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Review

Chemical Oxidation and Reduction Pathways of Mercury Relevant to Natural Waters: A Review

by
Lin Si
1,*,
Brian A. Branfireun
2 and
Jessica Fierro
1,†
1
Department of Chemistry, Auburn University at Montgomery, Montgomery, AL 36117, USA
2
Department of Biology, University of Western Ontario, London, ON N6A 5B7, Canada
*
Author to whom correspondence should be addressed.
†
Current address Epic System Corporation, 1979 Milky Way, Verona, WI 53593, USA.
Water 2022, 14(12), 1891; https://doi.org/10.3390/w14121891
Submission received: 6 May 2022 / Revised: 8 June 2022 / Accepted: 9 June 2022 / Published: 12 June 2022
(This article belongs to the Section Biodiversity and Functionality of Aquatic Ecosystems)
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Abstract

:
Mercury (Hg) pollution in the environment is a global issue and the toxicity of mercury depends on its speciation. Chemical redox reactions of mercury in an aquatic environment greatly impact on Hg evasion to the atmosphere and the methylation of mercury in natural waters. Identifying the abiotic redox pathways of mercury relevant to natural waters is important for predicting the transport and fate of Hg in the environment. The objective of this review is to summarize the current state of knowledge on specific redox reactions of mercury relevant to natural waters at a molecular level. The rate constants and factors affecting them, as well as the mechanistic information of these redox pathways, are discussed in detail. Increasing experimental evidence also implied that the structure of natural organic matter (NOM) play an important role in dark Hg(II) reduction, dark Hg(0) oxidation and Hg(II) photoreduction in the aquatic environment. Significant photooxidation pathways of Hg(0) identified are Hg(0) photooxidation by hydroxyl radical (OH•) and by carbonate radical (CO3−•). Future research needs on improving the understanding of Hg redox cycling in natural waters are also proposed.

1. Introduction

Listed as one of the ten leading chemicals affecting public health by the WHO, mercury (Hg) pollution has been a focus of research in recent years [1]. Mercury speciation significantly influence its cycling, transportation, bioavailability and potential toxicity in the environment [2]. Mercury (Hg) presents mainly in two oxidation states in natural environments, i.e., Hg(0) and Hg(II) [3]. Elemental mercury (Hg(0)) is the dominant form of mercury in the atmosphere. Hg(0) has a low vapor pressure and is capable of long-distance transport. Hg(0) can be oxidized to various Hg(II) species by a number of atmospheric oxidants such as ozone (O3). These oxidized Hg(II) compounds are subject to wet or dry deposition to aquatic and soil systems [4]. Dry deposition of Hg(0) onto vegetation surfaces and soils can also occur [5]. The major mercury species in natural waters are Hg(II)-complexes with abundant ligands (e.g., dissolved organic matter (DOM)) or methylmercury (CH3Hg+) compounds [2,6]. Methylmercury species have a very high toxicity and its exposure to humans generally occurs through eating Hg-containing fish [7,8].
The redox processes involving changes of oxidation states of Hg are crucial for the prediction and remediation of mercury contamination in various environmental compartments [9,10]. The redox reactions of Hg in the aquatic environment, i.e., reduction reactions of Hg(II) and oxidation reactions of Hg(0), may impact on two important Hg-cycling processes in the environment: volatilization of Hg(0) from natural waters and methylation of Hg(II) in the aquatic environment. When the reduction/oxidation reactions result in a net production of Hg(0) in natural waters, volatilization of Hg(0) from natural waters to the atmosphere occurs. This process may enhance Hg(0) levels and could elongate its lifetime in the atmosphere [11,12,13]. On the other hand, the redox reactions of Hg may affect the amount of Hg(II) available for the methylation process and its subsequent bioaccumulation in fish [14,15]. The transformations of various Hg species in natural waters are shown in Figure 1.
The abiotic redox reactions of Hg may occur photochemically or directly in the dark. Dark reduction or oxidation reactions may dominate in deep water and sediments and could be significant on a global scale [16]. Recent field measurements and modelling data implied that dark reduction of Hg(II) drove Hg(0) evasion from the ocean [17]. Ravichandran reviewed the role of DOM on the direct chemical reduction of mercury in the dark in 2004 [6]. However, a systematic review on kinetics and mechanisms of dark chemical reduction or oxidation of Hg(II) is lacking. On the other hand, photochemical redox reactions of mercury have been a research focus due to their much faster reaction rates than dark ones [18]. Numerous field studies have observed that sunlight could facilitate the reduction and oxidation reactions of Hg in various natural surface waters [2,19]. Identification of specific photoreduction and photooxidation reactions of mercury relevant to natural waters is crucial to explain the field observations and is key for environmental Hg models [20]. Most previous reviews of photoreactions of Hg have focused on field studies, except for Zhang 2006. Zhang summarized the controlled experimental studies on the specific photochemical redox pathways of Hg relevant to the aquatic environment in 2006 [21]. This review will focus on recent advances in the identified specific photoreduction and photooxidation reactions of mercury since 2006. To our knowledge, this is the first systematic review on specific abiotic dark redox pathways of Hg since 2004 and on photochemical redox reactions of Hg since 2006. Furthermore, the current state of knowledge on the redox reaction of Hg both in the dark and under light radiation together at a molecular level provides a synthetic scheme of Hg redox pathways under different natural water conditions.
The objective of this paper is to review recent controlled experimental studies on specific redox reactions of mercury relevant to natural waters. We herein provide a systematic review on the chemical redox reactions of Hg relevant to natural waters including (i) abiotic dark reduction of Hg(II); (ii) photochemical reduction of Hg(II); (iii) abiotic dark oxidation of Hg(0); and (iv) photochemical oxidation of Hg(0) in the aquatic environment. The rate constants and factors affecting them, as well as the mechanistic information of these identified redox pathways of Hg, are discussed in detail.

2. Abiotic Dark Reduction of Hg(II) Relevant to Natural Waters

To date, the major identified abiotic dark reduction pathway of Hg(II) relevant to natural waters is the reduction of Hg(II) by natural organic matter (NOM). Table 1 summarizes the up-to-date knowledge on the major kinetic and mechanistic findings of this dark reduction pathway of Hg(II) since 2000.
Table
Table 1. Identified Abiotic Reduction Pathways of Hg(II) in the Dark.
Table 1. Identified Abiotic Reduction Pathways of Hg(II) in the Dark.
ReductantRate ConstantspHFactors Affecting
Kinetics 1		Reaction MechanismReferences
Tropical River HS 2~0.1 h−1 &
0.02 h−1		2–11+/−: pH (max. at 8)
+: [HS]
N: Hg(II):HS ratio.		slow two-step, 1st-order reaction.[22]
Tropical River HS 5+: the ratio of phenolic/carboxylic groups;
−: the amount of S in the HS,
N: molecular sizes.		slow two-step, 1st-order reaction.[23]
Nordic Reservoir DOM 3 (IHSS) 40.0290 h−1 &
0.0005 h−1		~6Binding of DOM to Hg(II).N/A[24]
Reduced Elliott soil HA 5 ~7+/−: DOM:Hg(II) ratios.Binding of HA to Hg(II) at hight DOM:Hg(II) ratios may inhibit Hg(II) reduction.[25]
Four reduced NOM 6 isolatesInitial Rates: 0.4–5.5 h−1~7+/−: DOM:Hg(II) ratios.Reduction of Hg(II) by the reduced quinones and Hg(0) oxidized by thiols in NOMs.[26]
Three different HAs 3.6, 6.8, 7.2, 8.1+/−: pH, initial [Hg(II)] or [HA]
Y: the HA types (structures)
+: temperatures, light or aqueous phase.		N/A[27]
Suwannee River HA, Suwannee River FA 7 and Eliot Soil HA from IHSS 4.0–9.0−: pH, salinity ([Cl−]), S content in HS.Hg(II) reduction suppressed by complexation of Hg(II) with S groups in HS or with Cl−, and by the back oxidation of Hg(0) by thiols in HS.[28]
+: the −COOH/−OH ratio in HS.Hg(II) reduction impacted by both the types and positioning of the functional groups in HS.
Peat HA
Coal HA
Soil HA

Peat HA
Coal HA

Soil HA		0.18 h−1 8
0.22 h−1 8
0.35 h−1 8

0.003 h−1 9
0.0003 h−1 9
(biased)
0.006 h−1 9		6.8+: initial Hg(II), or [HA] at an identical Hg(II)/DOC 10 ratio
Y: Hg(II)/DOC ratio.		Structure-specific kinetics of the proposed Hg(II)-NOM complexes (Hg(OR)2 and RSHgOR).[18]
Suwannee River HA, Suwannee River FA and Eliot Soil HA from IHSS 8+: Hg(II)/HS ratio,
Y: the sources of HS and −OH/−COOH ratio in HS;
−: [Ca2+].		Hg(II) reduction mediated via the complexes of Hg(II) with HS.
Ca2+ may impede the intermolecular reaction between Hg(II) and HS.		[29]
Reduced NOM2.42 ± 0.53 h−1 (no Cl−)
0.14 ± 0.01 h−1 (with Cl−)		7−: [Cl−].Both the formation of Hg(II)-chloride complexes and the re-oxidation of Hg(0) to Hg(II), may contribute to the retarding effect of Cl− on the Hg(II) reduction.[30]
1 “+” sign means a positive effect on the reaction kinetics; “−” sign means a negative effect on the reaction kinetics; “+/−” means a positive effect followed by a negative effect on the reaction kinetics; “Y” means it’s a factor affecting the reaction kinetics; “N” means the factor has no impact on the reaction kinetics. 2 HS: humic substances. 3 DOM: dissolved organic matter. 4 IHSS: International Humic Substances Society. 5 HA: humic acids. 6 NOM: natural organic matter. 7 FA: fulvic acids. 8 The rate constants for the proposed Hg(OR)2 complexes. 9 The rate constants for the proposed RSHgOR complexes. 10 DOC: dissolved organic carbon.
A number of studies have been performed to advance our knowledge on the direct reduction of Hg(II) by NOMs since it was first reported by Alberts et al. in 1974. We herein review the recent studies on the dark Hg(II) reduction by NOM since the review by Ravichandran (2004) [6].

2.1. The Rate Constants

The dark reduction of Hg(II) by NOM generally proceeded initially at a relatively fast rate and then slowed down. The pseudo-1st-order rate constants were determined with NOM in a large excess relative to Hg(II). Faster and slower pseudo-1st-order rate constants of Hg(II) reduction were generally obtained from a fitting of kinetic data. The large pseudo-1st-order rate constants (k1) ranged from 0.0290 h−1 to 5.5 h−1, while the smaller one (k2) ranged from 3 × 10−4 h−1 to 0.037 h−1. Although it was generally believed that these vastly different rate constants were related to how Hg(II) binds with NOM, different models were proposed to fit the kinetic data. Zheng and Hintelmann described the reduction kinetics of Hg(II) by Nordic Reservoir NOM with the rate constants to be k1 = 0.0290 ± 0.0068 h−1 and k2 = 0.0005 ± 0.0001 h−1 at pH ~6 [24]. It was suggested that these parallel pseudo-1st-order reactions may be due to various types of reducing moieties or reactive sites in NOM. Zheng et al. obtained the multiple parallel pseudo-1st-order rate constants by assuming the simultaneous reduction and oxidation of Hg((II) by four different reduced NOM isolates [26]. The reduction rate constants were estimated to be 0.4–5.5 h−1 (k1) and 0.004–0.037 h−1 (k2) at pH ~7, while the oxidation rate constants were determined to be 2.2–5.0 h−1 (k1′) and 0.02–0.18 h−1 (k2′). It was hypothesized that Hg(II) was reduced to Hg(0) by the reduced quinones on these NOM isolates and the reduction was offset by the oxidation of Hg(0) by thiols at high DOM:Hg(II) ratios. Jiang et al. determined the rate constants of the chemical reduction of Hg(II) by three different HAs in the dark for two types of proposed Hg(II)-NOM complexes, i.e., Hg(OR)2 and RSHgOR at pH = 6.8 and T = 295 ± 1 K [18]. Here, OR designated any O/N functional group on NOM. Another advance in this study is that the reported rate constants also accounted for the rearrangement of the binding sites on NOM to Hg occurring simultaneously with the Hg(II) reduction. The pseudo-1st-order rate constant for the reduction of Hg(OR)2 were 0.18, 0.22 and 0.35 h−1 for peat, coal and soil HA, respectively. For RSHgOR, the pseudo first-order rate constant was significantly lower, at 0.003, 0.006, and 0.0003 h−1 for peat, soil, and coal HA, respectively. Although a 3rd Hg(II)-NOM complexes, i.e., Hg(SR)2, was also proposed, its rate was too slow to be measured in the limited time of the experiment (53 h). More studies are clearly needed to develop a universal kinetic model on the Hg(II) reduction by NOMs.

2.2. Factors Affecting Dark Hg(II) Reduction Rates by NOMs

In an effort to better understand the reaction mechanism of Hg(II) reduction by NOMs, many researchers investigated how various environmental variables affected this reduction pathway. These factors include pH, the concentration of Hg(II) or NOM, NOM:Hg(II) ratios, NOM structures, salinity, and the presence of Cl− or Ca2+.

2.2.1. Effect of pH

The production of Hg(0) was found to be affected by the solution pH in many studies. Jiang et al. showed that the amount of Hg(II) reduction by three different humic acids (HA) maximized at pH ~7, while either highly acidic (pH 3.6) or basic (pH 8.1) conditions inhibited the Hg(0) production [27]. It was suggested that the decrease in Hg(II) reduction at a high pH was due to complexation of Hg(II) with OH− ions despite the enhanced extension of HA structures [22]. At a low pH, the formation of Hg(0) may be inhibited by the precipitation of HA rather than by competition of binding from H+ [27,31]. However, conflicting results on how the Hg(II) reduction efficiency varied with pH were obtained by Chakraborty et al. It was found that net production of Hg(0) by three different humic substances (HS) decreased as the pH increased from 4.0 to 9.0, which was likely due to increasing complexation between Hg(II) and HS [28]. In summary, multiple factors could contribute to the seemingly contradictory effect of pH on Hg(II) reduction efficiency by NOM [27]. More studies are needed to evaluate various factors contributing to the observed pH effect on this reduction pathway.

2.2.2. Effect of Hg(II)/NOM Ratio, the Concentration of Hg(II) ([Hg(II)]) or NOM ([NOM])

Both Gu et al. and Zheng et al. observed that the amount of the reducible Hg(II) and the reaction rate for the Hg(II) reduction by reduced NOMs initially increased at low NOM:Hg(II) ratios and then reduced at high NOM:Hg(II) ratios at pH ~7 and 4-h equilibrium time under anoxic, dark conditions [25,26]. Jiang et al. varied Hg(II) concentration from 0.6 to 60 µM at the HA concentration of 30 mg C/L and found that the %Hg(II) reduced with Hg(II) concentration until 6.0 µM, but declined as it was further increased. Similar trends were also observed when changing the concentration of HA at a constant Hg(II) concentration for all three HAs [27]. The amount of Hg(II) reduced by HA was shown to depend on the Hg(II)/HA ratios by Jiang et al. At an identical Hg(II)/HA ratio, the Hg(0) production increased with the concentration of either Hg(II) or HA [18]. The dark reduction of Hg(II) by several well-characterized HSs has been recently studied for 144 h at pH = 8 by Vudamala et al. The Hg(II) reduction rate by three different HSs increased with the Hg(II)/HS ratio [29].
The complexation of Hg(II) to weak and strong binding sites on NOM has been used to explain the variation of Hg(II) reduction with Hg(II)/NOM ratio, [Hg(II)] or [NOM] in these studies. The re-oxidation of Hg(0) induced by thiol functional groups on NOM may also impede Hg(II) reduction, as implied by Gu et al. and Zheng et al. [25,26] Jiang et al. [18] attempted to quantitatively explain the kinetic data in some studies [23,25,26,31,32]. It was assumed that the mass of sulfhydryl (−SH) sites were equal to 0.0015 × NOM [31]. Combined with the Hg(II)/NOM ratio, the concentrations of Hg(II) and NOM used in these studies, the fraction of weak and strong complexes of Hg(II) with NOM were estimated and linked to the different rate constants observed. Hg(II) reduction may be also limited by electron donor groups on NOM at high Hg(II)/NOM ratios [18].

2.2.3. Effect of NOM Structure

Numerous studies have shown that Hg(II) reduction by NOM in the dark varied significantly with the types of NOM used [18,27,28,29,32,33]. For example, some researchers found that both the percentage of S content and the phenolic/carboxylic group (−OH/−COOH) ratios of NOM influenced Hg(II) reduction by NOM. Rocha et al. measured Hg(0) produced from the Hg(II) reduction by different size-fractions of the Rio Negro HS [23]. At pH 5 the reduction efficiency was found to increase with the ratio of phenolic/carboxylic groups and decrease with the amount of sulfur in the HS but was barely affected by molecular sizes. These findings were later confirmed by Chakraborty et al. and Vudamala et al. by investigating Hg(II) reduction in the presence of various well-characterized HSs from the International Humic Substances Society (IHSS) [28,29]. Chakraborty et al. observed that Hg(II) reduction decreased with the decreasing −COOH/−OH ratio or with the increasing sulfur content of HS at various pHs [28]. Similar trends were observed by Vudamala et al. at different Hg(II) loadings and varying Ca2+ concentrations [29]. Chakraborty et al. also found that the reduction of Hg(II) by all three HSs was not dependent on the concentrations of either −COOH or −OH groups in HS, which implied that both the types and the positioning of the functional groups such as −COOH and −OH on the HS studied may impact on the Hg(II) reduction [28].

2.2.4. Effect of Cl−

The presence of a chloride ion (Cl−) has been found to retard the Hg(II) reduction by various NOMs under dark conditions [28,30]. Chakraborty et al. showed that Hg(II) reduction by different HSs decreased with increasing additions of NaCl salt, which was most likely caused by complexation between Hg(II) and Cl− ions at high salinity [28]. Recently, Lee et al. determined the extent and rate of Hg(II) reduction by the reduced DOM with or without the presence of Cl− under dark, anoxic conditions [34]. The results indicated that the presence of a chloride ion significantly slowed down the Hg(II) reduction reaction. Additional experiments implied that both the formation of Hg(II)-chloride complexes and the re-oxidation of Hg(0) to Hg(II) may contribute to the observed inhibiting effect of Cl− on the reduction of Hg(II) by the reduced DOM used in this study. Hg(0) could be re-oxidized to Hg(II) by semi-quinone and Cl− [35,36]. Future studies should focus on quantification of the effect of Hg(II)-Cl− complexation on the redox reactions of Hg [30].

2.2.5. Effect of Temperature and Ca2+ Ion

The effects of various factors on abiotic reduction of mercury (Hg) by three different humic acids (HA) was evaluated by Jiang et al. [27] It was observed that higher temperatures could facilitate the reduction of Hg(II) by the HAs studied. Vudamala et al. evaluated the dark reduction of Hg(II) by several well-characterized HSs at various concentrations of Ca2+ ion. Increasing concentration of Ca2+ ion resulted in a lower reduction rate of Hg(II) under dark conditions. Since Ca2+ is a hard acid, the presence of Ca2+ may impede Hg(II) reduction by preventing the intermolecular transfer between Hg(II) and HS in the dark [29].

3. Photochemical Reduction of Hg(II) Relevant to Natural Waters

To date, the identified photochemical reduction pathways of Hg(II) relevant to the aquatic environment include the photo-reduction of mercuric hydroxide, the photoly-sis of HgS22−, and the photo-reduction of Hg(II) by DOM. Table 2 summarizes the current state of knowledge on identified photochemical reduction reactions of Hg(II) relevant to natural waters. The photo-reduction of Hg(OH)2, and the photolysis of HgS22−, have been previously reviewed by Zhang in 2006 [21]. We focus on discussing recent studies on the photo-reduction of Hg(II) by DOMs since 2006. Both photoreduction of Hg(II) by naturally occurring low-molecular-weight organic compounds (LMWOCs, MW < ~1000) and photoreduction of Hg(II) in the presence of heterogeneous DOM isolates from natural water samples, were investigated. Similar to dark reduction of Hg(II) in the presence of DOMs discussed earlier, (pseudo-)1st-order rate constants for the Hg(II) photoreduction in the presence of excess DOMs were determined in most studies.
Table
Table 2. Identified Photo-reduction Pathways of Hg(II) relevant to Natural Waters.
Table 2. Identified Photo-reduction Pathways of Hg(II) relevant to Natural Waters.
ReductantRate ConstantsLight
Source		pHFactors Affecting
Kinetics 1		Reaction MechanismReferences
Hg(OH)23 × 10−7 s−1simulated sunlight7N/APrimary reaction[37]
HgS22−~10−7 s−1simulated sunlight7N/A H g S 2 2 − → H g 0 + H g S + o t h e r   p r o d u c t s [37]
Oxalic acid1.7 × 104 M−1 s−1 (no Cl−)
1.1 × 104 M−1 s−1 (with Cl−)		simulated sunlight3.9 or 7−: Cl−Secondary Mechanism by HO2 radical[38]
Oxalic acidN/A N/APrimary reaction[39]
Dicarboxylic Acids (C2–C4)1.2 × 104 M−1 s−1 (oxalic)
4.9 × 103 M−1 s−1 (malonic)
2.8 ×103 M−1 s−1 (succinic)		UV or visible lamps3.0−: Cl− or O2Both primary and secondary reactions[40]
Serine0.640 h−1simulated sunlight3.6 or 3.8−: reduced sulfur groups on LMWOCsBoth primary and secondary reactions[41]
cysteine0.047 h−1
Salicylic acid2.1–8.3 h−1UV-B4.2Y: both the substituent functional groups and their positions on the benzene ringSecondary reactions by organic free radicals[42]
Anthranilic acid2.0–8.8 h−1
phthalic acid3.4 h−1
4-hydrobenzoic acid2.4 h−1
4-aminobenzoic acid7.6 h−1
Salicylic acid0.60 h−1UV-A4.2
Anthranilic acid8.4 h−1
4-hydrobenzoic acid0.33 h−1
4-aminobenzoic acid1.8 h−1
Thioglycolic acid (TGA)(2.3 ± 0.4) × 10−5 s−1UV4.0Y: pHHg(II) reduction may be mediated via Hg(II)-TGA complex
Heterogeneous processes may be involved		[43]
cysteine (Cys)1.13 ± 0.21 day−1 (highest, pH 7, oxic)Natural sunlight7 or 3.2Y: pH, light wavelength or dissolved O2
Y: functional groups on LMWOCs		 [44]
serine (Ser)29.94 ± 4.23 day−1 (highest, pH 7, oxic)
ethylenediamine (en)2.80 ± 0.36 day−1
(pH 7, oxic)
1. see the explanations in footnote 1 of Table 1.

3.1. Photoreduction of Hg(II) by LMWOCs

Because of their known chemical structures and binding constants with Hg(II), LMWOCs are frequently used as model compounds to help decipher the complicated photoreduction of Hg(II) in the presence of heterogeneous DOM [6,41,42,43,45]. The LMWOCs employed a focus on thiol containing compounds, aromatics, quinones, dicarboxylic acids and amino acids, which contain −SR, −OR, and/or −NR functional groups. Both the rate constants and reaction mechanisms vary with specific functional groups (−SR, aromatic, −OR or −NR) on these LMWOCs.

3.1.1. Reduced Sulfur Functional Groups (−SR)

As shown in Table 2, the (pseudo-)1st-order rate constants for LMWOCs without −SR groups (0.120–8.8 h−1) were typically higher than those with -SR groups (0.0014–0.047 h−1) despite a wide range of experimental conditions employed. Studies have shown that the presence of strong binding sites on LMWOCs to Hg(II) (e.g., −SH) may significantly slow down the Hg(II) photoreduction. Zheng and Hintelmann examined the mercury photochemical transformation by 12 different low-molecular-weight organic compounds (LMWOCs) [41]. The photoreduction of Hg(II) by LMWOCs without reduced sulfur groups (e.g., serine) were much faster than those with reduced sulfur groups (e.g., cysteine). The pseudo-1st-order rate constant for Hg(II) photoreduction by serine (no −SH group) was determined to be 0.640 h−1 at T = 293–295 K and pH 3.6, which was more than 10 times higher than that by cysteine (0.047 h−1, contains −SH group). Studies by Si and Ariya also reached similar conclusions for different LMWOCs [43,46]. It was found that the apparent rate constant increased ~50 times when replacing one −SH group (e.g., 1-propanethiol) with one −COOH group (e.g., thioglycolic acid) under similar experimental conditions.
Several mechanistic studies have implied that the photoreduction of Hg(II) by sulfur-containing organic compounds may be a primary reduction process [41,45]. Isotope fractionation experiments were performed for Hg(II) photoreduction in the presence of selected S-containing LMWOCs (cysteine, glutathione, cysteamine, methionine, and thiourea) by Zheng and Hintelmann, and the observation of a reversed magnetic isotope effect ((−) MIE) indicated direct photolysis of Hg(II)-organic complexes [41]. Systematic kinetic and product studies for photochemical reactions between Hg(II) and thioglycolic acid also suggested that the reduction may occur via Hg(II)-thiol complexes [43].
Most recently, Motta et al. investigated mercury isotope fractionation during the Hg(II) photoreduction in the presence of cysteine. With cysteine, the highest Hg(II) photoreduction rate was observed at pH 7.2 under oxic conditions with UV exposure. It was discovered that both the reaction rates and the sign of MIE were dependent on pH and dissolved oxygen [44].

3.1.2. Aromatics

The photochemical reduction of Hg(II) by aromatics may be a secondary reaction and its rate were affected by both the substituent functional groups and their positions on the benzene ring [42]. He et al. measured the pseudo-1st-order rate constants for the Hg(II) photoreduction in the presence of salicylic, 4-hydrobenzoic, anthranilic, 4-aminobenzoic, or phthalic acid, at pH 4.2 or 8, T = 23 ± 1 °C under UV-A or UV-B irradiation and oxygen-free conditions. These photoreduction rates were affected by both the substituent functional groups (−COOH, −NH2 or −OH) and their positions on the benzene ring. Hg(II) photoreduction rates were positively related to UV absorption spectra and the concentrations of these aromatics, which indicated a secondary reaction mechanism where Hg(II) may be reduced by organic free radicals formed upon photolysis of these aromatic compounds.

3.1.3. Oxygenated (−OR) and/or Amine (−NR) Functional Groups

The isotope fractionation patterns for Hg(II) photoreduction by six sulfurless LMWOCs were recorded by Zheng and Hintelmann [41]. These sulfurless LMWOCs contain −OR (−COOH, −OH) and/or −NR functional groups. The results implied that both direct photolysis and secondary reduction are possible mechanisms for these LMWOCs containing oxygenated and/or amine groups only. Systematic kinetic and product studies for photochemical reactions between Hg(II) and selected dicarboxylic acids also suggested that both primary and secondary reductions could occur [40].
Recently, Motta et al. performed kinetic and mechanistic studies on Hg(II) photoreduction by cysteine, serine or ethylenediamine at various pHs, light wavelengths and dissolved oxygen [44]. With serine, the Hg(II) reduction was fastest at pH 7 under oxic conditions with UV-vis radiation. The order of the photoreduction rate of Hg(II) was serine > ethylenediamine > cysteine. The (+) MIE was consistently observed for Hg(II) reduction by serine upon UV irradiation with no impact from either pH or dissolved oxygen. Secondary reduction may take place for photoreduction of Hg(II) complexed to −OR under visible radiation, evidenced by a small (−) MIF. The (−) MIE was obtained for Hg(II)-NR complexes at pH 7 with UV-vis radiation. However, the effects of pH light wavelength and dissolved oxygen were not evaluated for ethylenediamine that contains -NR functional groups only.

3.2. Photoreduction of Hg(II) in the Presence of Natural DOM Isolates

Contradictory results exist in terms of the relationship between DOM concentrations and Hg(II) photoreduction. Some scientists have observed the decrease in Hg(0) formation with increasing DOM concentrations [12,47,48]. A positive relationship between DOM concentrations and Hg(II) photoreduction rates was reported by O’Driscoll et al. [49]. Lee et al. showed that when varying the HA concentration from 0 to 4 mg/L, the photoreduction rate of Hg(II) in presence of a humic acid salt increased the most from 0 to 1 mg/L, and then slowed down at higher HA concentrations [34]. Other studies revealed no correlation between Hg(0) production and DOM concentrations [50,51]. It has been suggested that this discrepancy may be due to variations in the DOM structures used in these studies [52,53]. It is the DOM composition rather than the threshold in DOM concentration that could significantly impact the photoreduction of Hg(II) in the environment [10,50,53,54,55].
Despite the difficulty in identify the exact structure of heterogeneous DOMs in natural waters [56,57], several recent studies have shed light on how certain functional groups on DOMs influence the kinetics and reaction mechanisms of Hg(II) photoreduction in presence of natural DOMs. Zheng and Hintelmann studied the photoreduction of Hg(II) in filtered natural water samples [32]. The obtained kinetic data had a better fit as a combination of multiple different photoreduction reactions corresponding to various DOM binding sites to Hg(II). Wang et al. studied the influence of Hg/DOM ratios on the photoreduction of Hg(II) using the same natural water sample at the same time. It was discovered that either a low Hg/DOM ratio (<1134 ng/mg) or a high Hg/DOM ratio (>1134 ng/mg) hindered Hg(II) photoreduction upon sunlight radiation at room temperature. The Hg(II) photoreduction at a low Hg(II)/DOM ratio was possibly impeded due to the strong binding of Hg(II) to the reduced sulfur groups in DOM, while at a high Hg(II)/DOM ratio, this may be caused by less solar energy available for a single-DOM bond [58]. Jeremiason et al. observed similar Hg(II) photoreduction rates in the presence of various DOM isolates only, thiol-containing LMWOCs plus DOM, or only thiol-containing LMWOCs at pH 7, which supported a primary reaction mechanism for the photoreduction of Hg(II) bound to thiol sites on DOM [45]. O’Driscoll et al. suggested that the increase in carboxylic functional groups (−COOH) in DOC structures may result in an increased number of weak binding sties on DOM to Hg and therefore, more photo-reducible Hg(II) by DOMs from four different lakes in central Quebec [52,59].

3.3. Effect of Light Wavelength and Intensity on Hg(II) Photoreduction

Many previous studies have observed high photoreduction rates of Hg(II) upon UV radiation and with increased light intensity [47,60]. Recently, Oh et al. measured the Hg(II) photoreduction rates upon photolysis of the filtered water samples spiked with 100 ppt Hg(II), at UVA intensities of 0.5, 1.5, and 3 mW cm−2 and at UVB intensities of 0.02, 0.05, and 0.2 mW cm−2, respectively. Despite the lower intensities used for UVB compared with UVA, higher reduction rates were obtained for UVB rather than for UVA radiation. Similar results were obtained by Lee et al. Hg(0) production evidently increased under UVA, UVB, or UVC radiation, while UVB irradiation was most effective for the photoreduction of Hg(II) by humic acid sodium salt from Aldrich [34].

4. Chemical Oxidation Pathways of Hg(0) in the Dark

The identified abiotic Hg(0) oxidation pathways in the dark are chemical oxidation of Hg(0) by low-molecular-weight (LMW) thiols and by reduced NOMs. The kinetic data and mechanistic information were summarized in Table 3. Thiol functional groups have been shown to be crucial for abiotic dark Hg(0) oxidation reactions in natural waters [25,26,61,62].
Table
Table 3. Identified Abiotic Oxidation Pathways of Hg(II) in the Dark.
Table 3. Identified Abiotic Oxidation Pathways of Hg(II) in the Dark.
OxidantRate ConstantspHFactors Affecting
Kinetics 1		Reaction MechanismReferences
Reduced Elliott soil HA 2 from IHSS 2N/A7+: [reduced HA]ligand-induced oxidative complexation of Hg(0) by thiol groups in HA[25]
Reduced NOMs 2Initial Rates: 2.21–5.42 h−17+: DOM:Hg(II) ratios 2Same as above[26,63]
Various thiol organic ligands0.01–2.18 h−17Y: thiol species and thoil to Hg ratiosSame as above[64]
Reduced NOM0.14 ± 0.01 h−1 (no Cl−)
0.25 ± 0.07 h−1 (with Cl−)		7−: Cl−N/A[30]
1 See explanations in footnote 1 of Table 1. 2 See explanations in footnotes 3–6 of Table 1.
Experimental evidence suggested that dark Hg(0) oxidation occurred via the adsorption of LMW thiols [62] or thiol sites of reduced NOMs [25,26,64] onto Hg(0), followed by cleavage of S-H bonds and charge transfer to electron acceptors. However, the organic compounds containing either oxidized sulfur (e.g., disulfides) or no sulfur, showed little reactivity with Hg(0) [62].
Kinetic experiments showed that the oxidation of Hg (0) by reduced NOMre proceeded rapidly at the outset but slowed down later [26]. Similar multistage oxidation kinetics were also found by Zheng et al. in 2019 [61]. The determined pseudo-1st-oder rate constants of chemical oxidation of Hg(0) varied widely (Table 3). The initial rate constants of the abiotic dark oxidation of Hg(0) with the reduced NOM from various sources at pH 7 was determined to be between 2.21–5.42 h−1 [26]. The rate constants of dark Hg(0) oxidation by various LMW thiols were reported to be 0.01–2.18 h−1 at pH 7 under dark, anoxic conditions by Zheng et al. in 2013 [62]. The large variations of these rate constants were found to be affected by the chemical and structural properties of the oxidant (LWM thiols or reduced NOMs), the concentration of NOM [25], the NOM to Hg(0) ratio [26], or the presence of additional electron acceptors [62] or chloride ion [30].

5. Photochemical Oxidation of Hg(II) by DOM in Natural Waters

To date, there are two photooxidation pathways identified as having contributed to the observed photooxidation of Hg(0) in natural waters, i.e., Hg(0) photooxidation by hydroxyl radical (•OH) and Hg(0) photooxidation by carbonate radical (•CO3−).
In the photooxidation of Hg(0), Lalonde et al. evaluated the contribution of •OH radical to the Hg(0) oxidation in water samples from St. Lawrence estuary [36]. Upon addition of an •OH scavenger, the Hg(0) photooxidation rate decreased by 25% in brackish waters from the St. Lawrence River and 19% in artificial saline water containing p-benzoquinone. Thus, it was concluded that the observed photooxidation of Hg(0) was partially caused by •OH radicals, but not likely by O2− radicals. Oh et al. later reported that spiking the natural water samples with high levels of NO3− significantly decreased Hg(0) formation compared with the original water samples. This may be a result of Hg(0) oxidation by •OH radicals formed upon the photolysis of NO3−v [50]. The 2nd-order rate constant for Hg(0) photooxidation by •OH radicals have been determined to be 1.0–2.4 × 109 M−1 s−1 and the reaction intermediate was proposed to be •HgOH [63,64].
The photooxidation of Hg(0) by carbonate radicals (•CO3−) was identified by He et al. [65] The photooxidation rate of Hg(0) in a freshwater sample and a pH 8 buffer solution was significantly increased only when both CO32− and NO3− were present. It has been shown that •OH generated from photolysis of NO3− reacts with CO32− to produce carbonate radicals (•CO3−). Therefore, it was concluded that Hg(0) could be primarily oxidized by •CO3−. The presence of DOM greatly reduced the Hg(0) oxidation rate by •CO3− due to a possible quenching effect.
However, the photooxidation of Hg(0) by many other oxidants has not been evaluated and should be a focus of future studies.
The mechanistic scheme of the identified photoreduction and photooxidation reactions of Hg is illustrated in Figure 2.

6. Concluding Remarks and Future Research Needs

Figure 3 shows the current model of the identified dark and photo-redox pathways of Hg relevant to various aquatic environments. Up to date, dark Hg(II) reduction by NOMs remained the major identified dark Hg(II) reduction pathway relevant to the aquatic environment. Multistage slow pseudo-1st-order rate constants have been determined but the values vary widely from one study to another. The rate constants were affected by pH, concentrations of Hg(II) and NOMs, Hg(II)/NOM ratios, the structures of NOMs, Cl−, temperature, and Ca2+. Dark Hg(II) reduction by NOMs could be affected by pH but the exact mechanism is unknown. The seemingly conflicting effect of the concentrations of NOM and Hg(II) and Hg(II)/NOM ratios on Hg(II) reduction efficiency, may be due to different binding sites and electron donor groups on NOMs to Hg(II) and the re-oxidation of Hg(0). Both Hg(II)-Cl- complexation and re-oxidation of Hg(0) could contribute to the retarding effect of chloride ions on the Hg(II) reduction rate. The Hg(II) reduction rate increased with the temperature, but decreased with the concentration of Ca2+ by impeding the intramolecular transfer between Hg(II) and NOMs. Studies on the dark Hg(0) oxidation by NOM showed that the oxidation rates varied with the species of thiol-containing compounds. The oxidation of Hg(0) may occur via thiol-induced oxidation of dissolved Hg(0).
The photochemical reduction pathways of Hg(II) identified relevant to natural waters include the photoreduction of Hg(OH)2, the photoreduction of HgS22− and the photoreduction of Hg(II) by DOM. The photolysis rates of Hg(OH)2 and HgS22− may be too slow to be significant for atmospheric waters but could be important for large water bodies. Both the kinetics and possible mechanisms of the photoreduction of Hg(II) by LMWOCs with identifiable structures seem to depend on the species of LMWOCs, especially the functional groups on them. Recent studies on Hg(II) photoreduction using DOMs in various natural water samples have suggested that the DOM structure rather than the threshold in DOM concentration could significantly impact the photoreduction of Hg(II). Other factors affecting the photoreduction of Hg(II) included the light wavelength and intensity used, pH, and the presence of Cl− or dissolved O2. The major identified photooxidation pathways of Hg(0) were Hg(0) photooxidation by hydroxyl radical (•OH) and by carbonate radical (•CO3−).
Future research needs to better understand chemical redox reactions of Hg in natural waters include:
  • To reduce the uncertainties in the determination of the rate constants under a wide range of environmental conditions.
  • To study how the reaction kinetics and mechanisms varied with the functional groups on NOMs.
  • To decipher the mechanisms on how various environmental variables (e.g., Cl−) impact the chemical redox reactions of Hg.
  • To characterize NOM composition at a molecular level.
  • To identify more abiotic redox reactions of Hg that contribute to the observed field data.

Author Contributions

All the authors made significant contributions. In particular, L.S. conceived and wrote the manuscript; B.A.B. reviewed, commented and revised the paper; and J.F. contributed to Section 1, Section 2, Section 3 and Section 4. All authors have read and agreed to the published version of the manuscript.

Funding

This paper is partially supported by a grant from the Auburn University at Montgomery Research Grant-in-Aid Program (102001-220265).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Not applicable.

Acknowledgments

We would like to thank the reviewers for their valuable comments. Lin Si appreciates financial support from Auburn University at Montgomery.

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. The transformations of various Hg species in natural waters. The chemical processes circled in red are the focus of this review.
Figure 1. The transformations of various Hg species in natural waters. The chemical processes circled in red are the focus of this review.
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Figure 2. Current mechanistic scheme of photo-redox reactions of Hg relevant to natural waters.
Figure 2. Current mechanistic scheme of photo-redox reactions of Hg relevant to natural waters.
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Figure 3. Current model of the dark and photo-redox pathways of Hg under different natural water conditions.
Figure 3. Current model of the dark and photo-redox pathways of Hg under different natural water conditions.
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Si, L.; Branfireun, B.A.; Fierro, J. Chemical Oxidation and Reduction Pathways of Mercury Relevant to Natural Waters: A Review. Water 2022, 14, 1891. https://doi.org/10.3390/w14121891

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Si L, Branfireun BA, Fierro J. Chemical Oxidation and Reduction Pathways of Mercury Relevant to Natural Waters: A Review. Water. 2022; 14(12):1891. https://doi.org/10.3390/w14121891

Chicago/Turabian Style

Si, Lin, Brian A. Branfireun, and Jessica Fierro. 2022. "Chemical Oxidation and Reduction Pathways of Mercury Relevant to Natural Waters: A Review" Water 14, no. 12: 1891. https://doi.org/10.3390/w14121891

APA Style

Si, L., Branfireun, B. A., & Fierro, J. (2022). Chemical Oxidation and Reduction Pathways of Mercury Relevant to Natural Waters: A Review. Water, 14(12), 1891. https://doi.org/10.3390/w14121891

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