Partitioning of mercury onto suspended sediments in estuaries

Abstract

Radiochemical partitioning experiments using ²⁰³Hg have been undertaken with mixtures of river, seawater and sediment samples taken from three geochemically contrasting UK estuaries: the Plym, Beaulieu and Mersey. Species of dissolved Hg were determined using reversed-phase C₁₈ chelating columns and particulate species were determined by sequential leaching with 1 M NH₄OAc and 1 M HCl. Mercury had a high particle reactivity with partition coefficients, K_Ds, ranging from 10⁴ to 5 × 10⁵ ml g⁻¹, depending on salinity, the chemical composition of the end-member waters, and on the physico-chemical characteristics of the sediment. Dissolved organic matter present in the waters (humic substances and/or anthropogenic compounds) was found to be the main factor governing the forms of dissolved Hg and their reactivity. From the spiked ²⁰³Hg, up to 95% of the dissolved metal was retained on the C₁₈ columns for the Mersey waters, whereas this fraction was <60% in the Plym and Beaulieu waters. Quasi-irreversible adsorption of Hg onto particles from each estuary was observed over a time-scale of a few hours and <20% of total particulate Hg was released by the sequential leach. In this paper, physico-chemical processes are proposed to explain the estuarine behaviour of Hg and the results are discussed in terms of Hg availability in estuarine systems.

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Introduction

The behaviour and speciation of metals in estuaries play a major role in their biogeochemical cycling,^1,2 an understanding of which is essential in order to assess the impact of metal pollution on coastal and marine ecosystems. In particular, contamination of the environment by Hg is a major global issue^3,4 owing to the acute toxicity of this element. Recent advances in analytical techniques⁵ have enabled accurate measurements of low concentrations of various Hg species encountered in aquatic systems. A number of intensive investigations have been devoted to assessing Hg contamination and bioaccumulation in polluted sites.^6–10 However, there remains an incomplete understanding of the factors controlling the speciation and particle–water reactivity of Hg in estuaries, where reaction conditions change markedly on going from river to sea.

The work presented in this paper has attempted to address this question by controlled laboratory experiments in which the γ-emitting radioisotope ²⁰³Hg (t_1/2 = 47 days) was used as a tracer. The experiments were based on a simulation of estuarine processes, using salinity as an indicator of mixing. Water and sediment samples were obtained from three UK estuaries with contrasting geochemical properties, such as dissolved organic carbon (DOC), suspended particulate matter (SPM) concentrations and the type of SPM. The radiochemical partitioning experiments replicated instantaneous particle–water interactions, and therefore allow direct insight into trace-metal estuarine reactivity.^11,12 The advantages of using radioisotopes to simulate the behaviour of their stable analogues are: (a) the possibility of working at ultra-trace levels (pico- to nanomolar) offered by the low limits of detection of γ-spectrometry; (b) the elimination of contamination problems; and (c) the direct analysis of particulate and dissolved samples.¹³ A novel approach was used in this work to complement the partitioning experiments, by applying trace-metal speciation protocols to the ²⁰³Hg radioisotope in both the dissolved and the particulate phases. The experiments were designed to answer the following questions: (1) What are the physico-chemical factors controlling the speciation of dissolved Hg in estuarine waters? (2) What processes control the particle–water interactions of Hg in estuaries? (3) What are the environmental implications of Hg reactivity in estuarine systems?

Methodology

Sampling, sample treatment and analysis were performed using ultra-clean protocols. Ultra-pure water was obtained from a Millipore (Watford, Hertfordshire, UK) Milli-Q/Milli-Ro system (18.2 MΩ resistivity). All glassware and plasticware used in the radiochemical experiments were previously soaked for a least 24 h in a bath containing 10% Decon (Merck BDH, Poole, Dorset, UK), then in 3 M HCl and finally thoroughly rinsed with ultra-pure water. All reagents used were AnalaR or AristaR grade (Merck BDH). When possible, sample handling was performed in a laminar flow hood.

Sampling

Sampling sites description. For this study, water and sediment samples were collected from three UK estuaries selected for their diversity in geochemical and geographical conditions. The physico-chemical characteristics of the end-member water samples and of the sediment collected are given in Table 1.

Table 1 Physico-chemical characteristics of river and marine end-members samples and of sieved (<63 µm) estuarine sediment used in the radiochemical experiments^a

		Plym (May 1998)	Mersey 1 (June 1998)	Mersey 2 (October 1998)	Beaulieu (March 1998)
a DOC: dissolved organic carbon; POC: particulate organic carbon; SSA: specific surface area. b ND: not determined. c Total mineralisation using HNO3 + HCl + HF (96 °C). d Average and standard deviation of three replicates.
River water	pH	6.74	7.03	7.39	6.93
	DOC/mg C l^−1	1–3	9.7	9.33	17.7
	Conductivity/mS	0.5	0.63	0.38	0.27
Seawater	PH	8.11	8.07	7.88	8.09
	DOC/mg C l^−1	1–2	1.8	1.9	1.6
	Salinity	32.8	32.5	31.2	33.9
Sediment (<63 µm)	SSA/m^2 g^−1	6.1	9.3	6.0	6.9
	POC (%)	2.1	7.5	ND^b	6.9
	Total metal^c
	Al^d/mg g^−1	69.2 ± 1.4	64.0 ± 2.6	14.0 ± 1.5	33.7 ± 2.7
	Fe^d/mg g^−1	25.0 ± 0.4	48.0 ± 2.1	8.27 ± 0.9	36.7 ± 2.9
	Mn^d/mg g^−1	0.53 ± 0.25	1.20 ± 0.32	1.19 ± 0.13	<0.28
	Fe ∶ Al	0.36	0.75	0.59	1.09
	Mn ∶ Al × 10^3	7.7	19	85	<8.3

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The River Plym drains a moorland catchment area with granitic rock and various soil types. The estuary is characterised by relatively low DOC and SPM concentrations (Table 1). By contrast, the River Beaulieu exhibits elevated DOC concentrations (Table 1), since it drains an afforested and moorland catchment area. Whereas these two estuaries possess relatively ‘clean’ drainage basins, the Mersey estuary is located in an intense industrial area. The river drains the largest urban area in the UK and the estuary is, or has been, subject to untreated discharges of domestic sewage and industrial wastes.^14,15 Water and sediment contamination of the estuary includes a variety of partially identified organic compounds such as surfactants and organochlorines, and inorganic pollutants such as heavy metals, nitrogen and sulfides. In particular, Hg contamination still remains a major concern due to previous indirect discharges in dissolved and particulate form from chloroalkali plants and paper mills.¹⁶

Sample collection and pretreatment. Four sampling campaigns were performed on the three estuaries between May and October 1998; two campaigns were performed on the Mersey in June and October 1998.

Freshwater samples were collected from the shore at the limit of the saline intrusion and seawater samples offshore at the mouth of the estuary, using acid-cleaned 5 l Nalgene (Nalge Ltd, Hereford, UK) bottles. The samples were filtered within 24 h of collection through acid-cleaned 0.45 µm pore size cellulose acetate membranes (47 mm diameter, Millipore) using a Nalgene filtration kit, and stored at 4 °C. For DOC analyses, additional water samples were collected in 250 ml borosilicate glass bottles previously heated at 550 °C for 8 h. These samples were filtered through ashed 0.7 µm pore size GF/C membranes (47 mm diameter, Whatman, International Ltd, Maidstone, Kent, UK) using a glass filtration unit. After addition of 0.3% v/v H₃PO₄, the filtrates were analysed for DOC using a Shimadzu (Milton Keynes, UK) TOC-5000 total organic carbon analyser. Other physico-chemical parameters were measured during sampling, i.e., pH, salinity and temperature, dissolved oxygen and conductivity.

Surface estuarine sediments from the Plym and Beaulieu estuaries were collected manually from inter-tidal mud using a PTFE spatula, whereas sediment samples from the Mersey were collected at the same location as the river water using a Shipek grab. The sediments were separated across an acid-cleaned 63 µm nylon sieve using a small quantity of the native river water and stored at 4 °C until required for the incubation experiments. A few grams of the <63 µm sediment fraction were dried in a laminar flow hood at room temperature then at 60 °C overnight, ground in a mortar and stored in a vacuum desiccator. Total iron, aluminium and manganese concentrations were determined by GFAAS in these samples after HF mineralisation following a protocol described by Rantala and Loring.¹⁷ The dried sediments were also used for specific surface area (SSA) measurements using a N₂ gas adsorption method in a Micromeritics (Micromeritics Instrument Corporation, Norcross, GA, USA) Gemini 2360 analyser. Sediments for particulate organic carbon (POC) analysis were stored in sterilised borosilicate bottles, then separated across a 63 µm sieve and filtered through 0.7 µm pore size GF/C membranes (47 mm diameter; Whatman). After drying at room temperature, the samples were ground in an agate mortar and determined combustimetrically using a Shimadzu TOC-5000 CHN analyser.

Radiochemical experiments

The overall protocol for partitioning and speciation of ²⁰³Hg is shown in Fig. 1.

Fig. 1 Schematic flow chart of the radiochemical partitioning experiments, and of dissolved and particulate speciation procedures.

Partitioning experiments. The filtered fresh and seawater samples were mixed in varying proportions in 125 ml Nalgene bottles in order to simulate the estuarine salinity gradient. To 100 ml of these solutions of known salinity, a small quantity of ²⁰³Hg radioisotope (HgCl₂/HCl; Amersham International plc, Buckinghamshire, UK) was added in order to obtain a final isotope activity of approximately 100 Bq ml⁻¹ in each vial, corresponding to concentrations of around 20 nmolHg l⁻¹. Changes in pH resulting from the isotope additions were monitored by measuring the pH at the end of each experiment, and were found to be below 0.2. The spiked solutions were left to equilibrate overnight in the dark at a constant controlled temperature of 20 °C. A 20 ml aliquot was then pipetted from each incubation reactor into 20 ml scintillation vials for γ-counting (A₀). Another 20 ml collected from the reactors were passed through C₁₈ columns in order to isolate hydrophobic complexes formed with the added isotope, following the protocol described below. To the solutions remaining in the incubation bottles, a constant amount of sieved estuarine sediment was then added to reach a final concentration of approximately 100 mg l⁻¹ in each reactor. Preliminary incubation experiments using various proportions of suspended solid and added ²⁰³Hg showed that this solid ∶ Hg ratio was too high for any saturation of the particle binding sites to occur. Following equilibration for 24 h, under mechanical shaking, at 20 °C and in the dark, the particulate and dissolved fractions were separated by filtration through pre-weighed 0.45 µm pore size cellulose acetate membranes (22 mm diameter; Sartorius AG, Göttingen, Germany). Preliminary kinetic studies showed that 24 h of incubation were sufficient for the partitioning process to reach equilibrium, due to the high particulate reactivity of mercury. The filters and 20 ml of the filtered solutions were then placed into scintillation vials to measure particulate and dissolved ²⁰³Hg activity, A_P and A_S, respectively (Fig. 1). The filters were dried at room temperature and weighed for SPM determinations, and then each filter was sequentially leached following the procedure described below.

Gamma-energies (279 keV) were counted using a high performance Wallac 1480 Wizard 3″ γ-counter with NaI detection (Perkin-Elmer, Cambridge, UK); each sample was counted three times for 1000 s and automatically background corrected. Given the relative length of the overall experiments and the relatively short half-life of ²⁰³Hg, all activities were corrected for radioactive decay.

Mercury partition coefficients, K_Ds (ml g⁻¹), were calculated as follows:^11,12

where A_P and A_S are the blank corrected ²⁰³Hg activities on the filter and in solution, respectively; V is the volume of solution in ml; m is the dry mass of sediment on the filter in mg; and f is the geometry correction factor, determined from the ratio of measured ²⁰³Hg activity per spike absorbed in a blank filter and dissolved in 20 ml of ultra-pure water.

Speciation of dissolved mercury. Hydrophobic complexes formed between the added ²⁰³Hg and the organic compounds present in the reactors were isolated by passing 20 ml of the spiked solutions by gravity (flow rate around 20 ml h⁻¹) through reversed-phase Sep-Pack C₁₈ columns (200 mg resin C₁₈ Bond Elut LRC; Merck Eurolab S.A., Fontenay-sous-Bois, France). The procedure for conditioning and recycling the columns has been reported elsewhere.^18,19 Solutions collected at the end of the columns were placed in scintillation vials for γ-counting; the measured activity, A_C, corresponded to the so-called ‘non-organic’ fraction of ²⁰³Hg (Fig. 1). The ‘organic’ (or ‘Hg-C₁₈’) fraction was calculated from the difference between the initial ²⁰³Hg activity (A₀) and A_C. Each column was then rinsed with 5 ml ultra-pure water and eluted by 10 ml 2 M HNO₃ into scintillation vials for γ-counting; the Hg-C₁₈ fraction could then be separated into ‘eluted’ (A_E) and ‘resistant to 2 M HNO₃’ fractions.

Speciation of particulate mercury. After counting, the dried and weighed filters were placed in 20 ml polyethylene tubes where they were submitted to a two-step sequential extraction procedure,²⁰ consisting of a first extraction by 20 ml 1 M NH₄OAc for 5 h and a second extraction by 20 ml 1 M HCl for 5 h. Between each extraction, both filters and leachates were placed into scintillation vials for γ-counting, then the filters were dried at room temperature for 48 h and weighed. The respective ‘exchangeable’ and ‘leachable’ fractions were calculated from the activities measured in the leachates after each extraction, the activities on the filters being used for mass balance calculations. ²⁰³Hg remaining on the filters after the two extractions was considered as ‘residual’.

Analytical figures of merit and mass balance calculations. Mercury recoveries were calculated from the initial activity A₀ and the sum of the activities in the particles, A_P, and in solution, A_S. Results of five replicate partitioning experiments gave an average ²⁰³Hg recovery of 83 ± 11%. Owing to its high particle reactivity, losses of ²⁰³Hg by wall adsorption were sometimes observed during an experiment. These losses were assessed by a kinetic study of ²⁰³Hg wall adsorption and by rinsing the incubation reactors with concentrated HCl after the partitioning experiments. A maximum of 15% of the original amount of ²⁰³Hg was lost to the walls. However, this percentage was significantly reduced with increasing salinity and by the presence of suspended material. Most of the ²⁰³Hg adsorbed onto the walls during the overnight equilibration period was recovered by the addition of sediment, giving recoveries as high as 100% in some cases. In the experiments where a significant amount of the added mercury was lost to the walls, it was not possible to estimate whether this loss would affect ²⁰³Hg activity in the particles or in solution. Furthermore, activities measured in solution were sometimes close to the background value, increasing the error on the K_D. For all of these reasons, the values for the K_D standard deviation [s(K_D)] were estimated using the recovery values obtained at each salinity, as follows: s(K_D) = ±K_D  × % recovery.

Validation of the C₁₈ speciation scheme was performed using breakthrough and elution studies of five replicates. Saturation of the C₁₈ columns was not observed after passing through 20 ml of spiked solutions. γ-counting of 2 ml fractions eluted by a total volume of 10 ml HNO₃ revealed that 90% of the total eluted ²⁰³Hg was found in the first 2 ml of eluate collected. However, recoveries calculated from the initial activity, A₀ (Fig. 1), and the sum of the activities at the end of the columns, A_C, and in the eluate, A_E, were only 30% on average. The Hg-C₁₈ fraction, operationally defined as the organically bound mercury, was therefore calculated from the difference between the activities at the top and at the end of the columns (A₀ − A_C). This fraction includes all the metal retained on the columns, either by chelation between mercury hydrophobic complexes and the C₁₈ functional groups, or by adsorption of ²⁰³Hg onto the columns. A_E was used to calculate the eluted and resistant to 2 M HNO₃ fractions. The Hg-C₁₈ percentages given have a relative standard deviation of 6% (n = 5).

Recoveries of particulate mercury after the partitioning experiments were calculated from the initial activity on the filters and the sum of the activities on the filters and in the leachates after each extraction. Reproducibility of the procedure was better than ±3% (n = 5), with an average ²⁰³Hg recovery of 95 ± 3%.

Results and discussion

Sample characteristics

Physico-chemical characteristics of the samples used in the partitioning experiments are given in Table 1. The diversity of the catchment area between the three estuaries was reflected by the contrasting organic contents of the river water samples collected, with DOC concentrations ranging from 1 mg C l⁻¹ in the Plym river to 17.7 mg C l⁻¹ in the Beaulieu. Intermediate DOC concentrations were found in the Mersey river for both June and October surveys.

The estuarine sediment samples collected also exhibited varied characteristics. The Mersey particles from the June survey possessed the highest POC, SSA and total iron and manganese concentrations (Table 1). These three parameters are interrelated, associations of Fe/Mn oxides and organic matter forming coatings at the surface of particles; they are thus an important criterion for particle surface reactivity.^2,21 Aluminium normalisation can be used to assess the importance of clays and alumino-silicates with regard to metal sorption compared to other reactive phases, such as ferromanganese oxides and POC.²² The Mn ∶ Al ratios were higher in the Mersey sediments than in particles from the other estuaries, again suggesting their higher surface reactivity due to the presence of Mn oxides.

Partitioning and speciation of mercury on estuarine mixing

From thermodynamic modelling, Hg can be characterised by highly stable soluble complexes, either by binding with humic materials in fresh waters,^23,24 or by complexation with seawater chloride ions to form mainly the four species HgCl⁺, HgCl₂, HgCl₃⁻ and HgCl₄²⁻.²⁵ However, whereas a predominance of dissolved forms of Hg would be expected from these predictive calculations, Hg partitioning in the three estuaries studied was dominated by its high particle reactivity, with K_D > 10⁴ ml g⁻¹ in all cases, as shown in Fig. 2. Previous experiments conducted with the ²⁰³Hg radioisotope similarly provided typical K_Ds around 10⁵ ml g⁻¹.^10,26–28 Another feature of Hg partitioning was the general increase of K_D with increasing salinity (Fig. 2), despite the stability of chlorocomplexes. Although the relative standard deviation on the K_D values in seawater was as high as 30% in the Plym and the Beaulieu, due to low ²⁰³Hg counts in solution, the increase of K_D with salinity was significant.

Fig. 2 Mercury partition coefficients (K_Ds) in the Plym, Mersey and Beaulieu estuaries as a function of the salinity; error bars represent standard deviation of five replicates.

The apparent contradiction between the stability of dissolved Hg complexes and its strong sorptive behaviour may be explained by the presence of the thermodynamically favoured HgOH⁺ hydrolysed species in oxic waters and in the range of pH considered.¹¹ Hg hydrolysis competes with the formation of other stable complexes such as chlorocomplexes and is enhanced at higher pHs, providing one possible explanation for the increased K_Ds in seawater for the Plym, Beaulieu and Mersey (June) estuaries (Fig. 2). No increase was observed with the October Mersey samples, possibly due to the small pH difference between the river and seawater samples collected (Table 1).

In addition, mercury rich colloids can be formed either by coagulation and aggregation of Hg hydrolysed species or by metal sorption onto existing colloids. This ‘colloidal pumping’, evidenced by Stordal et al.²⁷ using ²⁰³Hg incubation experiments, is particularly effective in an estuarine environment where the increase of ionic strength can cause flocculation of macromolecular organic matter and associated metals,²⁹ and thus their transfer into the particulate phase.

The speciation of dissolved Hg performed through the use of reversed-phase C₁₈ columns confirmed its affinity for hydrophobic organic compounds, with the Hg-C₁₈ fractions exceeding 40% of the total dissolved Hg in the waters of the three estuaries studied, up to a salinity of approximately 20 (Fig. 3). The Hg-C₁₈ fraction significantly decreased with increasing salinity for all estuaries, as the result of chlorocomplexation in seawater. However, the DOC concentrations of the waters could not be a factor explaining the variability of Hg–organic complexation, for three reasons. Firstly, a marked decrease of Hg-C₁₈ from 50% in the river to less than 1% in seawater was observed in the Plym estuary, but the difference in DOC concentrations between the fresh and seawaters in this estuary (Table 1) was too low to account for this decrease. Secondly, the Hg-C₁₈ percentages in fresh water did not follow the order of the organic contents of the river waters, i.e., Beaulieu > Mersey > Plym (Table 1). Thirdly, in spite of the similar DOC values in the June and October Mersey waters, a 30% difference in average was observed between the Hg-C₁₈ fractions of the two surveys; the curves otherwise followed the same trend along the salinity gradient (Fig. 3). For these reasons, the type of DOC present in the waters appeared to be the factor dominating dissolved Hg speciation, rather than purely DOC concentrations. It appears that organic complexes formed with a natural form of organic matter, such as humic acids, were less hydrophobic than organic complexes formed in the waters of the Mersey estuary, where the natural DOC may be mixed with anthropogenic organic compounds. Indeed, the Hg-C₁₈ fraction was generally lower in the Plym and Beaulieu waters than in those from the Mersey, where up to 95% of mercury was retained on the C₁₈ columns in the October survey. In addition, the organic complexes formed in the Mersey waters were more strongly bound to the C₁₈ columns, since less than 4% could be eluted by 2 M HNO₃ (Fig. 3). This resistant fraction was on average 30% in the Beaulieu estuary and 40% in the Plym.

Fig. 3 Percentage of Hg-C₁₈ in filtered (<0.45 µm) estuarine waters as a function of salinity: A, Plym; B, Mersey (June); C, Mersey (October); and D, Beaulieu; dark and light grey shades represent, respectively, the eluted by 2 M HNO₃ and resistant fractions; error bars represent standard deviation of five replicates.

The role of natural organic matter (NOM)

Solid/solution partitioning of Hg appeared related to its organic complexation, i.e., to the hydrophobic character of the complexes formed, and to their behaviour in the salinity gradient. Firstly, less organically bound Hg in seawater (Fig. 3) corresponded to an increased sorption onto the particles (Fig. 2). Furthermore, in the Mersey waters in October, where the highest organic complexation was found, the K_Ds were the lowest, below 4 × 10⁴ ml g⁻¹, and constant at all salinities. Mercury solubility could therefore be explained by competitive processes between its organic binding in solution, chlorocomplexation and particle sorption. The increased Hg sorption onto the particles from a salinity of 20 may be due to the adsorption of organic complexes onto the particles.^7,28,30 Indeed, whereas adsorption of NOM and associated metals onto solid surfaces was expected to decrease with increasing pH,¹ in seawater this trend may have been reversed by NOM salting out at high salinity and high ionic strength. The ‘salting out’ process, described by Turner and Rawling,³¹ arises in seawater because less water molecules are present to stabilise NOM in solution due to their association with the hydration sphere of dissociated seawater ions, and because of the compression of surface electrical double layer at high salinity.

However, it should be noted that Hg complexing capacities were not measured in this study. If any saturation of the waters complexing capacity by the added ²⁰³Hg occurred, dissolved Hg speciation and its binding to particles would have been greatly affected.³² The diminution of Hg complexing capacity with salinity could also be a factor explaining the increase of K_Ds.

The increased sorption of Hg with salinity was observed in all cases except in the Mersey in October (Fig. 2). Similar studies³³ also carried out on Mersey waters in June and October 1998 showed the same type of Hg distributions, and K_Ds higher in June than in October. No correlation with biological processes was found, although a dependency of Hg species to seasonal variations was established in other aquatic systems.^34,35

Mercury particle reactivity

The rapid, almost irreversible adsorption of Hg onto particles in aquatic systems has been demonstrated in a number of in situ^4,6,16,35 and laboratory^11,27,28 studies. In this work, the high particle reactivity of Hg resulted in elevated K_Ds (Fig. 2), and additional evidence was given by chemical extractions performed on the ²⁰³Hg incubated sediments. After only 24 h of incubation, more than 80% of total particulate mercury was found irreversibly bound to the solids and could not be released even after extraction with 1 M HCl (Fig. 4).

Fig. 4 Distribution of particulate mercury between the exchangeable, leachable and residual fractions in estuarine sediments: A, Plym; B, Mersey (June); C, Mersey (October); and D, Beaulieu as a function of salinity.

Several processes can be proposed to explain ²⁰³Hg irreversible migration into the solid lattice: sorption on high energy binding sites, isotopic exchange with Hg already present in the original particles, or formation of stable associations with particulate organic compounds.^7,28,36 Mercury may also bind ferromanganese oxides by surface sorption or co-precipitation and then become refractory by ageing of the oxides.^{8,23,35,37,38} Mercury surface precipitation³⁹ was observed in our laboratory at nanomolar Hg concentrations, using incubation experiments of the ²⁰³Hg radiotracer with several synthetic Fe and Mn oxides at various pHs. Another factor limiting Hg bioavailability in oxic waters may be the presence of metal sulfide phases in the particles. Recent research has shown that sulfide complexes are stable in saline waters,^40,41 and exchange reactions between Hg and other metals present in these phases are likely to influence Hg partitioning.

Our data suggested no salinity-induced desorption of mercury from the particles (Fig. 4), in accordance with the results from Jones⁴² who found very little desorption from natural Hg-contaminated sediments subjected to seawater over a 7 day period. This is consistent with the observation that Hg sorption onto the particles generally increased with salinity (Fig. 2). The difference in Hg partitioning between the two Mersey campaigns was attributed to the variability of the organic compounds present in the estuary. The particle characteristics were another parameter that had to be taken into account. The particles collected in June exhibited a higher total Fe concentration (48 mg g⁻¹ compared to 8.3 mg g⁻¹ in October) and a higher SSA (9.3 m² g⁻¹ compared to 6.0 m² g⁻¹ in October; Table 1). Their higher surface reactivity could therefore account for the 3- to 10-fold higher K_D values in June than in October. The higher reactivity of the Mersey particles compared to those from the two other estuaries was also reflected by the significantly lower Hg availability from these particles, i.e., the exchangeable and leachable fractions (Fig. 4).

The predominance of Hg in a refractory form in the particles (>80%) has important implications for the fate of Hg in the estuarine environment. From our results, Hg solubility and bioavailability would be severely limited in natural waters, as the majority of the metal was found in a refractory form unlikely to be released from the particles. The long-term consequences of such a particle reactivity would be the immobilisation and accumulation of Hg in estuarine sediments. However, surface sediments can be a major source of bioavailable Hg by biologically mediated methylation.^10,24,35 In estuaries, although the methylation rate diminishes at increased pH and salinity,⁷ inputs of inorganic or methylated Hg into the water column can occur following sediment resuspensions, particularly in macrotidal estuaries such as the Plym, Beaulieu and Mersey.

Acknowledgements

S.L.R. is very grateful to the Whistler 2000 Speciation Symposium organisers for the travel bursary awarded for her attendance at the conference. Many thanks are due to the following persons: Dr. P. D. Jones (Environment Agency for England and Wales) and Miss S. Watts (University of Plymouth) for assistance with sampling on the Mersey estuary, Mr D. Henon (University of Plymouth) for the HF digests, and Dr. A. R. Bowie (University of Plymouth) for proof-reading the final manuscript.

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Footnote

† Presented at the Whistler 2000 Speciation Symposium, Whistler Resort, BC, Canada, June 25–July 1, 2000.

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