Trace Materials in Air, Soil, and Water : Mercury-Thiourea Complex Ion

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Chapter 6

Mercury-Thiourea Complex Ion Chromatography: Advances in System Chemistry and Applications to Environmental Mercury Speciation Analysis Todd A. Olsen,1,4 Tina H. Huang,2 Ramdas Kanissery,3 and Robert J. M. Hudson*,3 1Department

of Civil and Environmental Engineering, University of Illinois at Urbana-Champaign, Urbana, Illinois 61801 2Department of Chemistry, University of Illinois at Urbana-Champaign, 367B Noyes Lab, 601 S. Mathews Avenue, Urbana, Illinois 61801 3Department of Natural Resources and Environmental Sciences, University of Illinois at Urbana-Champaign, W-503 Turner Hall, 1102 S. Goodwin Avenue, Urbana, Illinois 61801 4Current address: Division of Environmental Sciences, Oak Ridge National Laboratory, P.O. BOX 2008 MS 6036, Oak Ridge, Tennessee 37931 *E-mail: [email protected].

Mercury-thiourea complex ion chromatography is the core of a relatively new approach to Hg speciation analysis that is sensitive enough to accurately quantitate monomethyl Hg at ultratrace environmental levels. In this chapter, a detailed description of an updated system chemistry and operating conditions for performing mercury speciation analysis using inductively coupled plasma mass spectrometry are presented. The new operating conditions are very stable, highly-sensitive, and free of problems that afflicted earlier versions of the system. The potential for obtaining accurate results using the system is demonstrated by the excellent performance metrics obtained and the close agreement of these results with consensus values of monomethyl Hg in water and sediment reference materials and inter-lab comparison samples. The ability to quantitate © 2015 American Chemical Society Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

mercuric and monoethyl along with monomethyl forms of Hg is a feature of this approach that likely will prove useful in future studies of environmental systems.


Environmental Hg Speciation Despite the analytical challenges of accurately measuring multiple chemical forms of mercury at ultratrace levels, environmental chemists, hydrologists, and oceanographers have made much progress in understanding the distribution and biogeochemical cycling of this element in aquatic ecosystems over the past 25-30 years. The central fact established by their work is that the levels of Hg are very low, except in systems receiving direct discharges of Hg pollution. In most of the atmosphere, the partial pressure of elemental mercury is ~10–13 atm (1). In marine and fresh surface waters, dissolved Hg occurs at nanogram per liter (picomolar) levels or less (2). In sediments and soils, Hg levels are roughly 1000-fold higher on a mass basis but still fall in the tens of nanograms per gram range (3). In biota, levels commonly reach into the hundreds of nanograms per gram, making tissues the easiest to analyze of the main environmental media (4). The fact that Hg naturally occurs at such low levels not only makes chemical analysis more difficult, but it explains why humans have been able to significantly perturb Hg levels in environmental systems ranging from local to global in scale (5). Environmental chemists have also shown that most Hg in the environment occurs in a few distinct chemical forms – elemental, mercuric, monomethyl and dimethyl – with the proportions of each varying between environmental compartments (6, 7). Their investigations of Hg speciation have employed a wide variety of analytical methods in different environmental media. Our goal in this chapter is not to describe them all but to report on the potential of one new approach to Hg analysis to expand our understanding of environmental Hg speciation in aquatic ecosystems. To see this potential clearly, it is helpful to appreciate what has and has not been learned by employing the most widely-used methods. Major Forms of Dissolved Hg Environmental chemists seeking to fully characterize the speciation of dissolved mercury in natural waters generally divide the total (THg) among four major chemical forms: elemental (Hg0), mercuric (HgII), monomethyl (MMHg), and dimethyl (DMHg) (Figure 1). Using widely-known analytical methods, ambient concentrations of all forms except HgII are directly measureable. Monoethyl mercury (ETHg) has been reported only rarely and is generally neglected. Unlike most metals, Hg is “atmophilic.” As a result, not only do volatile forms of Hg occur in the atmosphere, but they also dissolve into natural waters. In freshwater systems, the only dissolved gaseous species typically present is Hg0. Thus, it can be measured as dissolved gaseous mercury (DGM) without performing the chromatographic step needed to distinguish DMHg and Hg0 in 116 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


marine waters. However, since Hg0 is not directly relevant to assessing the toxicity of Hg pollution, DGM is generally not measured in freshwater samples. Furthermore, since Hg0 oxidizes to HgII during sample storage, measurements of either one in such samples cannot be assumed to reflect ambient conditions. A further complication in Hg speciation analysis arises from the fact that measurements of “dissolved” species in natural waters are by necessity operationally-defined, i.e., by filtration using a filter with pores of diameter 0.4-µm or some similar size. Thus, it comes as no surprise that some chemists also report detecting nanoparticulate Hg capable of passing through these pores (Figure 1). Methods of quantifying HgNP involve isolation of the fine particulate fraction by ultrafiltration or solid-phase extraction, since at least some of the nanoparticles are hydrophobic. It should be noted that this Hg fraction is not often measured but can be a significant part of the total “dissolved” pool (8, 9).

Figure 1. Chemical forms of dissolved mercury in natural waters: nanoparticulate (HgNP), mercuric (HgII), elemental (Hg0), dimethyl (DMHg), monomethyl (MMHg), and monoethyl (ETHg). Aggregate analytes: Total (THg) and Dissolved Gaseous Mercury (DGM). Inferred: Inorganic (InHg). Directly-measurable forms in red italics.

The complete speciation scheme of Figure 1 is impractical for widespread use in environmental monitoring. Instead, the current practice in environmental assessment work is to monitor both MMHg, due to its great toxicological relevance, and THg. Datasets comprising these measures often derive a concentration of “inorganic Hg” (InHg), which includes HgII plus HgNP and any Hg0 that was originally present in the sample, from the difference between THg and MMHg:

Note that this nomenclature (Figure 1; Table 1) follows widely-adopted conventions of environmental mercury chemists. These parameters refer to measurable quantities of different analytes, which in several cases comprise groups of species, e.g., all species containing the element Hg or all complexes of monomethyl mercury. The same notation is also used when referring to the analytical method. 117 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 1. Nomenclature Used in Defining Forms of Hg Quantifiable by Analysis as Well as Distinct Chemical Species


Analytical parameters THg

Total mercury (all species of Hg)

HgII

Mercuric mercury (except nanoparticulate)

HgI

Mercurous mercury

MMHg

Monomethyl mercury (all complexes of MeHg+)

DMHg

Dimethyl mercury (Me2Hg)

ETHg

Monoethyl mercury (all complexes of EtHg+)

InHg

Inorganic Hg (all except MMHg, DMHg, and ETHg)

DGM

Dissolved gaseous mercury (Me2Hg and Hg0)

HgNP

Nanoparticulate Hg (including HgS(s))

Distinct species (complexes not shown) Hg0

Elemental mercury

Hg2+

Mercuric ion

MeHg+

Monomethyl mercuric ion

Me2Hg

Dimethyl mercury

EtHg+

Monoethyl mercuric ion

MeHgEt

Methylethyl mercury

Hg22+

Mercurous ion

Hg-binding ligands considered OH–

Hydroxide ion

Cl–

Chloride ion

Br–

Bromide ion

SH–

Bisulfide ion

TU

Thiourea

HSRAA

Thiol-containing amino acids

HSRDOM

Thiol moieties of natural dissolved organic matter

These symbols and acronyms, however, are not necessarily the clearest for describing individual ionic or complex species and conflict with other naming conventions, such as “Me” for methyl. Thus, herein we employ a second convention – Hg2+, MeHg+, and EtHg+ – when referring to distinct species or the common central ion of a group of complexes. Complexes of these ions with 118 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

relevant ligands (Table 1) are written as normal central ion-ligand combinations. Finally, we also refer to groups of complexes with the same central metal ion and ligand, but variable ligand numbers as , where zL indicates the appropriate net charge of the complexes containing metal ion M and x molecules of ligand L.


Complexation of Dissolved Hg One central goal of environmental chemists is to characterize the speciation of each “truly dissolved” form of Hg (Figure 2). While knowing the concentrations of dissolved HgII and MMHg is enough to quantify their respective transport fluxes, knowing the concentrations of the aquo Hg2+ and MeHg+ ions and their coordination complexes in different environmental compartments is crucial to predicting their reactivities (10). The distribution of Hg2+ and MeHg+ among their various weak and strong complexes affects their bioavailabilities as well. The former include complexes with hydroxide (OH–) and chloride (Cl–) anions while the latter include complexes with bisulfide ion (SH–) and the thiol moieties of amino acids (HSRAA) and natural dissolved organic matter or DOM (HSRDOM).

Figure 2. Important complexes of Hg2+ and MeHg+ included in HgII and MMHg, along with Hg in nanoparticles. Arrows represent relatively rapid (solid) and slow (dashed) reactions. Free ligands not depicted. (see color insert) 119 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


Methods for characterizing the organic complexation of Hg2+ are not as developed as they are for other environmentally-significant divalent metal ions, such as Cu2+. The few measurements that exist suggest that this ion, as well as MeHg+, are strongly bound (11–13) and it is widely accepted that this is more often the case than not (14). Note that essentially all measurements of Hg2+ complexation by natural ligands are limited by the operational nature of the measures of HgII employed. Those using THg measurements include both MMHg and HgNP in HgII when they should not, and those using SnCl2-reducible Hg may underestimate HgII. In fact, strong complexation by natural organic matter must also routinely be overcome in the sample preparations used in quantifying dissolved THg and MMHg in natural waters.

Major Forms of Hg in Sediments Saturated sediments are especially important compartments within aquatic ecosystems for Hg biogeochemistry because the anaerobic conditions that commonly form within them support the methylation of mercury (15). As these conditions typically do not develop in surface waters, sediments are usually the main source of MeHg+ to aquatic ecosystems. Thus, measuring the concentrations of the substrate for the methylation process – HgII – and the product – MMHg – both in bulk sediment samples and porewater has proven to be important in environmental work. Now many of the same Hg species and complexation reactions occur in sediment porewater as in surface waters, although adsorption reactions take on much greater significance. Adsorption is essentially a class of metal complexation reactions by the ligands located on the surfaces of particles (16). Thus, adsorption of both MeHg+ and Hg2+ by thiols in sediment organic matter and by sulfide minerals play a very significant role (10, 17). For Hg2+, formation of pure mineral phases, including cinnabar (α-HgS) and metacinnabar (β-HgS), can also limit porewater concentrations of HgII (18–20). While the use of direct solid phase speciation methods for Hg is increasing (21), much analysis of aquatic sediments still relies on chemical digestions to extract the analytes from the complex sample matrices (22). The initial stages of these sediment sample preparations usually differ markedly from those for water; by the final steps where the different forms of Hg are quantified, the methods are often related. In addition, one important consideration in designing methods is similar: how can any particular Hg species (the analyte) be extracted from an environmental matrix that strongly binds it? Thus, in this chapter, we also consider the analysis of MMHg in sediments.

Hg Speciation Analysis The understanding of Hg speciation in the environment described above was made possible by methodological advances in several different areas, including: i) sample collection and handling, ii) sample preparation, iii) analyte 120 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


preconcentration, iv) chromatographic separation of Hg species, and v) sensitive detection. All of these advances have been essential for obtaining relevant data. Here, however, we focus on the subject areas ii) – iv) since they are most closely related to the analysis of Hg speciation in aqueous sample preparations. Perhaps surprisingly, the bulk of the key data on Hg in environmental media was obtained using just two core analytical methods. The first method entails measuring HgII as Hg0 following a reduction process and preconcentration via gold amalgamation (23). The second method quantifies MMHg, usually by aqueous ethylation followed by purge and trap of the derivatized MeHg+ with gas chromatographic separation from other volatile Hg species (24). The ability of these methods to measure Hg in different environmental media – air, water, and solids – depends on coupling them with appropriate methods of preparing samples for analysis, preconcentrating analytes, and on having sufficiently sensitive instruments for detecting the Hg. Measurement of Hg0 is the key quantitation step in several widely used methods of analyzing total Hg, labile HgII species, and Hg0 itself. Total Hg is commonly measured by reducing to Hg0 the products of exhaustive oxidation of Hg in environmental samples (23). Such methods can be applied to different physical fractions of water or sediment samples, but cannot make chemical distinctions between the main forms of Hg present within the samples or fractions themselves. In addition, operationally-defined, labile HgII fractions have been measured by adding a reductant such as SnCl2 directly to water or sediment samples (12). Finally, sparging samples and trapping Hg0 is a direct speciation method used in studies where gaseous forms of Hg need to be quantified (7). The distillation/ethylation gas chromotagraphy (GC) method (25), which is the only widely-used method for assaying a particular ionic mercury species at ambient environmental levels, includes multiple steps (Figure 3A). As implied by the schematic of MeHg+ complexation equilibria in natural waters (Figure 2), it is necessary to first release the ion from strong organic complexes in the sample matrix before the preconcentration, chromatographic, and detection steps in analysis. Since accurate quantitation requires extremely high instrumental sensitivies to measure Hg species at the sub-parts per trillion levels found in the environment, both of the above methods usually quantify gaseous forms of Hg using atomic fluorescence spectrometry (AFS) or inductively coupled plasma mass spectrometry (ICP-MS). The main steps in using the distillation/ethylation-GC method to analyze dissolved MMHg closely parallel those in the alternative methodology described herein (Figure 3B). Both first steps, distillation and thiourea-catalyzed solid-phase extraction (TU-SPE), effectively remove the analyte(s) from the sample matrix. TU-SPE also affords a substantial factor of preconcentration that is deferred to the second step in the standard method. Both second steps concentrate and trap the analyte(s) for injection into the chromatographic system. Both third steps, gas or ion chromatography (IC), separate the analyte(s) from other Hg species. Finally, both detection steps quantify Hg in a cold vapor stream using the same detectors. The mercury-thiourea complex ion chromatography (HgTU-IC) system described below is used during the second through fourth steps in Figure 3. 121 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


Figure 3. Main steps in dissolved MMHg analysis by A) distillation/ethylation-GC and B) thiourea-catalyzed solid-phase extraction/Hg thiourea complex-ion chromatography.

The differences between the chemistries of distillation/ethylation-GC and TU-SPE/HgTU-IC are evident from the ligand substitution reactions of aqueous Hg2+ amd MeHg+ complexes, both present in the natural sample and formed during the process of analysis. By manipulating the ligands coordinating each Hg species, the Hg2+ and MeHg+ can be i) removed from the original sample matrix, ii) pre-concentrated, and iii) separated or derivitized prior to detection. The reactions occurring in the sample matrix can be conceptualized as the competition for Hg between a strong natural ligand, such as a thiol moiety in dissolved organic matter (HSRDOM), and the ligands (AL) introduced in order to aid in analysis. The reactions for Hg2+ and MeHg+ with a monodentate analytical ligand can be written as:

Depending on the relative Bronsted basicities of AL and , the favored equilibrium state of the Hg species in these reactions may have significant pH dependence, making the manipulation and control of this variable central to the analytical process. 122 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

MMHg Analysis by Distillation/Ethylation-GC


The standard method of measuring MMHg in natural water samples, as described in United States Environmental Protection Agency (USEPA) Method 1630 (25), involves an elaborate sample preparation before the GC-based speciation and Hg detection steps. The most common sample preparation entails distillation of an acidified sample, which releases MeHg+ from the strong Hg-binding sites in dissolved organic matter (DOM) that can cause the sample matrix to interfere with direct ethylation. In DOM-rich samples, recoveries of MeHg+ internal standards can be 80% or lower with direct ethylation (26). The key reactions in water vapor distillation are ligand exchange with the analytical ligand, typically a monovalent anion such as chloride or bromide:

followed by bubbling with N2 to effect phase transfer:

and then trapping by dissociation in the condensed distillate:

It should also be noted that under the conditions in the distillation vessel, the predominant halide complexes of Hg2+ – [Hg(AL)3]− and [Hg(AL)4]2− – are themselves anionic. Thus, the distillation acts as a quite effective, if incomplete, isolation step for MeHg+. After buffering the distillate near pH 4.5, sodium tetraethyl borate is added to generate the volatile methyl ethyl mercury species (MeHgEt) from MeHg+, while any Hg2+ in the distillate reacts to form diethyl mercury (Et2Hg):

The resulting volatile alkylmercurials are purged from the ethylation vessel with N2 and preconcentrated onto a Carbotrap. Later, they are desorbed into an Ar gas stream by heating the trap and the MeHgEt separated from Et2Hg by GC. Post separation, the volatile Hg species are detected either by ICP-MS or by AFS, after pyrolysis to Hg0. Distillation/ethylation-GC is the standard method for determination of MMHg in environmental work due to its high sensitivity and low method detection limit of 0.01-0.02 ng/L for water samples (25). Note that while the distillation step helps to overcome interference by DOM, it is tedious, time-consuming, and limits the sample size that can be easily processed to ~45-60 mL (25). This keeps the detection limit higher than is desirable for marine and large lake ecosystems, although a detection limit of 0.003 ng L–1 has been achieved with speciated isotope dilution (27). 123 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Recognizing this challenge, numerous researchers have looked and are searching for an alternative way to isolate and preconcentrate dissolved MeHg+. A major focus of work on Hg speciation analysis in this laboratory since 2005 has been to develop and refine a new method of preconcentrating MeHg+ from natural waters so that large volumes can be processed and detection limits further decreased.


Speciation Analysis by Thiourea-Catalyzed Solid-Phase Extraction/HgThiourea Complex-Ion Chromatography In contrast to the GC-based analytical processes, which employ irreversible alkylation reactions to form distinct, volatile Hg species prior to preconcentration and chromatographic separation, in HgTU-IC both Hg2+ and MeHg+ are reduced to volatile Hg0 after preconcentration and chromatographic separation. Central to this process is the reversible manipulation of the ligands coordinating the Hg species. In these reactions, the key analytical ligand is TU, which by virtue of its own neutral charge, forms complexes with Hg2+ and MeHg+ (Figure 4) whose respective charges reflect the central metal ion and thus can be separated on the basis of this difference (28).

Figure 4. Structures of Hg(TU)22+ and MeHgTU+complexes.

Thiourea-Catalyzed Solid-Phase Extraction The TU-SPE process entails a series of ligand exchange reactions of Hg2+ and MeHg+ complexes with TU, the key analytical ligand. For example, to free either Hg species from the strong Hg-binding sites in DOM, water samples are acidified and amended with TU, to create conditions that favor dissociation:

Pre-concentration is then affected in an off-line preparatory step by adsorption onto thiol-functionalized resins (HSRResin) at pH ~ 4.0:

124 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

In these reactions, TU is not consumed, but its presence facilitates the quantitative adsorption of Hg species onto the resin, presumably because the ligand exchange reactions of TU complexes are more rapid. The sorbed Hg species are then eluted from the resin using strongly acidic (pH ~ 0) TU solutions (see system eluent below). This preconcentration process was therefore named “thiourea-catalyzed solid phase extraction (29).”


pH-Modulated Thiol-Thiourea Switch The pH dependence of the reactions comes about because the pKa of TU is less than zero, while that of a thiol is typically about 9, making it a simple matter to manipulate the relative affinities of the Hg species for the resin and the aqueous phase by altering the pH of the solution. This pH-modulated switch (Figure 5) was first employed by Shade (28) in order to permit loading of Hg from sample preparations onto the on-line thiol trap of the HgTU-IC system.

Figure 5. Thiol-thiourea switch: A) OFF (pH < 2): Thiols fully protonated and MeHg+ is complexed by TU in solution. B) ON (pH > 3) Partially-deprotonated thiols on resin outcompete TU for MeHg+.

The ligand exchange reactions of the on-line concentrator are exactly analogous to reactions (10) and (11) of the off-line TU-SPE process, with analyte loaded at pH 4 and then eluted into the system at the pH of the eluent, i.e., the 0 to 1 range. 125 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Ion Chromatographic Separation


In HgTU-IC, transport of complexes is retarded more than the MeHgTU+ complex due its stronger tendency to adsorb to the sulfonate moieties . Of course, when contacting an aqueous solution, these sites on the resin on the resin bed are matched by counter ions. In the HgTU-IC system, H+ is the only cation added in more than incidental amounts. Thus, the adsorption of Hg species can be written as one of the following ion exchange reactions with H+:

As a result, the retention of Hg complexes on the column must be pH-dependent. By optimizing the pH of the eluent, complete (baseline) separation between the leading MeHg+ and trailing Hg2+ peaks can be achieved. Post-Column Chemistry and Hg Detection To obtain high-sensitivity detection, the separated Hg species are transferred from the mobile phase to an Ar carrier gas stream using a multi-step, on-line reaction process. First, the difficult-to-reduce MeHg+ is oxidized to Hg2+ by a post-column reaction:

The same oxidation step also breaks down the TU, which otherwise would inhibit Hg2+ reduction. After quenching excess oxidant and neutralizing the acid, the Hg2+ is reduced to Hg0 under alkaline conditions before the sample stream is fed into a gas-liquid separator (GLS), where Hg0 is stripped into the gas phase and the cold Hg vapor carried to the detector:

The cold vapor generated by the HgTU-IC system is compatible with detection by atomic fluorescence after drying the carrier gas stream (28) or directly by ICPMS as shown here (Figure 6).

Key Changes in the HgTU-IC System The fundamental chemistry and procedures for using the HgTU-IC methodology have been published in papers by this group (28, 29) and by Shade (30). The methodology has been adopted in toto in a few instances (31) and other groups (32) have adopted major aspects of the HgTU-IC chemistry. Several analysts have made use of the pH-modulated thiol-thiourea switch to elute MeHg+ from thiol resins embedded in diffusive gradients thin film (DGT) gels (33–35). 126 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

In this lab, problems with the original post-column chemistry arose in 2009 and were not fully resolved until recently. The remainder of this chapter describes work that we have done to transform the system into a reliable and accurate tool for Hg speciation analysis and deepen our understanding of its operation. The key changes from the original system can be summarized as: • •


• • •

Substitution of a commercially-available thiol resin for the customsynthesized resin used originally, Adjustment of [H+] in the mobile phase in order to better separate species of low charge, including MeHg+ and EtHg+, Substitution of KBrO3 for H2O2/UV irradiation in the oxidation step (30), Substitution of NaBH4 for alkaline Sn2+ in order to avoid formation of particles that caused perturbations in the baseline, Coupling to ICP-MS detection to permit speciated isotope dilution.

A complete description of the system configuration, its operating conditions, and essential background chemistry are presented below.

System Components and Reagents Upstream of the detector, the current system components and configuration (Figure 6) are similar but not identical to those employed in earlier work in this lab (28, 29) and at Quicksilver Scientific (30). The physical parts of the system include (1) an isocratic high pressure liquid chromatography (HPLC) pump (Chromtech Series III), (2) a 2-position, 10-port sample injection valve (Rheodyne), (3) a 4×50-mm ion chromatography guard column (Dionex CG-5A), (4) an oxidation loop, (5) an oxidant-quenching loop, (6) an acid-neutralization loop, (7) a Hg reduction loop, (8) a custom borosilicate glass, gas-liquid separator (GLS) (Allen Glass, Boulder, CO), and (9) a Hg detector (Agilent 7500S ICP-MS). Peristaltic pumps are used for loading a) samples (LP) and b) reagents (RP) and for c) draining the waste solution from the GLS (WP) (Figure 6). The custom medium-pressure, concentrator/thiol trap (TT) and stock sample injection loop (SL) are both located on the injection valve. Because the acid content of the eluent ([HClEL]) was relatively high, all mechanical components of the system have an all-PEEK flow path and all tubing is made of PEEK (high pressure end of the system) or Teflon PFA (low-pressure end). Aqueous solutions containing analytes are injected into the high-pressure end of the system either via the SL or TT mounted on the injection valve (Point 2 in Figure 6). Separation of the species occurs on the ion chromatography column (ICC; Point 3 in Figure 6), which contains a mixed resin designed for analyzing trace metals. The first post-column step in the on-line reaction system is oxidation of TU by bromine monochloride (BrCl) in the oxidation loop (Point 4 in Figure 6). ) with the HCl in the The BrCl is formed upon mixing of the KBrO3 solution ( post-column acid (HCl) stream and the mobile phase. The BrCl also demethylates 127 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


MeHg+ to Hg2+. In the antioxidant loop (Point 5 in Figure 6), excess oxidants are quenched by the sodium ascorbate solution (Asc), while any hydrophobic oxidation byproducts, mostly S80, are kept in solution by the Triton XTM in the reagent. Next (Point 6 in Figure 6), the acids in the sample stream are consumed by introducing a strong base (KOH) with the resultant heat being removed as the neutralization loop passes through an ice bath. The final reaction step (Point 7 in Figure 6) is the reduction of Hg2+ to Hg0 after mixing with the alkaline borohydride ). The mobile phase, now much altered from the original eluent, solution ( passes through a gas liquid separator (Point 8 in Figure 6), where the Hg0 is stripped into the argon stream and carried to the detector (Point 9 in Figure 6).

Figure 6. Configuration of the current generation HgTU-IC system. See Table 2 for reagent compositions. (see color insert)

The particular set of post-column reactants used here (Table 2) have been extensively optimized and found to yield stable baselines and consistent sensitivities day to day. As will be discussed in more detail below, different eluents should be used in measuring MMHg and HgII, with the eluent for the latter analyte remaining unchanged from the original recipe.

Reagent Cleanup In order to maintain low blanks and chromatographic baselines, it is essential to employ reagents that are as free of Hg as is practicable. In general, we find that reagent grade chemicals are adequate. As the system reagents make substantial use of HCl, we generally purchase the trace metal grade of this acid (Fisher). However, when batches of HCl have problematical Hg blanks, we clean up 2 M solutions by adding NaBH4(s) and sparging with Ar for 1 hour or more. 128 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 2. Optimized System Reagents (by Analyte)a HgII

MMHg

Temperature

20

20

Flow

0.5

0.5

[HCl]

1.0

0.1(0.05)b

[HAc]

1.75

1.55(1.6)b

[TU]

0.15

0.15

Flow

0

0.25

[HCl]

--

2.5

Temperature

40

40

Flow

1.00

1.00

[KBrO3]

0.17

0.17

Temperature

20

20

Flow

0.25

0.25

[NaAsc]

1.0

1.0

[Triton-X]

1%

1%

Temperature

0

0

Flow

0.25

0.25

[KOH]

4.5

4.0

Temperature

20

20

Flow

1.00

1.00

[KOH]

1.0

1.0

Eluent


Post-column acid (HCl)

Oxidant (

)

Antioxidant (Asc)

Base (KOH)

Reductant (

)

[NaBH4]

0.005 min–1,

All values of T in °C, Flow in mL and [X] in M units. used for greater MeHg+/EtHg+ separation. a

0.005 b

Low [HCl]EL formula

Successful operation of the system absolutely requires the use of TU that is unoxidized and has been cleaned of contaminating Hg. Upon purchase, even reagent grade TU must be examined to be sure it is free of breakdown products, which give off a strong sulfur smell or have a yellow color, and stored in a freezer. 129 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


In our experience, reagent grade TU always contains excessive Hg. Thus, we clean stock solutions (1 M TU) by adding 30-g pre-cleaned DOWEX 50W-X8 (100-200 mesh) per liter of TU solution and stirring for 1 h. Cleaned TU stocks can be stored at 4 °C for at least 14 days, or kept frozen, although we normally prepare eluent fresh each day. Oxidation of concentrated TU solutions can be detected by testing for the formation of white, S-containing polymers/particles after buffering the solution to pH 5. As Cu2+ catalytically oxidizes TU, the DOWEX must be cleaned of metals. First, DOWEX is washed by stirring in 1 M HCl for 1 h. After filtering the resin and rinsing with high-purity deionized water (DIW), it is then resuspended in 1 mM EDTA solution and neutralized with KOH. After 1 h of further stirring, the DOWEX is filtered and rinsed while on the filter with DIW. Finally, the clean TU stock is stored at 4 °C until use Additionally, we note that the thiol resin can also contain ppb levels of Hg. Most Hg can be removed by shaking batches of resin overnight in a solution of H2SO4 (2 M) and TU (150 mM) and then rinsing with DIW. Prior to use, we pump 20 mL of cleaning solution (2 M HCl + 150 mM TU) through the concentrator.

Preconcentration Using Thiol-Functionalized Resins The ability to efficiently concentrate Hg2+ and MeHg+ on thiolated resins using the pH-modulated, thiol-TU switch is an integral part of the original system that we still exploit both in off-line preconcentration from water samples and in trapping analytes in on-line concentrator prior to injection. In our earlier HgTU-IC work, we always employed a custom-synthesized, thiol-functionalized, divinyl benzene resin (28, 29). Recently, we have found that SiliaMetS®-Thiol (Silicycle, Montreal) makes an excellent replacement. From a consideration of the ligand exchange reactions (10) and (11) involved in preconcentration, one expects the trapping efficiency of the resins to depend on the pH and concentration of Hg-binding ligands in the solutions being loaded. To identify the proper pH range for efficiently trapping the analytes of interest from solutions containing our primary analytical ligand, we added MeHgCl and HgCl2 to solutions buffered to pH’s ranging from 0 to 5 with Na2SO4/H2SO4 (ionic strength 0.1 to 1) and amended with TU at 150 mM. The retention of analytes by our on-line concentrator containing ~100 mg of resin varied monotonically with pH (Figure 7). Both Hg2+ and MeHg+ were trapped completely at pH above 4 and less than 15% below pH 1, with Hg2+ being trapped to a greater extent between pH 1 and 4. The postulated competition between dissolved ligands and resin-bound thiols for the analytes during sample loading – equations (10) and (11) – also suggests that the trapping efficiency of the thiol resin depends on the concentrations of Hg-binding ligands in the solutions being loaded, i.e., bisulfide, thiosalicylic acid, glutathione, etc. To test the effect of these ligands on Hg trapping efficiency, isotopically-labeled MeHgCl and HgCl2 were added to pH 8 solutions containing 1 mM of each ligand. The experimental solutions were then pumped through the 130 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


on-line thiol concentrator. Any Hg retained by the trap was then eluted directly into the on-line reaction system and quantified. Consistent with a high affinity for Hg of the thiol resin, both analytes were completely trapped from 1 mM solutions of glutathione and thiosalicylic acid. MeHg+ was completely trapped from a solution containing 1 mM bisulfide, but only 70% of Hg2+ was retained. Since most samples are acidified for preservation, normally any sulfide present in a natural sample would be volatilized prior to preconcentration. However, to ensure that none remains behind, we recommend that samples in which the presence of sulfide is suspected should be treated by acidification and bubbling to strip H2S before attempting to trap Hg.

Figure 7. Effect of sample pH on retention of Hg species by on-line concentrator (thiol trap). [TU] = 0.15 M in loading solution.

Injection of the Analytes Analytes are injected into the high pressure end of the system via either i) a sample loop or ii) a thiol trap (concentrator) mounted on the injection valve. Each method of injection has distinct advantages and limitations that make them suitable for samples from different environmental matrices. Between the two methods, a very wide range of aqueous sample preparations can be analyzed. Sample Loop To inject via the sample loop, the sample matrix must match the eluent. This is not an inconvenience, at least for certain common sample preparation types in which system eluent can be used directly (see sample preparation). Due to limitations on the volume of sample that can be added via the loop – generally 131 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

between 10 and 100 µL but we have tested up to 1 mL – this means of injection is ideal for analyzing digests of sediments and biological tissues, but TU-SPE eluates of ambient water samples require further preconcentration.


On-Line Concentrator/Thiol Trap (TT) At present, we manually load samples onto the on-line concentrator, or thiol trap (TT), via a loading pump using a procedure adapted from Vermillion (29). It is designed to ensure i) complete trapping of analyte, ii) prevent carryover from previous injections, and iii) minimize the MeHg+ and Hg2+ blanks. For the current system (see TT in Figure 6), the concentrator comprises SiliaMetS resin held in a custom column assembled from a 30-mm piece of 1/8”-ID PEEK tubing to which medium pressure, SuperflangelessTM fittings have been attached to each end and screwed into union junctions (all connections are ¼”-28). A PEEK frit is inserted between the tube and fitting at the downstream end (www.idex-hs.com). The new column is less prone to breaking and easier to pack than the column used in the original system. Prepared samples and the three wash/rinse solutions used in the loading process – eluent ([HCl]EL = 1 M), citrate buffer (0.1 M Na3Citrate) and DIW – are all kept in clean vials in a workspace located under a HEPA filter-fan unit. All solutions are pumped from their respective vials to the TT via a FEP sipper tube (~10-cm), a section of flexible pump tubing (~20-cm of Tygon E-3603), and a piece of 0.5-mm ID PEEK tubing (~15-cm) connected to the same side of the injection valve as the TT. The total void volume of the tubing before the trap is about 0.5 mL. The injection valve is switched at appropriate times between the positions for 1) loading the trap and 2) injecting the trapped analytes. The loading sequence begins with the valve in position 2). The loading line is flushed with enough DIW to displace the eluent that the tube is kept filled with between samples. Next, a solution of citrate buffer (0.1 M) is pumped through the tubing. Once the citrate is seen to pass through the injection valve, the valve is switched to position 1) so that the buffer flows through the TT. One mL of the citrate is then pumped through the trap and immediately followed by the sample, which is buffered to pH 4 just before loading. Then, 1 mL of DIW is added to the sample vial as a rinse to push any sample remaining in the tubing onto the trap. Finally, just enough eluent to fill the loading lines is pumped in so that they are cleaned between samples. Before this eluent flush reaches the TT, the valve is switched to position 2) so the mobile phase flows through the TT and the analytes are injected. Note that the system operator must ensure that the final flush of eluent used to clean the loading line does not reach the TT before switching the valve, or it will wash the analyte into the waste. Summary Because of the flexibility afforded by the two different methods of sample injection, the system is compatible with the main sample preparations used in the analysis of Hg species in water, sediments, and biota. Samples prepared 132 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

using extraction into organic solvents, such as toluene, can be analyzed after back-extraction into pure system eluent with [HCl]EL = 1 M (see Sediment MMHg below).


Ion Chromatographic Separation Operation of the HgTU-IC system using the original eluent recipe, which contains [HCl]EL at 1 M (28–30), was optimized to achieve baseline separation between MeHg+ and Hg2+ at a reasonably short retention time for the latter ion. This permitted both forms of Hg to be quantified in a single chromatogram. However, when operating the system in this way, it was implicitly assumed that it was not necessary to separate MeHg+ from other Hg species of low charge, since few such species were expected to be present in typical water samples from aquatic ecosystems (Table 3). For example, Hg0 is known to be present but is rapidly oxidized in the presence of TU (data not shown) and thus would be detected as HgII. EtHg+ is another form of Hg that the system as originally operated does not distinguish from MeHg+ (29). As EtHg+ has only been reported in aquatic sediments from a few locations (36), its presence was deemed to be an unlikely source of artifactual MeHg+, at least in surface waters.

Table 3. Detection of Environmentally-Relevant Forms of Hg by HgTU-IC Species

Occurrence

Fate in HgTU-IC System

Hg0

Widespread

Oxidized to Hg2+ before injection

Me2Hg

Only in seawater

Decomposes to MeHg+ in acidic solutions (37) (Untested)

MeHg+

Widespread

Well-defined peak

EtHg+

Rare

Well-defined peak near MeHg+

Hg22+

Not reported

Untested

Hg2+

Assumed equal to InHg

Well-defined peak

HgNP

Limited reports

Untested

[HCl]EL-Dependence of Retention Times To investigate how the separation of MeHg+ from EtHg+ and Hg2+ depends on eluent pH, the times of transit from injection to the start of peaks for each species were measured over a range of [HCl]EL from 0.05 to 2.0 M using our normal 50mm CG-5A guard column. The retention time (RT) of a Hg species on the column itself is the difference between the total transit time and the time of transit without 133 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


a column, which is about 3 min. The variation in transit times of the MeHg+ peak as a function of the strong acid content of the eluent was fitted by an equation of the form:

Since the RT of MeHg+ is inversely proportional to [HCl]EL over the range 2 to 0.05 M (Figure 8), it is a simple matter to adjust eluent composition so that the system operates at any RT between 1 min for MeHg+ results in an RT for Hg2+ that is too long to be a practical method of simultaneously quantifying both species. Note also that Hg2+ retention is also more sensitive to [HAc]EL (Figure 8).

Figure 8. Effects of eluent acidity on the transit times of MeHg+, EtHg+ and Hg2+ using a single 50-mm Dionex CG-5A. (see color insert) 134 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


EtHg+ is another form of monovalent alkyl mercury that can be present in natural samples, although it is rarely reported. Since HgTU-IC relies on cation exchange to separate Hg species, one does not expect much separation of MeHg+ from EtHg+. With 0.5-1 M [HCl]EL and our standard 50-mm column, the separation between is negligible, but as [HCl]EL decreases, the relative retention of EtHg+ increases until at [HCl]EL~0.05 M there is complete separation of EtHg+ from MeHg+ (Figures 9 and 10). Note that this separation chemistry differs from the reverse-phase, ion-pairing separation of MeHg+ and EtHg+ in thiourea-based eluents reported elsewhere (38).

Figure 9. Effects of column length and eluent pH on transit time of MeHg+ (left axis) and separation of MeHg+ and EtHg+ peaks (right axis). (see color insert)

Although chromatographic separation occurs mainly on the ICC, MeHg+ and Hg2+ do elute differently from the thiol trap as well. The difference in elution profiles was directly observed by analyzing samples without the ICC in place. Both Hg species rapidly elute off the thiol trap at [HCl]EL of 1.0 M, but at 0.1 M MeHg+ elutes 1-3 min before Hg2+ with the latter exhibiting a pronounced tail. These differences likely contribute to the observed peak broadening of Hg2+ relative to MeHg+ when analyzing with the ICC. 135 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


Figure 10. Smoothed HgTU-IC-ICP-MS chromatograms for a sample containing 198-enriched MeHg+ and ambient EtHg+A) before B) after mathematical source deconvolution; [HCl]EL = 0.1 M; two Dionex CG-5A 50-mm columns. See Hg Detection by ICP-MS for explanation of deconvolution. (see color insert)

Column Length We also investigated the transit times of MeHg+ and EtHg+ using three different column configurations at three different eluent pH values (Figure 9). The 50- and 100-mm column data correspond to results with one CG-5A guard columns or two in series, while the 250-mm data corresponds to results with an analytical column containing the same stationary phase (Dionex CS-5A). As expected, transit times increase with column length, with the trend being most linear at the highest pH tested (Figure 9). The separation of EtHg+ and MeHg+ also increases with column length, but the trend is less than linear above 100-mm. Getting good separation of MeHg+ and EtHg+ requires using 100-mm of column at [HCl]EL = 0.1 M (Figure 10). Figure 10A shows the individual isotopic signals measured (i.e., 198Hg-202Hg). Figure 10B displays the post-deconvolution data where “Hg198” represents signal from the [Me198HgCl] source and “HgAMB” represents the signal from the ambient EtHg+ source. The decovoluted chromatogram (Figure 10B) shows a clear separation of MeHg+ and EtHg+. See Hg Detection by ICP-MS for a full explanation of deconvolution. 136 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


Fluctuations in [HCl]EL Chromatograms typically have a non-zero baseline due to a combination of instrument noise and the unavoidable presence of trace Hg2+ in the eluent and postcolumn reagents. We have observed that not only does [HCl]EL control retention times, but since the retention of Hg on the column is a cation exchange reaction – equations (12) and (13) – and H+ is the main mobile phase cation, fluctuations in [HCl]EL can perturb the baseline of chromatograms. When injecting via the sample loop, noticeable baseline perturbations can be avoided by carefully controlling the composition of the sample injected. For example, blanks and standards injected via sample loop in matrices that match the mobile phase show no “solvent dips” and peaks are highly symmetrical. But, injection of samples with matrices not matched to the eluent inserts a slug of solution with differing [H+] into the eluent stream. When a slug with low [H+] is injected, a brief (~15 s) dip in the signal baseline occurs at zero RT. Similarly, if a slug of sample with high [H+] is introduced, at zero RT the baseline is briefly raised due to the resulting perturbation in the partitioning of Hg2+ between the mobile phase and the ICC. Our qualitative observations suggest that the magnitude of the fluctuations is proportional to the background Hg2+ in the eluent. Some baseline perturbation also results from injecting analytes via the thiol trap, since loading the trap requires that the pH of the prepared sample be buffered to pH 4, i.e., 3-4 units higher than that of the eluent. While it is possible to decrease the differences between the composition of the solution filling the trap and the eluent stream, it is not possible to reliably eliminate the dip without going to great expense to reduce Hg2+ background levels in reagents.

Effects of Hg-Binding Ligands Just as injection can introduce a fluctuation in [HCl]EL, it is possible to introduce into the eluent stream a sample slug containing Hg-binding ligands that could also perturb the interaction between the mobile phase and Hg2+ sorbed to the ICC. However, as dispersion mixes the sample slug with the eluent, any introduced ligand will become increasingly dilute and less able to bind Hg2+ in the face of competition from the high concentrations of TU and H+ in the eluent. If such a ligand did outcompete TU under those conditions, it could pull Hg2+ off of the ICC and ultimately cause a brief increase in the signal baseline that might be difficult to distinguish from a peak. Samples containing Hg-binding ligands including 1 mM cysteine, glutathione or thiosalicylic acid were injected, via sample loop, into the on-line system and through the ICC at [HCl]EL = 0.1 M. None of the ligand injections caused baseline perturbations. When injecting a sample loop containing more TU than is in the mobile phase, a baseline dip resulted due to extra oxidation demand/incomplete TU oxidation.


Post-Column Chemistry


Original HgTU-IC Post-Column Chemistry The post-column redox chemistry of the HgTU-IC system employs the same two main reaction steps – oxidation of all Hg species to Hg2+ followed by reduction to Hg0 − as other flow-injection methods for total Hg analysis (39), but its reactant fluxes and compositions are adjusted to also fully oxidize the TU in the mobile phase and to attain each process’s requisite pH. Although smooth baselines and high sensitivities were routinely achieved in earlier work using H2O2/UV-oxidation and reduction by alkaline Sn2+ (29), subsequently we observed formation of two types of fine particles within the post-column reduction loop that randomly disturbed the baseline. The grey/black particles observed were likely Sn(OH)2(s), as stannous hydroxide becomes supersaturated within the reduction loop. The white precipitates were likely elemental sulfur or formamidine disulfide, which are known products of TU oxidation with limited solubilities (40). We had some success in avoiding particle formation by carefully selecting high quality reagents. The precipitation of Sn(OH)2(s) could be mitigated to some extent by i) neutralizing the acid in the mobile phase after oxidation prior to mixing with the highly alkaline Sn2+ reductant (30), and by ii) raising the hydroxide concentration to 10 M (unpublished results). In addition, adding Triton-X to the antioxidant reduced formation of S-containing particles (6). However, some fine particles were always formed after several hours of operation using these chemistries. In order to avoid this problem, alternative chemistries for the oxidation/reduction steps were investigated. In particular, oxidizing the S-II in TU more completely and finding an alternative to reduction by alkaline Sn2+ were deemed essential.

Oxidation of Thiourea That TU was not completely oxidized in the original HgTU-IC method was shown in an experiment conducted using an alternative reductant, 20% SnCl2 in 20% HCl (41). When following UV/H2O2 oxidation with reduction by acidic SnCl2, we observed no formation of Hg0 from MeHg+, implying that either MeHg+ had not been oxidized or that enough TU remained that it could inhibit the reduction of the Hg2+ formed from MeHg+ oxidation. Since reduction clearly occurs with alkaline Sn2+ and since MeHg+ is stable under alkaline conditions, while TU is not (42), this result implies that hydrolysis of incompletely-oxidized TU permitted Hg reduction to proceed in the original HgTU-IC method. Rather than attempting to further optimize the H2O2/UV chemistry, we tested oxidation by bromate. KBrO3 is commonly employed as a precursor of the BrCl oxidant used to measure total Hg in water (23), and was adopted in in HgTU-IC, Shade’s update of HgTU-IC chemistry (30). When using yellow-orange dissolved Br2 is visible after the oxidant stream mixes with the TU-containing mobile phase, as indicated in equation (17). Although the formation of white S-containing particles was diminished at the published ratio of 138 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


bromate-to-TU fluxes (0.74: 1) (30), some very fine white particles still appeared in the neutralization loop after running the system for several hours. To determine how much bromate is necessary to completely oxidize the S-II in TU, the system was operated using acidic SnCl2 as the reductant in order to avoid alkaline hydrolysis of TU. Various bromate concentrations in the oxidant were tested using constant flow rates of all reagents and constant composition of eluent (1 M [HCl]EL, 150 mM [TU]EL) and reductant (20% SnCl2 in 20% HCl). Peak areas for 100 pg MeHg+ standards were measured and are reported here (Figure 11).

Figure 11. HgTU-IC-ICP-MS system response to injected MeHg+ while varying bromate/TU molar flux ratio with reduction by acidic Sn2+.

The dependence of the system sensitivity on bromate flux observed in the experiment (Figure 11) shows that oxidation of TU’s S-II is essentially complete when the bromate and eluent streams are mixed at a 1.6: 1 molar ratio of bromate to TU fluxes. This ratio agrees closely with the reported stoichiometry of TU when the latter is present in excess (43): oxidation by

Subsequently, we operated the system using a : TU flux ratio that slightly exceeds the Simoyi stoichiometry and with the oxidation coil immersed in a 40 °C water bath to increase the reaction rate. Excess oxidant is needed to ensure /Br2/BrCl, NaAsc is injected into complete TU oxidation. To quench the extra the flow and allowed to react in the antioxidant loop located before the reduction step (Figure 6). 139 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Borohydride Reduction


While not used as widely as acidic SnCl2, borohydride has been employed as a post-column reductant in at least one reverse-phase-HPLC Hg speciation system (44), and suggested as a suitable replacement for alkaline Sn2+ in the HgTU-IC system (30). We found that at reduction loop transit times of ~30 s, NaBH4 is an effective reductant at concentrations as low as 5 µM and that system sensitivity increases only 20% when its concentration is raised by five orders of magnitude (Figure 12).

Figure 12. Effects of concentration and residence time within the reduction loop on HgTU-IC system sensitivity to MeHg+.

Neutralization of Acid Prior to Reduction Loop Although is an effective reductant under acidic conditions, the H2(g) bubbles formed at low pH add fluctuations to the signal. Thus, only alkaline NaBH4 is a suitable reagent in the HgTU-IC system. Since sensitivity varies inversely with pH, a balance was found to maximize signal while safeguarding against the formation of H2(g) bubbles. To do this, the acid flux in the eluent was matched by the base addition, which was then mixed with the 1 M KOH in the is followed with Hg2+ reductant. When the complete oxidation of TU by , the HgTU-IC system can operate with no particle formation reduction by over long periods (>10 h). An advantage this version of HgTU-IC has over other LC methods is the nature of the waste stream it produces, i.e., it contains no organic solvents or metals. By adding a small amount of acetone and bring the waste to pH 4, one , leaving acetone, acetate, Triton-X, Br–, Cl–, K+, can consume any excess + Na , ascorbate oxidation products, and urea in the waste. Also, the final pH of this waste is much less alkaline than waste from the previous HgTU-IC chemistries, making neutralization easier and less expensive. 140 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Summary


We have found the post-column chemistry described herein to be very stable, free of baseline pertubations, and at least as sensitive as the best results obtained with earlier versions of the HgTU-IC system (29). The fluxes of acids and bases are more carefully balanced (Table 4) to yield optimal sensitivity while also simplifying neutralization of wastes. The oxidant flux is also substantially higher than in previous versions in order to eliminate the formation of elemental sulfur particles.

Table 4. Acid-Base and Redox Balances in Different HgTU-IC System Chemistries Analytes

MMHg + HgII

MMHg + HgII

MMHg + HgII

System Version

2005 (28)

2007 (29)

2008 (30)

Proton Balance (Fluxes in meq

HgII

MMHg This work

min–1)

Eluent HCl

0.45

0.5

0.5

0.5

0.05

Eluent HAc

0.87

0.875

1.3

0.875

0.775

Oxidant HCl

0

0

0

0

0.625

Oxidation products

0.2

0.2

0.14

0.15

0.15

–3

–1.13

–1

Base Reductant

–5.7

–5.8

–0.75

–1

–1

Net protons

–4.18

–4.23

–1.81

–0.60

–0.4

Electron Balance (Fluxes in meq min–1) TU

0.8

0.8

0.56

0.6

0.6

Oxidant

–0.68

–0.52

–0.25

–0.85

–0.85

Net electrons

0.12

0.28

0.31

–0.25

–0.25

Ascorbate

0.28

0.28

0

0.3

0.3

Hg Detection by ICP-MS This work is the first documenting the use of HgTU-IC with cold vapor generation and ICP-MS detection; all previous work was performed using CV-AFS (28) or ICP-MS analysis of HgTU-IC eluent without post-column reaction (32). Just as shown by workers using ICP-MS with ethylation-GC, the HgTU-IC-ICP-MS system yields individual chromatograms for each of the different Hg isotopes (Figures 10A and 13A). Note that in this chromatogram, the natural abundance of the Hg isotopes in the ambient MeHgCl standard is perturbed 141 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


by the added Me198HgCl internal standard. Thus, as has been previously shown for THg and MMHg analysis (41), employing ICP-MS detection allows one to perform species specific stable isotope dilution, which enhances precision through the use of internal standards and facilitates the use of isotopically-enriched tracers to assay rates of species transformations in environmental media.

Figure 13. Smoothed HgTU-IC-ICP-MS chromatograms containing a mixture of ambient and 198Hg-enriched MeHg+ A) before B) after mathematical source deconvolution; [HCl]EL = 0.1 M. (see color insert) An important benefit of using ICP-MS for the HgTU-IC system is that it enables the direct comparison of peak shapes of analyzed Hg species in unknown samples with those of known isotopically-labelled internal standards. To make such comparisons, we perform a mathematical deconvolution using the ICP-MS counts for each isotope at every time point (0.1 s resolution) in the chromatogram (Figure 13A) prior to the integration of peak areas, a step that is not commonly done (41, 45). The deconvolution yields chromatographic traces expressed in terms of counts of ambient Hg (HgAMB) and those of the labelled internal standard (Hg198) (Figure 13B). While the trained eye can readily detect the perturbation of the natural abundance of the Hg isotopes from the relative size of the Me198Hg+ and Me202Hg+ peaks, the consistency of the peak shapes and transit times are much more evident after the deconvolution. In natural samples, we recommend using linear regression of the individual time points from the deconvoluted chromatograms to quantify the similarity of the ambient and internal standard peaks. In this case, the correlation coefficient (R2) of the data points during the peak was 0.992. If there were an ambient Hg species 142 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

that overlapped with MeHg+ under the system operating conditions, such as EtHg+, there would be a clear deviation from linearity over the latter part of the peak and a poor correlation coefficient. The regression can be used as a tool to ensure that no other Hg species is overlapping with MeHg+ under the operating conditions.


Measuring Dissolved MMHg and HgII by TU-SPE/HgTU-ICICP-MS To analyze HgII and/or MMHg at low levels in natural water samples, an off-line TU-SPE step is necessary to pre-concentrate Hg species and reduce the amounts of undesirable matrix components (DOM and other metals) injected into the on-line system. Our current off-line TU-SPE procedure is almost unchanged from the original (29), with the exception that we now use the Silia MetS-Thiol resin and add isotopically-labelled internal standards in the leaching step. Briefly, 20- to 1000-mL subsamples of previously filtered and acidified primary samples are weighed into clean borosilicate glass vials. Stock solutions containing clean TU and isotopically-labelled MeHgCl and HgCl2 internal standards are added to attain the desired concentrations, e.g., TU at 10-40 mM. Typically, the capped vials are then leached overnight at room temperature, although shorter leaching can be used (29). Samples can be buffered to pH 4 by addition of an appropriate volume of Na3Citrate stock (0.75 M) either before leaching or just before the TU-SPE step. On the day off-line TU-SPE is performed, we prepare TU-SPE columns by slurry-packing ~100-mg of thiol resin into each of several borosilicate glass chromatography columns (Kontes, 1×5-cm) and placing a small wad of glass wool on top of the resin bed. The packed column is then washed with 20 mL of cleaning solution (2 M HCl and 150 mM TU) and rinsed with 10 mL of DIW. The sample loading sequence begins with pumping: i) ethanol (10 mL), ii) eluent (10 mL at 1 M [HCl]EL), iii) DIW (10 mL), and iv) sodium citrate buffer (1 mL of 0.75 M) through the packed column. Then, the leached and buffered samples are pumped through the column. Following sample loading, the adsorbed analytes are eluted with 4 mL of eluent (1 M [HCl]EL), which is then kept frozen until analysis. Note that by preconcentrating the Hg species off-line, relatively little of the DOM and other metals from the original water sample end up in the prepared sample that is loaded on-line. Much of the DOM flows through the resin column without being adsorbed and most of the DOM that does adsorb is left on the resin when the Hg is eluted. This cleanup enables one to load samples on-line without plugging the TT frit and removes solutes that might degrade the performance of the on-line thiol resin or ICC. Performance Metrics To test method performance for dissolved MMHg and HgII, detection limit (MDL) studies were performed. The MDLs were determined by analyzing seven replicate samples cotaining 1 pg MeHg+ and 5 pg Hg2+. The standard deviation 143 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


from the replicates was multiplied by the student’s t-value appropriate for 99% confidence level to calculate the MDLs. The Hg stock solutions were diluted into 40 mL samples and TU-SPE was performed as described above. Both analytes were measured in a single chromatogram, while operating the system at 1 M [HCl]EL. We expect the variability in MMHg peak integrations to be greater under these conditions relative to operating at 0.1 M. The resultant MDLs are 0.003 ng L–1 for MMHg and 0.01 ng L–1 for HgII. Synthetic samples were created in order to test the effect of DOM on Hg preconcentration by TU-SPE. A stock solution containing ambient MeHg+, Hg2+, and Suwanee River DOM (IHSS) in the mass proportions 1 ng: 5 ng: 10 mg was added to Hg-free spring water at three different dilutions and internal standards added for both analytes. All samples in the study were run in triplicate; reported concentrations are the average values after correcting for recovery of the internal standards using standard isotope dilution calculations (45). The recoveries of the ambient MMHg and HgII from 30 mL samples, calculated as percent of the value expected based on the dilution of the sample, was unaffected by DOM between 2 and 20 mg L–1 or by the presence of EDTA at 1 mM (Table 5). The average recoveries of internal standards were 93% and 92% for MMHg and HgII respectively.

Table 5. Concentrations and Recoveries of Hg Species from Synthetic Samples Containing Suwanee River NOM and EDTA Using TU-SPE / HgTU-IC-ICP-MS [DOM]

2 mg

L–1

8 mg L–1

20 mg L–1

Units

MMHg

HgII

MMHg w/EDTA

HgII w/EDTA

ng L–1

0.203 ±0.001

1.038 ±0.043

0.194 ±0.004

0.987 ±0.005

Recovery

101.7%

103.8%

97.2%

98.7%

ng L–1

0.773 ±0.006

3.952 ±0.054

0.792 ±0.010

3.849 ±0.148

Recovery

96.7%

98.8%

99.0%

96.2%

ng L–1

1.992 ±0.034

9.954 ±0.132

2.026 ±0.035

9.636 ±0.077

Recovery

99.6%

99.5%

101.3%

96.4%

Since in this method, the recovery of Hg from the samples depends on the presence of TU, it was deemed important to check whether the method’s performance is affected by the presence of a known TU oxidizer commony found in natural waters, Cu2+ (46). At 1.5 mg/L Cu2+, there was a significant amount of TU oxidation as evident from the clouding of the samples during leaching at pH 4. The addition of Cu2+ without EDTA did not affect the recovery of MMHg, but 144 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

did reduce the HgII recovery to an average of 86.2%. In the samples containing both 1.5 mg/L Cu2+ and 1 mM EDTA there was no clouding during the leaching step and recovery of both MMHg and HgII was quantitative. Based on these results, it is recommended that samples known to or suspected of containing high levels of Cu be amended with enough EDTA to bind it.


Sediment MMHg To analyze total MMHg in aquatic sediment samples, we employ a modified version of Bloom’s digestion/extraction technique (22). In the original method, an H2SO4/CuSO4/KBr solution is mixed with sediment to leach sediment-bound MeHg+ into solution as the neutral complex [MeHgBr]0. A simultaneous extraction with dichloromethane (DCM) removes the complex from the aqueous phase. After the extraction, the DCM phase is subsampled into a new vessel containing deionized water and the DCM allowed to evaporate, leaving behind the MeHgBr0 dissolved in deionized water. The resulting aqueous sample can be analyzed by ethylation-GC. For coupling with HgTU-IC, the same procedure could be used with eluent replacing water in the last step. However, it is convenient to make an additional substitution of toluene for DCM in order to work with a less volatile solvent. The resultant sample preparation is ready to be buffered and loaded onto the HgTU-IC system. To apply this method, ~0.2-0.5 g of sediment is shaken for 1 h with a mixture of 5 mL of 18% (w/v) KBr + 5% (v/v) H2SO4 and 1 mL of 1 M CuSO4. An appropriate amount of isotopically-enriched MeHgCl internal standard is also added at this stage. The combination of acid and ligand leaches MeHg+ into solution, forming the neutral MeHgBr0 species. Then 10 mL of toluene is added to the mixture and shaken for 1 h to extract the neutral MeHgBr0 species into the toluene phase and drive the desorption reactions to completion. After mixing, the sample is centrifuged to break up any emulsion that forms and to separate suspended solids from the toluene phase. Next, 80-90% of the toluene is transferred into a new vessel containing 5 mL of eluent and shaken for 1 h. The [MeHgBr]0 dissolves back into the aqueous phase. There the Br– anion is replaced by TU to form the charged MeHgTU+ species, causing all of the [MeHgBr]0 to be transferred from the toluene into the aqueous eluent. A sub-sample of the eluent can now be injected via sample loop or buffered and loaded onto the thiol trap as described above.

Other Sample Preparations Total Dissolved Hg For THg analysis the sample oxidation step adapted from USEPA Method 1631 is used (23). The sample is brought to 1-5% BrCl in order to oxidize DOM and allowed to react for at least 48 h. To quench excess BrCl, hydroxylamine 145 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

(NH2OH) solution (30%) is typically added at 15 µL per 10-mL of sample. Finally, the digested sample is brought up to 50 mM TU to prevent the potentially substantial adsorption of Hg2+ to the sample vial and the sample introduction system that can occur once the sample is buffered to pH 4-5 for on-line loading. The loading procedure is the same as described above.


Total Sediment Hg A wide variety of methods for digesting sediments for total Hg are known. Essentially all can be adapted for use with HgTU-IC by quenching excess oxidant with NH2OH or NaAsc, adding TU to keep the Hg2+ in solution, and buffering with citrate or acetate at pH ~4 just prior to loading.

Biological Tissue Digestions A simple and effective procedure to extract MeHg+ and Hg2+ from biological tissues by digesting them overnight in eluent at 60 °C has been reported by Shade (30).

Recommended System Operating Conditions Conditions for Hg Speciation Analysis The key system variables that one can tune to optimize the tradeoff between peak separation and analysis time are i) eluent proton concentration, or [HCl]EL and ii) length of the column (ICC). Since lengthening the column also raises the pressure in the system, we normally adjust [HCl]EL so that the peaks corresponding to the Hg species we need to distinguish and/or quantify i) can be resolved from each other and ii) come long enough after the injection dip to allow the baseline to be reestablished. To measure MMHg in samples containing little or no EtHg+, it is convenient to use one 4×50-mm ICC with an eluent that contains [HCl]EL of 0.1 M or less. Although this setup affords only partial separation of MeHg+ and EtHg+, when combined with the use of isotopically-enriched interal standards and ICP-MS detection one can identify the presence of EtHg+ by comparing the internal standard and ambient Hg curves in the deconvoluted chromatograms. When using AFS detection, the identification of MeHg+ peaks influenced by EtHg+ is less certain. Thus, when analyzing MMHg by HgTU-IC-AFS or ETHg with either type of detector, one should either employ two 4×50-mm columns at [HCl]EL of 0.1 M or one 4×50-mm columns at 0.05 M [HCl]EL, to achieve complete separation. To analyze HgII in samples not subjected to oxidative preparations, we recommend using an [HCl]EL of 1.0 M (Table 2) to adequately separate Hg2+ from singly-charged Hg species without excessively long total analysis times. 146 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

Table 6. Results of MMHg Analysis Using HgTU-IC-ICP-MS for Intercomparison Samples and Reference Materials Sample Mass (g)

HgTU-IC Value

Consensus Value

Intercomp 2014 “UB” (47)

50

0.110 ± 0.022

0.126 ± 0.080

Intecomp 2014 “SP” (47)

50

0.118 ± 0.007

0.103 ± 0.050

Intercomp 2015 “UJ” (48)

50

0.038 ± 0.002

300

0.045 ± 0.006

50

0.240 ± 0.015

300

0.240 ± 0.001

50

0.041 ± 0.001

300

0.054 ± 0.004

Dissolved MMHg (ng L–1)a


Intercomp 2015 “LS” (48)

Intercomp 2015 “CC” (48)

0.043 ± 0.011

0.22 ± 0.087

0.049 ± 0.011

Sediment MMHg (ng g-dw–1)b BRI-1 (49)

0.32

0.168 ± 200.003

0.17 ± 0.06c

IAEA 158

0.30

1.40 ± 200.14

1.38 ± 0.27

CC 580

0.23

72.9 ± 204.5

75.5 ± 3.7

a Samples prepared using TU-SPE. b Samples prepared using H SO /CuSO /KBr/toluene 2 4 4 digestion+extraction. c Provisional value; Due to wide scatter in results, median is reported here instead of mean of 0.20.

Conditions for THg Analysis There are two main setups for the HgTU-IC system that one can use to analyze THg, i.e., Hg2+ in samples subjected to preparative procedures that exhaustively oxidize all Hg species. The choice between the two depends on the range of concentrations one needs to quantify. The simplest method is appropriate for preparations containing relatively high levels of Hg. With such samples, one only needs to dilute them into eluent and inject small volumes via the sample loop. The time of analysis can be minimized by operating without the column (ICC), since only Hg2+ is present. Samples of surface waters or other media containing low levels of Hg must be injected into the system via the thiol trap. However, this method of loading creates an injection dip that slightly precedes but is not separated from the Hg2+ peak. Thus, in order to separate the two and accurately integrate the peak, one must employ the ICC. By adjusting [HCl]EL in the range of 1.0-1.3 M, one can allow the baseline to be reestablished after the dip while minimizing the retention time of the Hg2+ peak. 147 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


Method Validation While it is conceivable that a new analytical method such as this could yield systematically different results than the standard methods if it recovered more or less of the analyte, that is not the case here. Results from participation in blind intercomparison studies for dissolved MMHg and MMHg in sediment reference materials indicate that our results with HgTU-IC-ICP-MS are all within the 95% confidence limits of the consensus or certified values (Table 6). The high accuracy of the system/sample preparation combination for biota has already been documented (30). Note that when analyzing water, excellent results were obtained for MMHg with both 50- and 300-mL samples. Since the recoveries of analytes were high even in the large volume samples, routine analysis of water samples with detection limits in the low pg L–1 range for MMHg should be possible.

Acknowledgments We are grateful for the financial support of this work provided by i) a Natural Resources Conservation Service Conservation Innovation Grant to the Iowa Soybean Association, ii) Electric Power Research Institute project EP-P30063/C14095, and iii) USDA National Institute of Food and Agriculture, Hatch project 875-913. R.K. was supported by a postdoctoral fellowship from The Camille and Henry Dreyfus Foundation, Inc.. The training in speciated isotope dilution analysis of MMHg afforded by H. Hintelmann and his research group during R.H.’s sabbatical at Trent University was enormously influential in this work.

References 1.

2.

3.

4.

5. 6.

Driscoll, C. T.; Mason, R. P.; Chan, H. M.; Jacob, D. J.; Pirrone, N. Mercury as a Global Pollutant–Sources, Pathways, and Effects. Environ. Sci. Technol. 2013, 47, 4967–4983. Watras, C. J.; Morrison, K. A.; Host, J. S.; Bloom, N. S. Concentration of Mercury Species in Relationship to Other Site-Specific Factors in the Surface Waters of Northern Wisconsin Lakes. Limnol. Oceanogr. 1995, 40, 556–565. Davis, A.; Bloom, N. S.; Que Hee, S. S. The Environmental Geochemistry and Bioaccessibility of Mercury in Soils and Sediments: A Review. Risk Anal. 1997, 17, 557–569. Watras, C. J.; Back, R. C.; Halvorsen, S.; Hudson, R. J. M.; Morrison, K. A.; Wente, S. P. Bioaccumulation of Mercury in Pelagic Freshwater Food Webs. Sci. Total Environ. 1998, 219, 183–208. Selin, N. E. Global Biogeochemical Cycling of Mercury: A Review. Annu. Rev. Environ. Resour. 2009, 34, 43–63. Fitzgerald, W. F.; Lamborg, C. H.; Hammerschmidt, C. R. Marine Biogeochemical Cycling of Mercury. Chem. Rev. 2007, 107, 641–662. 148 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.

7. 8.

9.


10.

11.

12.

13. 14.

15.

16.

17.

18.

19. 20.

Vandal, G. M.; Mason, R. P.; Fitzgerald, W. F. Cycling of Volatile Mercury in Temperate Lakes. Water Air Soil Pollut. 1991, 56, 791–803. Deonarine, A.; Lau, B. L. T.; Aiken, G. R.; Ryan, J. N.; Hsu-Kim, H. Effects of Humic Substances on Precipitation and Aggregation of Zinc Sulfide Nanoparticles. Environ. Sci. Technol. 2011, 45, 3217–3223. Pham, A. L.-T.; Morris, A.; Zhang, T.; Ticknor, J.; Levard, C.; Hsu-Kim, H. Precipitation of Nanoscale Mercuric Sulfides in the Presence of Natural Organic Matter: Structural Properties, Aggregation, and Biotransformation. Geochim. Cosmochim. Acta 2014, 133, 204–215. Skyllberg, U. Competition among Thiols and Inorganic Sulfides and Polysulfides for Hg and MeHg in Wetland Soils and Sediments under Suboxic Conditions: Illumination of Controversies and Implications for MeHg Net Production. J. Geophys. Res. 2008, 113, 1–14. Hintelmann, H.; Welbourn, P. M.; Evans, R. D. Measurement of Complexation of Methylmercury(II) Compounds by Freshwater Humic Substances Using Equilibrium Dialysis. Environ. Sci. Technol. 1997, 31, 489–495. Lamborg, C. H.; Tseng, C.-M.; Fitzgerald, W. F.; Balcom, P. H.; Hammerschmidt, C. R. Determination of the Mercury Complexation Characteristics of Dissolved Organic Matter in Natural Waters with “Reducible Hg” Titrations. Environ. Sci. Technol. 2003, 37, 3316–3322. Hsu, H.; Sedlak, D. L. Strong Hg(II) Complexation in Municipal Wastewater Effluent and Surface Waters. Environ. Sci. Technol. 2003, 37, 2743–2749. Black, F. J.; Bruland, K. W.; Flegal, A. R. Competing Ligand ExchangeSolid Phase Extraction Method for the Determination of the Complexation of Dissolved Inorganic Mercury (II) in Natural Waters. Anal. Chim. Acta 2007, 598, 318–333. Hintelmann, H. 11. Organomercurials. Their Formation and Pathways in the Environment. In Metal Ions in Life Sciences; Sigel, A., Sigel, H., Sigel, R. K. O., Eds.; Royal Society of Chemistry: London, 2010; Vol. 7, pp 365–401. Skyllberg, U.; Drott, A. Competition between Disordered Iron Sulfide and Natural Organic Matter Associated Thiols for Mercury(II)–An EXAFS Study. Environ. Sci. Technol. 2010, 44, 1254–1259. Jonsson, S.; Skyllberg, U.; Nilsson, M. B.; Westlund, P. O.; Shchukarev, A.; Lundberg, E.; Björn, E.; Bjo, E. Mercury Methylation Rates for Geochemically Relevant HgII Species in Sediments. Environ. Sci. Technol. 2012, 46, 11653–11659. Paquette, K. E.; Helz, G. R. Inorganic Speciation of Mercury in Sulfidic Waters: The Importance of Zero-Valent Sulfur. Environ. Sci. Technol. 1997, 31, 2148–2153. Jay, J. A.; Morel, F. M. M.; Hemond, H. F. Mercury Speciation in the Presence of Polysulfides. Environ. Sci. Technol. 2000, 34, 2196–2200. Bell, A. M. T.; Charnock, J. M.; Helz, G. R.; Lennie, A. R.; Livens, F. R.; Mosselmans, J. F. W.; Pattrick, R. A. D.; Vaughan, D. J. Evidence for Dissolved Polymeric Mercury(II)-Sulfur Complexes? Chem. Geol. 2007, 243, 122–127. 149 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


21. Qian, J.; Skyllberg, U.; Frech, W.; Bleam, W. F.; Bloom, P. R.; Petit, P. E. P. Bonding of Methyl Mercury to Reduced Sulfur Groups in Soil and Stream Organic Matter as Determined by X-Ray Absorption Spectroscopy and Binding Affinity Studies. Geochim. Cosmochim. Acta 2002, 66, 3873–3885. 22. Bloom, N. S.; Colman, J. A.; Barber, L. Artifact Formation of Methyl Mercury during Aqueous Distillation and Alternative Techniques for the Extraction of Methyl Mercury from Environmental Samples. Fresenius’ J. Anal. Chem. 1997, 358, 371–377. 23. Method 1631, Revision E: Mercury in Water by Oxidation, Purge and Trap, and Cold Vapor Atomic Fluorescence Spectrometry; U.S. Environmental Protection Agency, Office of Water: Washington, DC, 2002. 24. Bloom, N. S. Determination of Picogram Levels of Methylmercury by Aqueous Phase Ethylation, Followed by Cryogenic Gas Chromatography with Cold Vapour Atomic Fluorescence Detection. Can. J. Fish. Aquat. Sci. 1989, 46, 1131–1139. 25. Method 1630: Methyl Mercury in Water by Distillation, Aqueous Ethylation, Purge and Trap , and CVAFS (Draft); U.S. Environmental Protection Agency, Office of Water, Washington, D.C, 2001. 26. Hintelmann, H. Personal communication, 2007. 27. Jackson, B.; Taylor, V.; Baker, R. A.; Miller, E. Low-Level Mercury Speciation in Freshwaters by Isotope Dilution GC-ICP-MS. Environ. Sci. Technol. 2009, 43, 2463–2469. 28. Shade, C. W.; Hudson, R. J. M. Determination of MeHg in Environmental Sample Matrices Using Hg-Thiourea Complex Ion Chromatography with on-Line Cold Vapor Generation and Atomic Fluorescence Spectrometric Detection. Environ. Sci. Technol. 2005, 39, 4974–4982. 29. Vermillion, B. R.; Hudson, R. J. M. Thiourea Catalysis of MeHg Ligand Exchange between Natural Dissolved Organic Matter and a Thiol-Functionalized Resin: A Novel Method of Matrix Removal and MeHg Preconcentration for Ultratrace Hg Speciation Analysis in Freshwaters. Anal. Bioanal. Chem. 2007, 388, 341–352. 30. Shade, C. W. Automated Simultaneous Analysis of Monomethyl and Mercuric Hg in Biotic Samples by Hg-Thiourea Complex Liquid Chromatography Following Acidic Thiourea Leaching. Environ. Sci. Technol. 2008, 42, 6604–6640. 31. Oiffer, L.; Siciliano, S. D. Methyl Mercury Production and Loss in Arctic Soil. Sci. Total Environ. 2009, 407, 1691–1700. 32. Hong, Y. S.; Rifkin, E.; Bouwer, E. J. Combination of Diffusive Gradient in a Thin Film Probe and IC-ICP-MS for the Simultaneous Determination of CH3Hg+ and Hg2+ in Oxic Water. Environ. Sci. Technol. 2011, 45, 6429–6436. 33. Clarisse, O.; Hintelmann, H. Measurements of Dissolved Methylmercury in Natural Waters Using Diffusive Gradients in Thin Film (DGT). J. Environ. Monit. 2006, 8, 1242–1247. 34. Gao, Y.; De Canck, E.; Leermakers, M.; Baeyens, W.; Van Der Voort, P. Synthesized Mercaptopropyl Nanoporous Resins in DGT Probes for Determining Dissolved Mercury Concentrations. Talanta 2011, 87, 262–267. 150 Evans et al.; Trace Materials in Air, Soil, and Water ACS Symposium Series; American Chemical Society: Washington, DC, 2015.


35. Fernández-Gómez, C.; Bayona, J. M.; Díez, S. Comparison of Different Types of Diffusive Gradient in Thin Film Samplers for Measurement of Dissolved Methylmercury in Freshwaters. Talanta 2014, 129, 486–490. 36. Cai, Y.; Jaffé, R.; Jones, R. Ethylmercury in the Soils and Sediments of the Florida Everglades. Environ. Sci. Technol. 1997, 31, 302–305. 37. Conaway, C. H.; Black, F. J.; Gault-Ringold, M.; Pennington, J. T.; Chavez, F. P.; Flegal, A R. Dimethylmercury in Coastal Upwelling Waters, Monterey Bay, California. Environ. Sci. Technol. 2009, 43, 1305–1309. 38. Dórea, J. G.; Wimer, W.; Marques, R. C.; Shade, C. W. Automated Speciation of Mercury in the Hair of Breastfed Infants Exposed to Ethylmercury from Thimerosal-Containing Vaccines. Biol. Trace Elem. Res. 2011, 140, 262–271. 39. Leopold, K.; Foulkes, M.; Worsfold, P. Methods for the Determination and Speciation of Mercury in Natural Waters-A Review. Anal. Chim. Acta 2010, 663, 127–138. 40. Gherrou, A.; Kerdjoudj, H.; Molinari, R.; Drioli, E. Modelization of the Transport of Silver and Copper in Acidic Thiourea Medium through a Supported Liquid Membrane. Desalination 2001, 139, 317–325. 41. Hintelmann, H.; Ogrinc, N. Determination of Stable Mercury Isotopes by ICP/MS and Their Application in Environmental Studies. In Biogeochemistry of Environmentally Important Trace Elements; Cai, Y., Braids, C. O., Eds.; ACS Symp Series Volume 835; American Chemical Society Publishing: Washington, DC, 2003; Vol. 835, pp 321–338. 42. Norr, M. K. The Lead Salt-Thiourea Reaction. J. Phys. Chem. 1954, 65, 1278–1279. 43. Simoyi, R. H.; Epstein, I. R.; Kustin, K. Kinetics and Mechanism of the Oxidation of Thiourea by Bromate in Acidic Soluton. J. Phys. Chem. 1994, 98, 551–557. 44. Ilgen, R. F. G. Coupling of the RP C18 Preconcentration HPLC-UV-PCOSystem with Atomic Fluorescence Detection for the Determination of Methylmercury in Sediment and Biological Tissue. Fresenius’ J. Anal. Chem. 1997, 358, 407–410. 45. Meija, J.; Yang, L.; Caruso, J. A.; Mester, Z. Calculations of Double Spike Isotope Dilution Results Revisited. J. Anal. At. Spectrom. 2006, 21, 1294. 46. Zatko, D. A.; Kratochvil, B. Copper(II) Oxidation of Thioureas in Acetonitrile. Anal. Chem. 1968, 40, 2120–2123. 47. Creswell, J.; Metz, J.; Carter, A.; Davies, C. 2014 Brooks Rand Labs Interlaboratory Comparison Study for Total Mercury and Methylmercury ( Intercomp 2014 ); Seattle, 2014. 48. Creswell, J.; Kilner, P. I.; Davies, C. 2015 Brooks Rand Labs Interlaboratory Comparison Study for Total Mercury and Methylmercury ( Intercomp 2015 ); Seattle, 2015. 49. Creswell, J. BRI-1 Wisconsin Freshwater Sediment Reference Sample for Total Mercury and Methylmercury, Provisional Report; Seattle, 2015.


Trace Materials in Air, Soil, and Water : Mercury-Thiourea Complex Ion

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