Fumigant methyl iodide can methylate inorganic mercury species in natural waters

Subjects

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 runoff5,6. CH3I is known to have the capability of methylating several metals and metalloids, including Sn2+, As3+, Pb2+ and Hg0, following an oxidative addition mechanism7,8,9,10. 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 observed12,13,14, 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 waters15,16, 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 CH3I10,11,17, 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 CH3199Hg+ 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.
figure1

(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 photodemethylated21,22. 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 I3amylose 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.
figure2

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.
figure3

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 (kM, 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.
figure4

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 (kM, 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+.
figure5

The obtained photomethylation rate constant of Hg2+ by CH3I (kM=1.51 × 10−2 l mol−1 E−1 m2) and photodemethylation rate constant of CH3Hg+ (kD=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, neurotoxic18,26 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 CH3201Hg 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. CH3202Hg+, and CH3199Hg+ 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 sunlight21,40. 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 kD. 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, kD, 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), kD 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), kM 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.

Additional information

How to cite this article: Yin, Y. et al. Fumigant methyl iodide can methylate inorganic mercury species in natural waters. Nat. Commun. 5:4633 doi: 10.1038/ncomms5633 (2014).

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Acknowledgements

We thank Leo Scinto for providing the instrument for PAR measurement, and Bruce McCord and Watson Lees for commenting on an earlier version of the manuscript. This research was supported by the National Natural Science Foundation of China (21120102040, 20937002), National Basic Research Program of China (2013CB430002), the U.S. Department of Energy (DE-FG01-05EW07033) and the Strategic Priority Research Program of the Chinese Academy of Sciences (XDB01020300).

Author information

Y.Y., Y.L., Y.C. and G.J. designed research; Y.Y., Y.L. and C.T. performed research; Y.Y., Y.L., Y.C. and G.J. wrote the paper.

Correspondence to Yong Cai or Guibin Jiang.

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Supplementary Information

Supplementary Figures 1-9, Supplementary Tables 1-2, Supplementary Note 1 and Supplementary References (PDF 613 kb)

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Yin, Y., Li, Y., Tai, C. et al. Fumigant methyl iodide can methylate inorganic mercury species in natural waters. Nat Commun 5, 4633 (2014). https://doi.org/10.1038/ncomms5633

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