Fumigant methyl iodide can methylate inorganic mercury species in natural waters

Abstract

Methyl iodide or iodomethane (CH₃I) has recently been registered as a fumigant in many countries, although its environmental impacts are not well understood. Here we report the results of a study on the methylation of mercury by CH₃I in natural water by incubation experiments using both Hg (¹⁹⁹HgCl₂ and )- and hydrogen (CD₃I)-stable isotope addition techniques. We find that methylation of Hg⁰, and Hg²⁺ by CH₃I can occur in natural water under sunlight, while only Hg⁰ and can be methylated in deionized water. We propose that the methylation of Hg by CH₃I in natural waters is mediated by sunlight and involves two steps, the reduction of Hg²⁺ to Hg⁰/ and the subsequent methylation of Hg⁰/ by CH₃I. Further quantitative assessment suggests that CH₃I-involved methylation of inorganic Hg could be an important source of CH₃Hg⁺ in an environment where CH₃I has been used in large amounts as a fumigant.

Introduction

Methyl iodide or iodomethane (CH₃I) is one of the most abundant organoiodine compounds in natural environments¹. Anthropogenic input of CH₃I to the environment has been rapidly increasing since the 1990s, after it was suggested as a replacement for the ozone-depleting fumigant methyl bromide (CH₃Br). CH₃I is preferred over CH₃Br because it is similar to, or more effective than, CH₃Br in controlling a wide variety of soil pests and weeds², and it has less impact on the ozone layer owing to its rapid photodegradation in the troposphere². The U.S. EPA approved the use of CH₃I as a fumigant in 2008 (ref. 3). A commercially available CH₃I formulation was then registered in many countries, such as the USA, Japan and New Zealand.

The potentially negative impact of using CH₃I has not been adequately addressed⁴, despite its increasing use on farmland. Compared with CH₃Br, CH₃I degrades more slowly in soil, hence increasing its chance of being transported to an aquatic environment via runoff^5,6. CH₃I is known to have the capability of methylating several metals and metalloids, including Sn²⁺, As³⁺, Pb²⁺ and Hg⁰, following an oxidative addition mechanism^7,8,9,10. Among these metals/metalloids, mercury deserves special attention because its methylation product, methylmercury (CH₃Hg⁺), is among the most widespread of all highly toxic contaminants. Previous work has demonstrated that CH₃I can be used as a methylation reagent for synthesizing CH₃Hg⁺ from Hg⁰ in a laboratory setting¹⁰. However, such a methylation reaction has not been confirmed in natural waters with low concentrations of Hg⁰ and CH₃I (ref. 11). The undetectable CH₃Hg⁺ in the previous study could be attributed to the possible degradation of CH₃Hg⁺, which was not monitored during the experiment¹¹. It is understandable that direct methylation of Hg²⁺ by CH₃I has not been observed^12,13,14, as this reaction is theoretically impossible according to the mechanism of the oxidative addition reaction. However, recent studies have indicated that a rapid conversion of Hg²⁺ to Hg⁰ can occur in natural waters^15,16, suggesting that Hg²⁺ could also play an important role in the CH₃I-involved methylation of inorganic mercury in the environment.

In this study, we show that CH₃I has the potential of methylating inorganic mercury in aquatic environments under sunlight. We used both mercury (¹⁹⁹HgCl₂ and )- and hydrogen (CD₃I)-stable isotopes to examine the possible methylation of inorganic mercury species by CH₃I in natural water. Various concentrations of ¹⁹⁹HgCl₂, Hg₂(NO₃)₂, Hg⁰ and CH₃I were added to deionized water or natural water and incubated with or without sunlight. , spiked in the incubation media, was used as a control to correct for the degradation of produced CH₃Hg⁺. CD₃I was used to validate the source of the methyl group in CH₃Hg⁺. Our results demonstrate that methylation of inorganic mercury by CH₃I could be an important source of CH₃Hg⁺ in CH₃I-contaminated surface water.

Results

Methylation of inorganic mercury by CH₃I in natural water

Natural water samples were collected from a pond located at the Florida International University (FIU). Methylation experiments were carried out in a closed 0.5-l Teflon bottle under natural sunlight or in darkness. The methylation of inorganic mercury species (Hg²⁺, and Hg⁰) by CH₃I was measured (Fig. 1). Significant production of CH₃Hg⁺ was observed in waters exposed to sunlight for all tested mercury species. For samples tested in the dark, no significant formation of CH₃Hg⁺ occurred for Hg²⁺ and Hg⁰ (P>0.1) (Fig. 1a,c), and a 90% reduction in the formation of CH₃Hg⁺ was observed for (P<0.05) when compared with that under sunlight (Fig. 1b). Most previous studies also showed that sunlight is necessary for the methylation of Hg⁰ by CH₃I^10,11,17, although contrast results were also observed¹⁸. In sunlight, the methylation rates of these three mercury species (r, equation (9), see Methods) were in the order of Hg⁰ (1.67±0.09 pmol l⁻¹ E⁻¹ m²)> (0.424±0.025 pmol l⁻¹ E⁻¹ m²)>Hg²⁺ (0.125±0.010 pmol l⁻¹ E⁻¹ m²). These results suggest that methylation of mercury by CH₃I can occur in natural water and this process is driven by sunlight. The 1-day methylation yield of Hg⁰ by CH₃I was calculated to be 0.6%, comparable with the results of Celo et al.^18,19 and Hall et al.¹¹ (1.1 and 0.29%, respectively). Although methylation rates of Hg⁰ and were ~13- and 4-fold higher than that of Hg²⁺, methylation of Hg²⁺ by CH₃I should be more important as Hg²⁺ is the main inorganic mercury species in natural water. The formation of CH₃ ¹⁹⁹Hg⁺ with the addition of ¹⁹⁹Hg²⁺ demonstrated that the generated CH₃Hg⁺ was from the methylation of spiked ¹⁹⁹Hg²⁺ (Supplementary Fig. 1a). Further experiments were conducted to examine the origin of the methyl group of CH₃Hg⁺ using a CD₃I tracer method. When natural water containing Hg²⁺ was spiked with CD₃I and incubated under sunlight for 4 days, formation of CD₃Hg⁺ was detected using gas chromatography–mass chromatography (GC–MS) (Supplementary Fig. 1b) (detailed description is given in Supplementary Note 1). Based on the oxidative addition theory, Hg²⁺ cannot be directly methylated by CH₃I. It is therefore proposed that the observed methylation of Hg²⁺ in natural water may proceed through a two-step reaction mechanism, the reduction of Hg²⁺ to Hg⁰ or followed by the methylation of Hg⁰ or by CH₃I (Equations (1, 2, 3, 4, 5)). Considering the electronegativity of the iodine atom, CH₃I is expected to dissociate to CH₃./I. or CH₃⁺/I⁻, but not CH₃⁻/I⁺ (ref. 20). CH₃. and CH₃⁺ can then methylate or Hg⁰. According to the density functional theory, the methylation reaction of Hg⁰ and CH₃I is thermodynamically favourable¹⁷. The requirement of sunlight for this reaction indicates the existence of a high activation barrier¹⁷. This proposed mechanism of Hg²⁺ methylation is supported by the facts that 95.2 pmol l⁻¹ Hg⁰ was detected in the pond water spiked with 4.99 nmol l⁻¹ Hg²⁺ (Supplementary Fig. 2) and that the methylation rate of Hg²⁺ was only 7.5% and 29.5% compared with that of Hg⁰ and (Fig. 1). The formation of during Hg²⁺ photoreduction in the presence of dissolved organic matter (DOM, Suwannee River humic acid) was also confirmed using Raman spectrometry (Supplementary Fig. 3). These results provide strong evidences to support the proposed two-step reaction mechanism of CH₃I-involved methylation of mercury in natural water.

Figure 1: Methylation of inorganic Hg by CH₃I in pond water.

(a) Hg²⁺, (b) and (c), Hg⁰. Solid line and dashed line represent the methylation of inorganic Hg under sunlight and in the dark, respectively. CH₃I was spiked into the pond water to form a final concentration of 1 mmol l⁻¹. The final concentrations of Hg²⁺, and Hg⁰ were 4.99, 4.99 and 4.24 nmol l⁻¹ (as Hg), respectively. The methylation rates for Hg²⁺, and Hg⁰ were calculated to be 0.125±0.010, 0.424±0.025 and 1.67±0.09 pmol l⁻¹ E⁻¹ m², respectively. All the tests were conducted in triplicate, and the error bars indicate the s.d. of three repeated measurements.

To evaluate the importance of CH₃Hg⁺ formed through CH₃I-involved process in aquatic environments, additional experiments were performed in natural waters from the Florida Everglades. In these experiments, the methylation of Hg²⁺ by CH₃I was also observed (Supplementary Fig. 4), indicating that this process may widely occur in aquatic environments. Methylation rates (r) were calculated to be 0.025±0.001 pmol l⁻¹ E⁻¹ m² and 0.125±0.010 pmol l⁻¹ E⁻¹ m² for surface water of the Florida Everglades and FIU pond water, respectively.

Methylation of inorganic mercury by CH₃I in deionized water

Experiments were conducted in deionized water, where factors (for example, DOM, Fe²⁺) possibly facilitating the reduction of Hg²⁺ are absent, to validate the aforementioned mechanism of photoinduced methylation of Hg²⁺ by CH₃I. The presence of CH₃Hg⁺ under sunlight was observed after 1 day of incubation when or Hg⁰ was used as a precursor, while the methylation of Hg²⁺ was not detected (Supplementary Fig. 5). The formation of CH₃Hg⁺ was not observed for all three inorganic mercury precursors in the dark. These results demonstrate that Hg²⁺ cannot be directly methylated by CH₃I.

In deionized water, the experimentally generated CH₃Hg⁺ became undetectable after 3 days (Supplementary Fig. 5). This loss of CH₃Hg⁺ is not probably the results of photodegradation as previous studies showed that CH₃Hg⁺ present in deionized water cannot be photodemethylated^21,22. During the course of the experiments, rapid degradation of CH₃Hg⁺ was always accompanied with a change in solution colour from clear to brown. The brown substance was identified to be I₂ based on the results of the following two experiments. A purple colour appeared in the organic phase when an extraction was performed with CH₂Cl₂ (ref. 23), indicating this substance could be I₂ (Supplementary Fig. 6). The brown solution turned blue when starch was added, indicating the formation of I₃⁻amylose complex. Furthermore, ultraviolet-visible spectra analysis showed that the brown substance in water and the purple substance in CH₂Cl₂ gave a maximum absorbance similar to that of I₃⁻ in water and I₂ in CH₂Cl₂, confirming the formation of I₂ (Supplementary Fig. 6). I₂ may be produced from the combination of I·or through the oxidation of I⁻ by dissolved oxygen (equations 6 and 7)⁵. Thus, it was assumed that I₂, originating from CH₃I, caused the degradation of CH₃Hg⁺ in deionized water. To test this hypothesis, I₂ or I⁻ solution was added to deionized water containing . For trials with I₂, immediate analysis of the solution indicated that disappeared completely (Supplementary Fig. 7a). An increase in the amount of ²⁰¹Hg⁰ and ²⁰¹Hg²⁺ was also observed, confirming the degradation of . However, the concentration of showed no significant change in the presence of I⁻ (Supplementary Fig. 7b), indicating that I⁻ could not degrade .

In contrast to deionized water, the degradation of in natural water showed no significant difference (P>0.1) in the presence or absence of CH₃I (Fig. 2). In addition, the water colour did not change to brown during the 5 days of incubation (Supplementary Fig. 6). It is proposed that any I₂ formed would be quickly reduced to I⁻ in the presence of humic substances²⁴, thus inhibiting the degradation of CH₃Hg⁺ by I₂. This assumption was partially supported by the fact that the rapid degradation of in the presence of CH₃I was not observed when Suwannee River humic acid was added to deionized water (Supplementary Fig. 8). The degradation of in Suwannee River humic acid solution correlated with cumulative photosynthetic active radiation (PAR) photon flux (Supplementary Fig. 8), indicating that sunlight was the driver for the degradation of . The rapid degradation of CH₃Hg⁺ by I₂ would explain why the methylation of Hg⁰ by CH₃I was not detected under sunlight in deionized water in a previous study¹¹.

Figure 2: Degradation of in pond water.

Solid line and dashed line represent the degradation of CH₃Hg⁺ in the presence (1 mmol l⁻¹) or absence of CH₃I, respectively. The degradation rate of was calculated to be 0.015±0.001 m² E⁻¹ in the presence of CH₃I and 0.013±0.002 m² E⁻¹ without CH₃I. The degradation of CH₃Hg⁺ showed no significant difference in the presence or absence of CH₃I (P>0.1). All the tests were conducted in triplicate, and the error bars indicate the s.d. of three repeated measurements.

Pathway of methylation of inorganic mercury by CH₃I

Based on the results obtained in deionized and natural waters, a pathway of photoinduced methylation of inorganic mercury by CH₃I in natural water was proposed (Fig. 3). Under the influence of solar irradiation, Hg²⁺ can be reduced to Hg⁰ or (probably facilitated by DOM) in natural waters²⁵, while CH₃I can be decomposed to CH₃·/I· or CH₃⁺/I⁻. CH₃· radicals and CH₃⁺ ions can then methylate or Hg⁰ through oxidative addition reactions. DOM may play an important role in this process. On one hand, DOM can be an effective scavenger of I₂, therefore inhibiting the degradation of CH₃Hg⁺ induced by I₂. On the other hand, DOM may facilitate the reduction of Hg²⁺ to Hg⁰ or under sunlight²⁵, subsequently enhancing the oxidative methyl transfer.

Figure 3: Proposed pathway of photoinduced methylation of inorganic Hg by CH₃I in natural water.

The methylation of Hg by CH₃I is proposed to include two steps, the reduction of Hg²⁺ to or Hg⁰ and the subsequent methylation of or Hg⁰ by CH₃. or CH₃⁺ (solid line), which are photo-decomposition products of CH₃I (dashed line). I₂, which is produced from the combination of I· or oxidation of I⁻ by dissolved oxygen, can rapidly degrade the generated CH₃Hg⁺. However, this process is negligible in natural waters as I₂ can be quickly reduced to I⁻ by DOM.

Assessment of CH₃I-involved CH₃Hg⁺ formation in environments

After being applied in farmland, CH₃I could be transported into surrounding ponds or lakes via runoff water. The levels of CH₃I in runoff water were estimated by laboratory column-leaching experiments (Supplementary Fig. 9). CH₃I applied to soils can be quickly leached out with concentration reaching 0.20–2.22 mmol l⁻¹ depending on the soil type used (Supplementary Fig. 9), which is two to five order of magnitude higher than its concentration in coastal waters²⁶. The reaction of Hg²⁺ and CH₃I in pond water can be described as a pseudo-second-order reaction (Fig. 4). The photomethylation rate constant of Hg²⁺ by CH₃I (k_M, see equation (13)) was then calculated to be (1.51±0.19) × 10⁻² l mol⁻¹ E⁻¹ m² in pond water. By using the obtained methylation rate constant and assuming that the increased CH₃I concentration in aquatic water was in the range of 0–7 mmol l⁻¹ (comparable to its concentration in leachates obtained in column experiments), the differences in CH₃Hg⁺ concentration at steady state in natural water with or without the presence of CH₃I were estimated (equation (17)). Figure 5 showed that the increased steady-state concentration of CH₃Hg⁺ is highly dependent on the concentration of CH₃I and Hg²⁺. If we assume that the concentration of CH₃I entering natural water is 1 mmol l⁻¹ (comparable with that obtained in leachates) and Hg²⁺ concentration is 6.1–3265.4 pmol l⁻¹ (literature reported mercury levels in natural waters)^27,28, the steady-state concentration of CH₃Hg⁺ in natural waters was estimated to increase by 0.005–2.966 pmol l⁻¹ through the CH₃I-involved methylation of mercury (see Supplementary Table 1). This increase in CH₃Hg⁺ is comparable to the concentration of CH₃Hg⁺ in these natural waters (<0.025–5.085 pmol l⁻¹)^27,28. To evaluate the relative importance of CH₃I-involved methylation versus biotic methylation in sediments, we further compared the formation rate of CH₃Hg⁺ from photomethylation by CH₃I in water body and the diffusion rate of CH₃Hg⁺ from sediment (generated by biotic methylation) with water (see Supplementary Table 2). The results showed that the calculated formation rates of CH₃Hg⁺ from photomethylation by CH₃I are in the range of 0.53–258.5 pmol m⁻² d⁻¹, comparable with the diffusion rates of CH₃Hg⁺ from sediment (−3.0–327.0 pmol m⁻² d⁻¹) (Supplementary Table 2), indicating that CH₃I-involved formation of CH₃Hg⁺ could be comparable with that diffused from sediment (mainly generated by biotic methylation). Both estimations suggest that photomethylation of mercury by CH₃I could be an important source of CH₃Hg⁺ in areas with CH₃I fumigant application.

Figure 4: Relationship between the formation rate of CH₃Hg⁺ and the product of Hg²⁺ and CH₃I concentrations in a pond water.

To calculate the relationship between the formation rate of CH₃Hg⁺ (r) and concentrations of Hg²⁺ and CH₃I, various concentration of CH₃I (0, 0.1, 0.2, 0.5, 1.0 and 2.0 mmol l⁻¹) and ¹⁹⁹HgCl₂ (0, 1.0, 2.0, 3.0, 4.0 and 5.0 nmol l⁻¹) were prepared using pond water containing 1 ng l⁻¹ (5.0 pmol l⁻¹) . The formation rate was observed to be positively related to the product of Hg²⁺ and CH₃I concentrations. Photomethylation rate constant of Hg²⁺ by CH₃I (k_M, see equation (13) in main text) was calculated to be (1.51±0.19) × 10⁻² l mol⁻¹ E⁻¹ m². All the tests were conducted in triplicate, and the error bars indicate the s.d. of three repeated measurements.

Figure 5: The effects of CH₃I and Hg²⁺ on the increased steady-state concentration of CH₃Hg⁺.

The obtained photomethylation rate constant of Hg²⁺ by CH₃I (k_M=1.51 × 10⁻² l mol⁻¹ E⁻¹ m²) and photodemethylation rate constant of CH₃Hg⁺ (k_D=0.015 m² E⁻¹) in the pond water were used to calculate the increased steady-state concentration of CH₃Hg⁺ (equation (17)). The concentrations of Hg²⁺ and CH₃I are assumed to be in the range of 0–5 nmol l⁻¹ and 0–7 mmol l⁻¹, respectively.

Discussion

Methylation of mercury by sulphate-reducing²⁹ or iron-reducing bacteria³⁰ has been deemed to be the main source of CH₃Hg⁺ in aquatic environments. However, this study shows that methylation of inorganic mercury by CH₃I could be another important source of CH₃Hg⁺. This would be particularly true in an environment where CH₃I is used in large amounts as a fumigant. In addition to acting as a methylation agent, CH₃I can also serve as a potential mobilizing agent for mercury sulfide (HgS)¹². It has been demonstrated that CH₃I could methylate sulphide into volatile (CH₃)₂S and (CH₃)₂S₂, and then liberate free Hg²⁺ into water from insoluble HgS¹². Considering HgS is an important inorganic mercury species in anoxic environments such as soil³¹, sediment³² and sulphidic water³³, this dissolution process of HgS could increase the leaching and release of Hg²⁺ from soil and sediment into surface water, and therefore enhance the potential formation of CH₃Hg⁺. As CH₃Hg⁺ is a neurotoxin, the impact of CH₃I application to farmland on biogeochemical cycling of mercury should be carefully evaluated.

With respect to the production of CH₃Hg⁺, most previous studies focused on the methylation of Hg²⁺ in the environment; however, there is a lack of knowledge on the methylation of Hg⁰. Results of this study imply that methylation of Hg⁰ can occur in the environment (for example, by CH₃I). Although Hg⁰ is present at low concentrations in water and sediment, Hg²⁺, through photomediated reduction in natural waters, provides a sufficient and continuous source of Hg⁰. Additionally, Hg⁰ is the main Hg species in the atmosphere, and the co-occurrence of Hg⁰ and released CH₃I from fumigated fields by volatilization, makes the photochemical methylation of Hg⁰ possible in the atmosphere. Therefore, the methylation of Hg⁰ by CH₃I and its contribution to the CH₃Hg⁺ pool needs to be considered.

Fumigants are widely applied in agricultural fields to control pests and weeds. With the phase out of CH₃Br, the use of CH₃I is expected to increase dramatically in the future. CH₃I is a pulmonary and dermal irritant²⁶, and is also suspected to be carcinogenic³⁴, neurotoxic^18,26 and endocrine disrupting⁴. This study shows that CH₃I could also indirectly threaten the health of humans and wildlife by forming a toxic chemical, CH₃Hg⁺, suggesting the necessity of a more comprehensive risk assessment of CH₃I use as a fumigant.

In summary, we have demonstrated that the registered fumigant CH₃I could methylate inorganic Hg species, including Hg⁰, and Hg²⁺, into toxic CH₃Hg⁺ in natural water under sunlight. The photoinduced CH₃Hg⁺ formation mechanism by CH₃I is proposed, which involves the reduction of Hg²⁺ to Hg⁰/ and the subsequent methylation of Hg⁰/ by CH₃I, following an oxidative addition reaction. Quantitative assessment suggests that methylation of inorganic Hg by CH₃I could be an important source of CH₃Hg⁺ in CH₃I-contaminated surface water.

Methods

Reagents

CH₃Hg⁺ standard was purchased from Ultra Scientific (North Kingstown, RI). Enriched ²⁰¹HgO and ¹⁹⁹HgO were purchased from Oak Ridge National Laboratory (Oak Ridge, TN). was synthesized from enriched ²⁰¹HgO using methylcobalamin³⁵. Enriched ¹⁹⁹HgCl₂ solution was prepared by dissolving ¹⁹⁹HgO in 10% HCl. Elemental mercury (Hg⁰) and Hg₂(NO₃)₂ were purchased from Fisher Chemicals (Pittsburgh, PA). Hg⁰ stock solution was freshly prepared following the method by Whalin and Mason³⁶. One drop of Hg⁰ was first transferred into a piece of silicon tubing. After capping at both ends with Teflon plugs, the tubing was immersed into deionized water. Next, the gaseous Hg⁰ was equilibrated with water for over 48 h under anoxic conditions. The concentration of Hg⁰ in water was determined before the experiment was conducted. CH₃I and CD₃I were purchased from Fisher Chemicals and Hengye Zhongyuan Chemicals (Beijing, China), respectively. Deionized water was prepared from a Barnstead NANOpure Diamond Life Science (UV/UF) ultrapure water system (Barnstead International, Dubuque, IA).

Methylation of inorganic mercury by CH₃I

Water samples were collected from a pond in FIU (25°45′N, 80°22′W) and the Florida Everglades (25°45′N, 80°44′W), and stored in pre-cleaned 2-L Teflon bottles using trace metal clean procedures. The samples were kept in a cooler and transported to the laboratory within 3 h. Methylation experiments were conducted immediately on arrival at the laboratory. The methylation experiments were carried out in a 0.5-l Teflon bottle under natural sunlight or in darkness (wrapping bottles with aluminum foil). ¹⁹⁹HgCl₂, Hg₂(NO₃)₂ or Hg⁰ was spiked into each bottle to form a final concentration of about 1,000 ng l⁻¹ (4.99 nmol l⁻¹) as mercury, respectively. CH₃I and CH₃ ²⁰¹Hg were then added to each bottle to form a final concentration of 141.9 mg l⁻¹ (1 mmol l⁻¹) and 1 ng l⁻¹ (4.975 pmol l⁻¹), separately. For the experiments conducted in deionized water, ¹⁹⁹HgCl₂, Hg₂(NO₃)₂ or Hg⁰ was spiked to each bottle to form a final concentration of about 1,000 ng l⁻¹ (4.99 nmol l⁻¹) as mercury, respectively. CH₃I were then added to each bottle to form a final concentration of 141.9 mg l⁻¹ (1 mmol l⁻¹). To investigate the effects of Hg²⁺ and CH₃I concentrations on the methylation rate, various concentration of CH₃I (0, 0.1, 0.2, 0.5, 1.0 and 2.0 mmol l⁻¹) and ¹⁹⁹HgCl₂ (0, 1.0, 2.0, 3.0, 4.0 and 5.0 nmol l⁻¹) were spiked into pond water containing 4.99 pmol l⁻¹ . Bottles were incubated under ambient temperature and light conditions for 4 days. During the experiment, the intensity of ambient PAR was measured at 15min intervals routinely using LI-192 Quantum Sensor (LI-COR Biosciences, Lincoln, NE). Triplicates (three separate bottles) were employed for each trial. CH₃ ²⁰²Hg⁺, and CH₃ ¹⁹⁹Hg⁺ in the incubated samples were determined after 0, 0.5, 1, 2, 3 and 4 days of incubation using the aqueous phenylation purge-and-trap followed by the GC inductively coupled plasma–MS (ICP-MS) detection. A CD₃I addition technique was adopted to investigate the origin of the methyl group in the generated CH₃Hg⁺. HgCl₂ and CD₃I were added to 2 l of environmental water to form a final concentration of 9.97 μmol l⁻¹ as Hg²⁺ and 20 mmol l⁻¹, respectively. The bottle was then incubated under ambient temperature and light conditions for 4 days. The formation of CD₃Hg⁺ was verified by the aqueous phenylation–solid phase microextraction (SPME)–GC–MS.

Analysis of CH₃Hg⁺

Water samples (20 ml), L-ascorbic acid (100 mmol l⁻¹, 1 ml), concentrated HCl (1 ml) and CH₂Cl₂ (5 ml) were added into a 50-ml polypropylene centrifuge tube (Corning Inc, Lowell, MA) sequentially. After shaking for 30 min, the samples were left to stand for 5 min to let CH₂Cl₂ separate from water. CH₂Cl₂ (2.5 ml) phase was then transferred to a 50-ml polypropylene centrifuge tube containing 35 ml 1% (v/v) HCl. The tube was purged with N₂ at 150 ml min⁻¹ to completely volatilize CH₂Cl₂ and release CH₃Hg⁺ into aqueous phase. CH₃Hg⁺ in the aqueous phase was analysed by aqueous phenylation and purge-and-trap pre-concentration followed by GC–ICP–MS³⁷. The artificial formation of CH₃Hg⁺ using this method has been demonstrated to be negligible, attributing to the omission of the distillation procedure and the replacement of the derivatization reagents (from sodium tetraethylborate or sodium (n-propyl)borate to sodium phenylborate (NaBPh₄))³⁷.

Aqueous phenylation reactions were performed in 300-ml glass bubblers from Exeter Scientific Glass Co. (Birdsboro, PA). The prepared sample (35 ml) was added to a 100 ml solution containing 30% (w/v) NaCl and 0.17% (w/v) NaOH. Two millilitres of citric buffer (pH 5.0, 2 mol l⁻¹) and 1 ml of 1% (w/v) NaBPh₄ (Fisher Chemicals, Fairlawn, NJ) were added. After 15 min of reaction, the phenylation products were purged from the bubbler at 45 °C for 45 min with 200 ml min⁻¹ of N₂, and trapped on a Tenax trap. The trap was dried for 5 min by 200 ml min⁻¹ of N₂. Finally, the trapped mercury species were thermally desorbed and separated using a fused silica capillary column (15 m × 0.53 mm i.d., coated with DB-1 film of 1.5-μm thickness) and then measured with an Elan DRC-e ICP-MS (PerkinElmer Sciex, Waltham, MA).

Phenylation–SPME–GC–MS³⁸ was used to validate the origin of the methyl group in produced CH₃Hg⁺. Natural waters (2,000 ml) were transferred to 2-l Teflon bottles and spiked with 4 mg Hg²⁺ and 40 mmol CD₃I. The solution was incubated under ambient temperature and sunlight for 4 days. The produced CD₃Hg⁺ was extracted from each 40-ml water sample after adding 1 ml concentrated HCl and 5 ml CH₂Cl₂. The CH₂Cl₂ phase (4 ml) was transferred into a 100-ml polypropylene centrifuge tube containing 35 ml 1% HCl and purged with N₂ at 150 ml min⁻¹ to remove CH₂Cl₂. Then, pH was adjusted to 7 using 0.17% (w/v) NaOH. Sample solution (35 ml), 2 mol l⁻¹ HAc-NaAc buffer (5 ml, pH 4.9) and 15 g NaCl were added to a 100-ml glass vial sealed with a polytetrafluoroethylene-coated septum. One millilitre 1% (w/v) NaBPh₄ solution was injected into the vial using a syringe. The vial was stirred by a magnetic stirring bar and maintained at 55 °C for 30 min in an oven. Next, headspace SPME extraction was carried out at 55 °C for 10 min. After extraction, the fibre was introduced into the GC injection port (200 °C) for thermal desorption and GC–MS analysis in splitless mode. The GC temperature programming is as following: 80 °C for l min, increased to 280 °C at a rate of 8 °C min⁻¹ and held there for 10 min.

Analysis of Hg⁰

Hg⁰ in water was purged and trapped onto a home-made gold trap and detected by atomic fluorescence spectrometer. Hg²⁺ solution was used as the calibration standard. Hg²⁺ standard and 0.5 ml SnCl₂ solution (10% (w/v)) were added to 100 ml 1% (v/v) HCl. The mixture was left to react for 15 min and then purged for 20 min with 400 ml min⁻¹ of N₂ at room temperature to let Hg⁰ completely volatilize and adsorb on the gold trap. The trap was dried for 5 min by 400 ml min⁻¹ of N₂. Finally, the trapped Hg⁰ was thermally desorbed and determined by a PS Analytical mercury speciation system (Orpington, Kent, UK). The analytical procedure for purgeable Hg⁰ in water samples was similar to that for the Hg²⁺ standards, except that SnCl₂ and HCl were not added.

Characterization of using Raman spectrometry

Formation of during the photoreduction of Hg²⁺ was investigated using surface-enhanced Raman spectrometry. Suwannee River humic acid and Hg²⁺ were added into deionized water to form a final concentration of 40 mg l⁻¹ dissolved organic carbon and 0.5 mmol l⁻¹ Hg²⁺, respectively. After incubation under sunlight for 3 days, Raman-scattering spectra of in the solution with and without the addition of KI (1.0 mol l⁻¹) were collected using the following method. Au nanoparticle (AuNPs) solution was prepared by heating the solution containing 1 mol l⁻¹ AuCl₄⁻ and 5 mg l⁻¹ dissolved organic matter (pH 6.8) at 80 °C for 24 h³⁹. AuNP-coated Si wafer was then made by dropping AuNPs on Si wafer and then dried in a vacuum desiccator at room temperature. Next, water samples containing mercury species were dropped on the wafer and dried in a vacuum desiccator at room temperature. The AuNPs-enhanced Raman-scattering spectra of samples were obtained by using a Leica microscope in a confocal Raman spectroscopy system (Renishaw InVia Raman microscope, Britain). Spectra collections were carried out using a × 50 objective lens and 60-s scan between 155.48 and 1,132.18 cm⁻¹ with a 532-nm laser (5 mW) and 2,400 lines per mm grating.

Column experiments for evaluating the leachability of CH₃I

The leachability of CH₃I was investigated using columns (16 cm length × 5 cm i.d.) containing vegetable soil, sandy soil or quartz sand. To minimize the losses of soil during the experiment, a well-permeable quartz plate was placed at the bottom of the column and the top 3 cm of the columns was filled with quartz sand. The columns were pre-saturated with deionized water and then 27.36 mg CH₃I (corresponding to an application rate of 125 lb ac⁻¹) was added into the columns. The columns were then leached with deionized water at a rate of 4.1 ml min⁻¹ for 3 h. Leachates (50 ml) were continuously collected at an interval of ~12 min.

Leachates (10 ml) in 15-ml polypropylene centrifuge tube were extracted with CH₂Cl₂ (2 ml) (Corning Inc, Lowell, MA). After shaking for 30 min, the samples were left to stand for 5 min to let CH₂Cl₂ separate from water. One ml CH₂Cl₂ phase was collected for GC–MS analysis.

The determination of CH₃I was performed using an Agilent 7,890 gas chromatograph system, coupled to a quadruple Agilent 5,975 electron ionization mass spectrometric detector (Agilent Technologies, Palo Alto, CA, USA) equipped with a DB-624 fused silica capillary column (30 m × 0.25 mm i.d., 1.4-μm film thickness). The gas chromatograph system was equipped with a split/splitless injection port operating in split mode (50:1) at 120 °C. The injection volume was 2 μl. The oven temperature was programmed from 40 °C (2 min) to 90 °C at 5 °C min⁻¹, and then to 240 °C at 15 °C min⁻¹. The carrier gas was helium with a constant flow of 1 ml min⁻¹. The mass spectrometer was operated in SIM mode (m/z=127 and 142). Solvent delay was set at 2 min. Quantification was conducted using an external standard curve method.

Data analysis

The methylation rates of Hg⁰, and Hg²⁺ (r) were calculated by nonlinear regression of CH₃Hg⁺ concentration against t using equation (9) (Origin, Version 6.0 for Windows, OriginLab Corp., Northampton, MA). The net production rate of CH₃Hg⁺ can be described as the difference of the photomethylation rate of inorganic mercury by CH₃I (r × PAR) and the photodegradation rate of CH₃Hg⁺ () at t time (equation (8)). Next, the function of at t time could be obtained (equation (9)) by integrating equation (8). Degradation of CH₃Hg⁺ can occur in natural waters and this process was deemed to be driven by sunlight^21,40. In this study, the spiked was also observed to be rapidly degraded under sunlight in both tested waters, while degradation of was not observed in the dark. The effect of CH₃Hg⁺ photodegradation was taken into account in the calculation, represented by the photodegradation rate constant k_D. A model based on the first-order chemical kinetics was used to describe the photodegradation of in water (equation (10))^21,40. The rate constant of photodegradation, k_D, was calculated by linear regression of against t.

where and are the concentrations of CH₃Hg⁺ at the beginning and t time, respectively (pmol l⁻¹), k_D is the rate constant of CH₃Hg⁺ photodegradation (m² E⁻¹), PAR is the photosynthetically active radiation (E m⁻² d⁻¹), J is the cumulative PAR photon flux (E m⁻²) and r is the PAR normalized formation rate of CH₃Hg⁺ (pmol l⁻¹ E⁻¹ m²).

The differences in CH₃Hg⁺ concentrations at steady state with or without the presence of CH₃I were calculated and used to evaluate the influence of CH₃I on CH₃Hg⁺ production. In natural waters, the concentration of CH₃Hg⁺ was assumed to be determined by several key processes, including photodegradation of CH₃Hg⁺, input of CH₃Hg⁺ from sediment and runoff, methylation in water body and methylation by CH₃I. Next, the changing rate of CH₃Hg⁺ could be calculated by the photodegradation rate , the CH₃Hg⁺ input rate from sediment and runoff (R₀), the methylation rate (R′) and the rate of photomethylation by CH₃I (r × PAR). As r was found to be positively related to the product of and (equation (13), Fig. 4) and equation (14) could be obtained by combining equation (13) with equation (12).

where R₀ is the rate of CH₃Hg⁺ input from sediment and runoff (pmol l⁻¹ d⁻¹), R′ is the non-CH₃I involved methylation rate in water body (pmol l⁻¹ d⁻¹), k_M is the PAR-normalized formation rate constant of CH₃Hg⁺ (l pmol⁻¹ E⁻¹ m²), is the concentration of Hg²⁺ (pmol l⁻¹), is the concentration of CH₃I (pmol l⁻¹). The predicted concentration of CH₃Hg⁺ at steady state in the presence of CH₃I () can be calculated by making the right side of equation (14) equal to zero (equation (15)).

The predicted concentration of CH₃Hg⁺ at steady state in the absence of CH₃I () can be calculated by making of equation (15) equal to zero (equation (16)).

Next, the differences in CH₃Hg⁺ concentrations at steady state with or without the presence of CH₃I () can be calculated by equation (17).

The parameters for calculating are given in Supplementary Table 1.

To evaluate the relative importance of CH₃I-involved methylation versus biotic methylation, the CH₃Hg⁺ photomethylation rate within the entire water column was calculated as follows. The sunlight intensity at Z depth (m) is calculated from the light intensity in the surface water (PAR (0)) by exponentially decreasing it with depth according to the Beer-Lambert equation. Combined with the calculation of the photomethylation rate of CH₃Hg⁺ in the surface water, the photomethylation rate at Z depth can be described as equation (18). Next, mercury methylation is integrated with respect to water column depth to obtain the methylation rate within the entire water column (equation (19)).

where is the concentration of CH₃Hg⁺ at depth Z (pmol l⁻¹), PAR(0) is the PAR above the surface of the water (E m⁻² d⁻¹) and k is the light attenuation coefficient of sunlight (m⁻¹).

Then,

The parameters for calculating are given in Supplementary Table 2.