, Li Y, Zhong Y, et al. (2025) sp2/sp3-Hybridized nitrogen-mediated electrochemical CO2 capture and utilization. Sci. Adv., 11, eadw6592. https://doi.org/10.1126/sciadv.adw6592

16. Cauwenbergh R, Goyal V, Maiti R et al (2022) Challenges and recent advancements in the transformation of CO2 into carboxylic acids: straightforward assembly with homogeneous 3d metals. Chem Soc Rev 51:9371-9423. https://doi.org/10.1039/D1CS00921D

17. Chu C, Yan Y, Ma J et al (2023) Implementation of laser flash photolysis for radical-induced reactions and environmental implications. Water Res 244:120526. https://doi.org/10.1016/j.watres.2023.120526

18. Xu P, Wang S, Xu H et al (2023) Dicarboxylation of alkenes with CO2 and formate via photoredox catalysis. ACS Catal 13:2149-2155. https://doi.org/10.1021/acscatal.2c06377

19. Fang L, Liu K, Xu L et al (2024) Rechargeable carbonaceous supercapacitor for sustainable superoxide generation and pollutant abatement. Nat Water 2:485-495. https://doi.org/10.1038/s44221-024-00241-6

20. Minakata D (2024) Development of an elementary reaction-based kinetic model to predict the aqueous-phase fate of organic compounds induced by reactive free radicals. Acc Chem Res 57:1658-1669. https://doi.org/10.1021/acs.accounts.4c00021

21. von Gunten U (2024) Oxidation processes and me. Water Res 253:121148. https://doi.org/10.1016/j.watres.2024.121148

22. Wang Y, Liu Y, Zhang H et al (2025) Carbonaceous materials in structural dimensions for advanced oxidation processes. Chem Soc Rev 54:2436-2482. https://doi.org/10.1039/D4CS00338A

23. Fleming AM, Burrows CJ (2020) On the irrelevancy of hydroxyl radical to DNA damage from oxidative stress and implications for epigenetics. Chem Soc Rev 49:6524-6528. https://doi.org/10.1039/D0CS00579G

24. Kottapurath Vijay A, Sharma VK, Meyerstein D (2023) Overlooked Formation of Carbonate Radical Anions in the Oxidation of Iron(II) by Oxygen in the Presence of Bicarbonate. Angew. Chem., Int. Ed., 62, e202309472. https://doi.org/10.1002/anie.202309472

25. Horne GP, Zalupski PR, Daubaras DL et al (2020) Radiolytic degradation of formic acid and formate in aqueous solution: modeling the final stages of organic mineralization under advanced oxidation process conditions. Water Res 186:116314. https://doi.org/10.1016/j.watres.2020.116314

26. Chandimali N, Bak SG, Park EH et al (2025) Free radicals and their impact on health and antioxidant defenses: a review. Cell Death Discov 11:19. https://doi.org/10.1038/s41420-024-02278-8

27. Barrios B, Minakata D (2023) Aqueous-phase single-electron transfer calculations for carbonate radicals using the validated Marcus theory. Environ Sci Technol Lett 10:204-209. https://doi.org/10.1021/acs.estlett.2c00913

28. Gui Y-Y, Yan S-S, Wang W et al (2023) Exploring the applications of carbon dioxide radical anion in organic synthesis. Sci Bull 68:3124-3128. https://doi.org/10.1016/j.scib.2023.11.018

29. Wang S, Xu P, Zhu X (2023) CO2 radical anion in photochemical dicarboxylation of alkenes. ChemCatChem 15:e202300695. https://doi.org/10.1002/cctc.202300695

30. Xiao W, Zhang J, Wu J (2023) Recent advances in reactions involving carbon dioxide radical anion. ACS Catal 13:15991-16011. https://doi.org/10.1021/acscatal.3c04125

31. Matsuo B, Majhi J, Granados A et al (2023) Transition metal-free photochemical C-F activation for the preparation of difluorinated-oxindole derivatives. Chem Sci 14:2379-2385. https://doi.org/10.1039/D2SC06179A

32. Shen Q, Cao K, Wen X et al (2024) Formate salts: the rediscovery of their radical reaction under light irradiation opens new avenues in organic synthesis. Adv Synth Catal 366:4274-4293. https://doi.org/10.1002/adsc.202400682

33. Shah H, Hartman SP, Weinhouse S (1979) Formation of carbonyl chloride in carbon tetrachloride metabolism by rat liver in vitro. Cancer Res 39:3942-3947

34. Leonzio G, Hankin A, Shah N (2024) CO2 electrochemical reduction: a state-of-the-art review with economic and environmental analyses. Chem Eng Res Des 208:934-955. https://doi.org/10.1016/j.cherd.2024.07.014
35. Mayer JM (2011) Understanding hydrogen atom transfer: from bond strengths to Marcus theory. *Acc Chem Res* 44:36–46. https://doi.org/10.1021/ar100093z

36. Hietala J, Vuori A, Johnsson P, et al.: Formic Acid, in: Ullmann's Encyclopedia of Industrial Chemistry, 1–22, https://doi.org/10.1021/ja00775a014

37. Song L, Wang W, Yue J-P et al (2022) Visible-light photocatalytic di- and hydro-carboxylation of unactivated alkenes with CO2. *Nat Catal* 5:832–838. https://doi.org/10.1038/s41929-022-00841-z

38. Yan S-S, Liu S-H, Chen L et al (2021) Visible-light photoredox-catalyzed selective carboxylation of C(sp3)-F bonds with CO2. *Chem* 7:3099–3113. https://doi.org/10.1016/j.chempr.2021.08.004

39. Neupane P, Bartels DM, Thompson WH (2024) Exploring the unusual reactivity of the hydrated electron with CO2. *J Phys Chem B* 128:567–575. https://doi.org/10.1021/acs.jpcb.3c06935

40. Rybkin VV (2020) Mechanism of aqueous carbon dioxide reduction by the solvated electron. *J Phys Chem B* 124:10435–10441. https://doi.org/10.1021/acs.jpcb.0c07859

41. Armstrong DA, Huie RE, Koppenol WH et al (2015) Standard electrode potentials involving radicals in aqueous solution: inorganic radicals (IUPAC technical report). *Pure Appl Chem* 87:1139–1150. https://doi.org/10.1515/pac-2014-0502

42. Jiang Z, Clavaguéra C, Hu C et al (2023) Direct time-resolved observation of surface-bound carbon dioxide radical anions on metallic nanocatalysts. *Nat Commun* 14:7116. https://doi.org/10.1038/s41467-023-42936-6

43. Gordon S, Hart EJ, Matheson MS et al (1963) Reactions of the hydrated electron. *Discuss Faraday Soc* 36:193–205. https://doi.org/10.1039/DF9633600193

44. Grills DC, Lymar SV (2018) Radiolytic formation of the carbon dioxide radical anion in acetonitrile revealed by transient IR spectroscopy. *Phys Chem Chem Phys* 20:10011–10017. https://doi.org/10.1039/C8CP00977E

45. Neta P, Simic M, Hayon E (1969) Pulse radiolysis of aliphatic acids in aqueous solutions. I. Simple monocarboxylic acids. *J Phys Chem* 73:4207–4213. https://doi.org/10.1021/j100846a029

46. Buxton GV, Sellers RM (1973) Acid dissociation constant of the carboxyl radical. Pulse radiolysis studies of aqueous solutions of formic acid and sodium formate. *J. Chem. Soc., Faraday Trans.* 1, 69, 555–559. https://doi.org/10.1039/F19736900555

47. Flyunt R, Schuchmann MN, von Sonntag C (2001) A common carbanion intermediate in the recombination and proton-catalyzed disproportionation of the carboxyl radical anion, CO2-, in aqueous solution. *Chem Eur J* 7:796–799. https://doi.org/10.1002/1521-3765(20010216)7:4%3c796::AID-CHEM796%3e3.0.CO; 2-J

48. Hartman KO, Hisatsune IC (1966) Infrared spectrum of carbon dioxide anion radical. *J Chem Phys* 44:1913–1918. https://doi.org/10.1063/1.1726961

49. Janik I, Tripathi GNR (2016) The nature of the CO2– radical anion in water. *J Chem Phys* 144:154307. https://doi.org/10.1063/1.4946868

50. Zhang X, Zhang R, Ren P et al (2024) Carbon dioxide radical anion mediated dehalogenation kinetics and mechanisms of halogenated alkanes. *Water Res* 259:121799. https://doi.org/10.1016/j.watres.2024.121799

51. Martell AE, Smith RM: Carboxylic Acids, in: Critical Stability Constants: First Supplement, edited by: Martell, A. E., and Smith, R. M., Springer US, Boston, MA, 284–332, https://doi.org/10.1007/978-1-4615-6761-5_10, 1982.

52. Fojtik A, Czapski G, Hglein A (1970) Pulse radiolytic investigation of the carboxyl radical in aqueous solution. *J Phys Chem* 74:3204–3208. https://doi.org/10.1021/j100711a008

53. Cohen H, Meyerstein D (1972) Mechanism of reduction of cobalt(III) and ruthenium(III) hexaamine complexes by several aliphatic radicals. *J Am Chem Soc* 94:6944–6948. https://doi.org/10.1021/ja00775a014

54. Rumble J (2008) NIST 40. NDRL/NIST Solution Kinetics Database: Version 3.0.

55. Weller C, Horn S, Herrmann H (2013) Effects of Fe(III)-concentration, speciation, excitation-wavelength and light intensity on the quantum yield of iron(III)-oxalato complex photolysis. *J Photochem Photobiol A Chem* 255:41–49. https://doi.org/10.1016/j photochem.2013.01.014

56. Mangiante DM, Schaller RD, Zarzycki P et al (2017) Mechanism of ferric oxalate photolysis. *ACS Earth Space Chem* 1:270–276. https://doi.org/10.1021/acsearthspacechem.7b00026

57. Herburger A, Ončák M, Siu C-K et al (2019) Infrared spectroscopy of size-selected hydrated carbon dioxide radical anions CO2-(H2O) (n=2–61) in the C–O stretch region. *Chem Eur J* 25:10165–10171. https://doi.org/10.1002/chem.201901650

58. Ovenall DW, Whiffen DH (1961) Electron spin resonance and structure of the CO- 2 radical ion. *Mol Phys* 4:135–144. https://doi.org/10.1080/00268976100100181

59. Chiesa M, Giamello E (2007) Carbon dioxide activation by surface excess electrons: an EPR study of the CO2– radical ion adsorbed on the surface of MgO. *Chem Eur J* 13:1261–1267. http://doi.org/10.1002/chem.200600792

60. Villamena FA, Locigno EJ, Rockenbauer A et al (2006) Theoretical and experimental studies of the spin trapping of inorganic radicals by 5,5-dimethyl-1-pyrroline N-oxide (DMPO). 1. Carbon dioxide radical anion. *J Phys Chem A* 110:13253–13258. http://doi.org/10.1021/jp064892m

61. Jiang Z, Adjei D, Denisov SA et al (2023) Transient kinetics of short-chain perfluoroalkyl sulfonate with radiolytic reducing species. *Environ Sci Technol* Lett 10:59–65. https://doi.org/10.1021/acs.estlett.2c00837

62. Shoute LCT, Mittal JP, Neta P (1996) Reduction and defluorination of pentafluorophenol in aqueous solutions. *J Phys Chem* 100:3016–3019. https://doi.org/10.1021/jp9513374

63. Shandalov E, Zilbermann I, Maimon E et al (2004) Radical catalyzed debromination of bromo-alkanes by formate in aqueous solutions via a hydrogen atom transfer mechanism. *Tetrahedron* Lett 45:989–992. https://doi.org/10.1016/j.tetlet.2003.11.091

64. Gu X, Lu S, Fu X et al (2017) Carbon dioxide radical anion-based UV/S2O82-/HCOOH reductive process for carbon tetrachloride degradation in aqueous solution. *Sep Purif Technol* 172:211–216. https://doi.org/10.1016/j.seppur.2016.08.019

65. Lutze HV, Brekenfeld J, Naumov S et al (2018) Degradation of perfluorinated compounds by sulfate radicals – new mechanistic aspects and economical considerations. *Water Res* 129:509–519. https://doi.org/10.1016/j.watres.2017.10.067

66. Köster R, Asmus KD (1971) Die Reduktion von Tetrachlorkohlenstoff durch hydratisierte Elektronen, H-Atome und reduzierende Radikale / The Reduction of Carbon Tetrachloride by Hydrated Electrons, H-Atoms, and Reducing Radicals. Z Naturforsch B 26:1104–1108. https://doi.org/10.1515/znb-1971-1106

67. Huang L-Z, Lin J, Chen W et al (2024) Generation of carbon dioxide anion radical by UV/small molecular monocarboxylic acid system for reductive dechlorination of chlorinated alkanes. *J Environ Sci* 158:50–59. https://doi.org/10.1016/j.jes.2024.12.025

68. Zhang Q, Wang X, Wu N et al (2025) Advanced reduction processes initiated by oxidative radicals for trichloroacetic acid degradation: performance, radical generation, and mechanism. *Water Res* 268:122587. https://doi.org/10.1016/j.watres.2024.122587

69. Chiang MK, Su L, Raman K et al (2023) Reversible chlorite anion/chlorine dioxide redox couple for low-cost energy storage. *J Phys Chem C* 127:3921–3927. https://doi.org/10.1021/acs.jpcc.2c08136
70. Tilgner A, Schaefer T, Alexander B et al (2021) Acidity and the multiphase chemistry of atmospheric aqueous particles and clouds. Atmos Chem Phys 21:13483–13536. https://doi.org/10.5194/acp-21-13483-2021

71. Pöschl U, Shiraiwa M (2015) Multiphase chemistry at the Atmosphere-Biosphere Interface influencing climate and public health in the Anthropocene. Chem Rev 115:4440–4475. https://doi.org/10.1021/cr500487s

72. Abbatt JPD, Lee AKY, Thornton JA (2012) Quantifying trace gas uptake to tropospheric aerosol: recent advances and remaining challenges. Chem Soc Rev 41:6555–6581. https://doi.org/10.5194/acp-23-9765-2023

73. Herrmann H, Schaefer T, Tilgner A et al (2015) Tropospheric aqueous-phase chemistry: kinetics, mechanisms, and its coupling to a changing gas phase. Chem Rev 115:4259–4334. https://doi.org/10.1021/cr500447k

74. Abbatt JPD, Ravishankara AR (2023) Opinion: atmospheric multiphase chemistry – past, present, and future. Atmos Chem Phys 23:9765–9785. https://doi.org/10.5194/acp-23-9765-2023

75. Burkholder JB, Abbatt JPD, Barnes I et al (2017) The essential role for laboratory studies in atmospheric chemistry. Environ Sci Technol 51:2519–2528. https://doi.org/10.1021/acs.est.6b04947

76. Finlayson-Pitts BJ (2003) The tropospheric chemistry of sea salt: a molecular-level view of the chemistry of NaCl and NaBr. Chem Rev 103:4801–4822. https://doi.org/10.1021/cr020653t

77. George IJ, Abbatt JPD (2010) Heterogeneous oxidation of atmospheric aerosol particles by gas-phase radicals. Nat Chem 2:713–722. https://doi.org/10.1038/nchem.806

78. Jacob DJ (2000) Heterogeneous chemistry and tropospheric ozone. Atmos Environ 34:2131–2159. https://doi.org/10.1016/S1352-2310(99)00462-8

79. Kolb CE, Cox RA, Abbatt JPD et al (2010) An overview of current issues in the uptake of atmospheric trace gases by aerosols and clouds. Atmos Chem Phys 10:10561–10605. https://doi.org/10.5194/acp-10-10561-2010

80. McNeill VF (2015) Aqueous organic chemistry in the atmosphere: sources and chemical processing of organic aerosols. Environ Sci Technol 49:1237–1244. https://doi.org/10.1021/es5043707

81. McNeill VF, Grannas AM, Abbatt JPD et al (2012) Organics in environmental ices: sources, chemistry, and impacts. Atmos Chem Phys 12:9653–9678. https://doi.org/10.5194/acp-12-9653-2012

82. Pye HOT, Nenes A, Alexander B et al (2020) The acidity of atmospheric particles and clouds. Atmos Chem Phys 20:4809–4888. https://doi.org/10.5194/acp-20-4809-2020

83. Rudich Y (2003) Laboratory perspectives on the chemical transformations of organic matter in atmospheric particles. Chem Rev 103:5097–5124. https://doi.org/10.1021/cr020508f

84. Rudich Y, Donahue NM, Mentel TF (2007) Aging of organic aerosol: bridging the gap between laboratory and field studies. Annu Rev Phys Chem 58:321–352. https://doi.org/10.1146/annurev.physchem.58.032806.104432

85. Usher CR, Michel AE, Grassian VH (2003) Reactions on mineral dust. Chem Rev 103:4883–4940. https://doi.org/10.1021/cr020657y

86. Zhang Y, Richards DS, Grotemeyer EN, et al. (2022) Near-UV and Visible Light Degradation of Iron (III)-Containing Citrate Buffer: Formation of Carbon Dioxide Radical Anion via Fragmentation of a Sterically Hindered Alkoxyl Radical. Mol. Pharmaceutics, 19, 4026-4042. https://doi.org/10.1021/acs.molpharmaceutc.2c00501

87. Faust JA (2023) PFAS on atmospheric aerosol particles: a review. Environ. Sci.: Process. Impacts, 25, 133–150. 10.1039/D2EM00002D

88. Alpert PA, Dou J, Corral Arroyo P et al (2021) Photolytic radical persistence due to anoxia in viscous aerosol particles. Nat Commun 12:1769. https://doi.org/10.1038/s41467-021-21913-x

89. Stavrakou T, Müller JF, Peeters J et al (2012) Satellite evidence for a large source of formic acid from boreal and tropical forests. Nat Geosci 5:26–30. https://doi.org/10.1038/ngeo1354

90. Malecha KT, Nizkorodov SA (2016) Photodegradation of secondary organic aerosol particles as a source of small, oxygenated volatile organic compounds. Environ Sci Technol 50:9990–9997. https://doi.org/10.1021/acs.est.6b02313

91. Tilgner A, Herrmann H (2010) Radical-driven carbonyl-to-acid conversion and acid degradation in tropospheric aqueous systems studied by CAPRAM. Atmos Environ 44:5415–5422. https://doi.org/10.1016/j.atmosenv.2010.07.050

92. Kundu S, Kawamura K, Andreae TW et al (2010) Molecular distributions of dicarboxylic acids, ketocarboxylic acids and &alpha;-dicarbonyls in biomass burning aerosols: implications for photochemical production and degradation in smoke layers. Atmos Chem Phys 10:2209–2225. https://doi.org/10.5194/acp-10-2209-2010

93. Myriokefalitakis S, Tsigaridis K, Mihalopoulos N et al (2011) Incloud oxalate formation in the global troposphere: a 3-D modeling study. Atmos Chem Phys 11:5761–5782. https://doi.org/10.5194/acp-11-5761-2011

94. Sorooshian A, Wang Z, Coggon MM et al (2013) Observations of sharp oxalate reductions in stratocumulus clouds at variable altitudes: organic acid and metal measurements during the 2011 E-PEACE campaign. Environ Sci Technol 47:7747–7756. https://doi.org/10.1021/es4012383

95. Zhang H, Li R, Huang C et al (2023) Seasonal variation of aerosol iron solubility in coarse and fine particles at an inland city in northwestern China. Atmos Chem Phys 23:3543–3559. https://doi.org/10.5194/acp-23-3543-2023

96. Deguillaume L, Leriche M, Desboeufs K et al (2005) Transition metals in atmospheric liquid phases: sources, reactivity, and sensitive parameters. Chem Rev 105:3388–3431. https://doi.org/10.1021/cr040649c

97. Scheinhardt S, Müller K, Spindler G et al (2013) Complexation of trace metals in size-segregated aerosol particles at nine sites in Germany. Atmos Env 74:102–109. https://doi.org/10.1016/j.atmosenv.2013.03.023

98. González AG, Bianco A, Boutorh J et al (2022) Influence of strong iron-binding ligands on cloud water oxidant capacity. Sci Total Environ 829:154642. https://doi.org/10.1016/j.scitotenv.2022.154642

99. Tapparo A, Di Marco V, Badocco D et al (2020) Formation of metal-organic ligand complexes affects solubility of metals in airborne particles at an urban site in the Po valley. Chemosphere 241:125025. https://doi.org/10.1016/j.chemosphere.2019.125025

100. Anastasio C, McGregor KG (2001) Chemistry of fog waters in California's Central Valley: 1. In situ photoformation of hydroxyl radical and singlet molecular oxygen. Atmos Env 35:1079–1089. https://doi.org/10.1016/S1352-2310(00)00281-8

101. Lallement A, Vinatier V, Brigante M et al (2018) First evaluation of the effect of microorganisms on steady state hydroxyl radical concentrations in atmospheric waters. Chemosphere 212:715–722. https://doi.org/10.1016/j.chemosphere.2018.08.128

102. Herrmann H, Tilgner A, Barzaghi P et al (2005) Towards a more detailed description of tropospheric aqueous phase organic chemistry: CAPRAM 3.0. Atmos Env 39:4351–4363. https://doi.org/10.1016/j.atmosenv.2005.02.016

103. Dou J, Alpert PA, Corral Arroyo P, et al. (2021) Photochemical degradation of iron(III) citrate/citric acid aerosol quantified with the combination of three complementary experimental techniques and a kinetic process model. Atmos. Chem. Phys., 21, 315-338. https://doi.org/10.5194/acp-21-315-2021
104. Miller WL, Zepp RG (1995) Photochemical production of dissolved inorganic carbon from terrestrial organic matter: significance to the oceanic organic carbon cycle. Geophys Res Lett 22:417-420. https://doi.org/10.1029/94GL03344

105. Melssen E, Bish DL, Losovyj Y et al (2024) Photoreduction of nitrate to HONO and NOx by organic matter in the presence of iron and aluminum. ACS Earth Space Chem 8:2587-2598. https://doi.org/10.1021/acsearthspacechem.4c00252

106. Pandamkulangara Kizhakkethil J, Shi Z, Bogush A, et al. (2024) Aerosolisation of per- and polyfluoroalkyl substances (PFAS)

during aeration of contaminated aqueous solutions. Atmos. Env.,
334, 120716. https://doi.org/10.1016/j.atmosenv.2024.120716

107. Sha B, Johansson JH, Tunved P et al (2022) Sea spray aerosol (SSA) as a source of perfluoroalkyl acids (PFAAs) to the atmosphere: field evidence from long-term air monitoring. Environ Sci Technol 56:228-238. https://doi.org/10.1021/acs.est.1c04277

**Publisher's Note** Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.