ELSEVIER Marine Chemistry 59 (1997) 159-176

Behaviour of anthropogenic mercury in coastal marine sediments

Christian Gagnon a, * , l&lien Pelletier b, Alfonso Mucci ’

a Dipartement d’O&anographie, Uniuersite’ du Q&bee h Rimouski, Rimouski, Quebec, Canada b INRS-O&anologie, 310 Allie des Ursulines. Rimouski, Quebec, Canada G5L 3Al

’ Department of Earth and Planetary Sciences, McGill Uniuersity, 3450 University Street, Montreal, Quebec, Canada H3A 2A7

Received 24 January 1997; revised 9 June 1997; accepted 25 June 1997

Abstract

The diagenetic behaviour of anthropogenic mercury accumulated in sediments of the Saguenay Fjord (Canada) was investigated. Box-cores taken along its main axis and in the St. Lawrence estuary were analyzed for bulk sediment and porewater total and methyl-mercury concentrations as well as a number of other chemical variables. Mercury concentrations

as high as 10,000 ng g-’ (dry weight) were measured in a core taken at the head of the fjord attesting to the presence of large quantities of mercury discharged by a chlor-alkali plant operated two decades ago along the Saguenay River. Porewater mercury concentrations ranged from 17 to 500 ng 1-l but were not correlated to the Hg content of the solid phase. Most of

the mercury appears to be bound to organic matter, part of it is recycled with Mn and/or Fe oxides at the redox boundary whereas some may be adsorbed to or coprecipitated with the anomalously abundant acid volatile sulphides. These sulphides

are very susceptible to oxidation and provide a more reactive sink to Hg than would pyrite. Despite the closure of the

chlor-alkali plant in 1976 and relatively high sedimentation rates, surficial sediment Hg concentrations remain abnormally high. We investigated whether this observation could be explained by the diagenetic remobilization of Hg from the highly

contaminated sediments buried below or resulted from other processes. The remobilization of Hg from deeper layers appears to be too slow to account for the high surficial sediment concentrations. Resuspension of older, contaminated sediments upstream during spring runoff or submarine mass flow may explain these observations. Methylation increases the solubility and mobility of Hg in sulphidic sediments but the CH,Hg(II) flux (0.07 ng cm-* yr- ’ ) from the contaminated layers to the surface sediment is negligible and accounts for only 0.01% of the present accumulation rate of mercury at the sediment

surface. Dissolved and solid CH,Hg(II) profiles also indicate that this species may not diffuse through the thin oxic layer at the sediment-water interface. The estimated flux of Hg to the water column (20 ng cme2 yr- ‘1, however, could be underestimated since the activity of burrowing organisms would increase the exchange rate with the water column. 0 1997

Elsevier Science B.V.

Keywords; diageneais; mercury; methylmercury; sulphur; fluxes; remobilization; marine sediment

* Corresponding author. Present address: Environment Canada, Guidelines Division, 8th Floor PVM Bldg., Ottawa, Ontario, Canada KIA 0H3. Tel.: + I-819-9975929; fax: + l-819-9.530461; e-mail: christian.gagnon@ec.gc.ca

0304-4203/97/$15.00 0 1997 Elsevier Science B.V. All rights reserved. PIZ SO304-4203(97)00071-6

160 C. Gngnon et al. /Marine Chemistv 59 (I 997) 159- I76

1. Introduction

The importance of diagenetic reactions on remobilization of trace metals and organic contaminants in aquatic

systems is well established (Salomons and Forstner, 1984). In some cases (e.g., Andersson et al., 1990),

contaminated sediments accumulating in areas receiving industrial wastes were identified as a point source of trace metals to surrounding aquatic ecosystems. The Saguenay Fjord has been a site of industrial and municipal

waste discharge since the early 1930s (Loring et al., 19831, and thus serves, today, as a natural laboratory for

diagenetic studies of anthropogenic trace metals in organic-rich coastal marine sediments.

Previous studies on the contamination of Saguenay fjord sediments by heavy metals have highlighted

sediment enrichments in Hg, Pb and Zn dating back to the beginning of industrialization in this region, about 60

years ago (Barbeau et al., 198la; Loring et al., 1983, Gagnon et al., 1993). Mercury accumulation in these sediments took place between 1947 and 1976 during the operation of a chlor-alkali plant located along the

Saguenay river at Arvida about 24 km upstream of the head of the fjord (Smith and Loring, 1981). Of the more

than 300 metric tons of Hg used by the plant during its operation, an estimated 60 tons are buried in these sediments (Loring and Bewers, 1978). Inputs of Hg to the Saguenay river and fjord declined in the early 1970s

following the implementation of new governmental regulations (Loring et al., 1983) and closure of the plant in

1976. Analyses of surficial sediments of the fjord sampled in 1986 (Pelletier and Canuel, 1988; Pelletier et al., 1989), however, indicated that total Hg concentrations remained at high levels (610 f 330 ng g-l), well above pre-industrial levels (- 100 + 50 ng g-i; Barbeau et al., 198la; 25 k 10 ng g-‘, Louchouam and Lucotte,

1997) despite high sedimentation rates. A comparison of the vertical distribution of the relatively immobile

polyaromatic hydrocarbons (PAHs) (McGroddy et al., 1996) released during the same period by aluminum

plants also located at Arvida and total Hg (Pelletier et al., 1990) led to the hypothesis that the latter could have been diagenetically remobilized in these sediments.

Very few studies have been published on the diagenetic behaviour of Hg and its mobility in coastal marine sediments (e.g., Bothner et al.. 1980; Gobeil and Cossa, 1993). In most cases, key geochemical parameters such

as the distribution and speciation of Hg in both the dissolved and solid phases were not measured simultane-

ously, impeding the development of a diagenetic model which could explain and predict the mobility of mercury in recent sediments. In addition, questions about the role of sulphur geochemistry as well as the methylation

process on the diagenetic behaviour of mercury in natural environments remain unanswered (Gilmour and Henry, 1991; Gilmour et al., 1992; Matty and Long, 1995).

The purpose of this investigation was to identify the main geochemical factors contributing to the diagenetic mobility of anthropogenic mercury in coastal sediments and evaluate their relative importance with respect to

physical processes such as bioturbation and sediment resuspension. In order to provide a complete description of

the system in this paper, we included a synthesis of earlier results on the behaviour of methylmercury in the same environment (Gagnon et al., 1996b) as well as new data on the distribution of total Hg in the dissolved and

solid phases. The combination of all available data on mercury partitioning between the various components of

the sediments and information on sulphur species distribution in the sediments of the Saguenay fjord allowed us to elaborate a comprehensive geochemical box model which includes fluxes and reservoir sizes of Hg species in

these organic-rich coastal sediments.

2. Description of sampling area and sediment characteristics

The Saguenay fjord is a long (93 km), narrow (l-6 km) submarine valley incised in the Canadian Shield which joins the St. Lawrence Estuary at Tadoussac (Fig. 1) through a shallow sill (20 m). A deeper interior sill (80 m), approximately 18 km upstream, separates the fjord into two main basins. The seaward (outer) basin is 250 m deep whereas the landward (inner) and largest basin is 275 m deep over much of its length. The Saguenay fjord is a stratified estuary with a well-defined N 15 m fresh water lens which thins down and mixes

C. Gagnon et al./Marine Chemistry 59 (1997) 159-176 161

1’00’

Saguenay Fjord (Canada)

48’30’

Saguenay

SAG_& l -

Fig. 1. Map of the Saguenay fjord and the St. Lawrence Estuary showing sampling stations

progressively towards its mouth. The deep waters of the Saguenay Fjord are saline (S = 30) well-mixed, oxygenated, and their temperatures vary between 0.5 and 2°C annually (Schafer et al., 1990). Detailed

hydrographic characteristics of the fjord are given elsewhere (Schafer et al., 1990). Strong spatial variations in sedimentation rates and organic matter contents have been observed in Saguenay

fjord sediments (Edenborn et al., 1987). Sediment accumulation rates at the head of the fjord can reach 7 cm

yr-’ but decrease rapidly to 0.2-0.4 cm yr -’ in the landward basin (Smith and Walton, 1980). Sedimentation

rates, however, are highly variable over small horizontal scales because of the steep bottom topography and

frequent submarine mass flow events (Pelletier and Locat, 1993). The sediments contain 1.4 to 3.7% organic carbon (Co,,) whereas the inorganic carbon (C iNORG) generally accounts for less than 10% of the total carbon (C,,,) in the indigenous sediments (Gagnon et al., 1993, 1995). The settling organic material is mainly of

terrigenous origin (Pocklington and Leonard, 1979), particularly at the head of the fjord where the C:N ratio

reaches a value of 40 (Gagnon et al., 1993). The biochemical oxygen demand of the sediments is high so that

anoxic conditions and the onset of bacterial sulphate reduction are observed within a few cm below the sediment-water interface (Edenborn et al., 1987). Integrated sulphate reduction rates @RR) (O-30 cm depth) decrease rapidly from the head of the fjord (110 nmol SO, cm-* dd ‘) to the landward basin (32 nmol SO, cm -* dd ‘). Within the same depth interval and with the exception of sediments near the head of the fjord, porewater sulphate is not significantly depleted ( < 20%; Gagnon et al., 1995). In spite of measurable SRR, porewaters are H,S-poor ( < 6 PM: Gagnon et al., 1996a). This observation is attributed to the availability of

reactive iron which limits the buildup of H,S in porewaters as a result of iron sulphide precipitation. High acid

volatile sulphide:pyrite ratios (mean AVS:FeS,-S = 1.6) indicate that reduced sulphur is mainly accumulated in the form of metastable iron monosulphides and that the conversion to pyrite is comparatively slow in the absence of dissolved elemental sulphur or polysulphides (Mucci and Edenborn, 1992; Gagnon et al., 1995).

Sampled sediments were relatively homogeneous in texture and displayed a progressive decrease in their organic carbon content throughout most cores except for those recovered near the head of the fjord which

162 C. Gagnon et al. /Marine Chemistry 59 (1997) 159-l 76

contained layers of organic-poor marine clays originating from a major landslide which occurred 25 km upstream in May 1971 (Schafer et al., 1990). Since that time, normal sedimentation buried the landslide material deep within the sediment column. Previous observations of the position/depth of these layers (Edenborn et al., 1987; Gagnon et al., 19951, as well as their angle with respect to the sediment-water interface, indicate that some deposit layers may have been emplaced by subsequent submarine slumps several years after the original landslide. Wood fibers are often identifiable in cores recovered at the head of the fjord (Schafer et al., 19801 and most likely originate from the numerous sawmills operated upstream. Particle size distributions fall within the clay-silt fraction, with mean size ranging from 9 to 37 pm. Sediments are slightly coarser upstream but the mean grain sizes vary little with sediment depth. Sediment grain-size distributions support the sedimentation model proposed by Leclerc et al. (1986) which maintains that most of the suspended load enters the fjord from the Saguenay river.

Sediments were collected at four stations during two research cruises carried out in June 1991 and 1992. Three stations were located along the main axis of the Saguenay fjord (Fig. I): near the head of the fjord (SAG-5/1991; 90 m deep and SAG-6/1992; 100 m), at the junction of the Baie des Ha!Ha! in the landward basin (SAG-15; 180 m deep), and in the deepest portion of the landward basin (SAG-30; 265 m deep). Sedimentation rates (- 0.3 cm yr-’ > in the landward basin were taken from the work of Barbeau et al. (1981b) as established from 137Cs activity profiles. The sedimentation rate at SAG-5 and SAG-6 was estimated at N 1 cm yr- ’ by Mucci and Edenbom (1992) based on the position of the upper boundary of the landslide layer deposited in May 1971. A fourth sampling station was located offshore from Rimouski (Quebec) in the Laurentian Trough of the St. Lawrence Estuary (CL-l; 335 m deep), over 250 km downstream of the mouth of the Saguenay Fjord. Sedimentation rates at this station were estimated at 0.1 cm yr-’ (Silverberg et al., 1986).

3. Methods

Undisturbed sediments were collected using an Ocean Instruments@ Mark II box-corer. Cores were sub-sampled immediately, on-board, at various depth intervals in a nitrogen-filled glove-box (Edenbom et al., 1986) and samples stored at -20°C until analysis. Sediment porewaters from each sampling interval were extracted under nitrogen using ‘Reeburgh-type’ squeezers modified to filter the water through a 0.45 pm Millipore (HA) filter as it passed directly into a syringe (Mucci and Edenbom, 1992). About 10 ml of porewater were transferred to pre-cleaned glass bottles and acidified with a 1% equivalent volume of concentrated Ultrex @ nitric acid for total Hg determinations. The remaining porewaters (25-100 ml) were used for methyl-mercury, CH,Hg(II), determinations. These were transferred to pre-cleaned Teflon bottles in a N,-filled glove bag and frozen without acidification (Bloom, 1989). All manipulations was performed using rigorous trace metal clean protocols adapted from Gill and Fitzgerald (1987).

Total dissolved Hg analyses, including organic and inorganic species, were performed by a gold amalgama- tion preconcentration procedure, thermal desorption and cold vapour atomic fluorescence (CVAF) detection (Gill and Fitzgerald, 1987) after cold oxidation of organic mercury compounds by a BrCl solution (Bloom and Crecelius, 1983). The detection limit was 2 ng 1-l (3~ of the reagent blank) and the reproducibility was better than 6% for samples of 7 ml with concentrations above 50 ng l- ‘.

Total mercury in the solid phase was determined after acid digestion of the freeze-dried sediment (0.4 g> with a mixture of 6 ml HNO, and 0.6 ml HCl in Teflon reactors for 20 min in a microwave oven. Mercury was analyzed by cold-vapour atomic absorption spectrophotometry (CV-AAS) following its reduction to Hg” with stannous chloride. The precision determined with BCSS and BEST (NRC Canada) sediment standards was + 7% and the detection limit was 5 ng g-l.

The analytical methods for the determination of CH ,Hg(II) in sediment porewaters and solids were described previously (Gagnon et al., 1996b). Briefly, CH,Hg(II) in porewater samples was extracted with methylene chloride after treatment with copper sulphate and an acidic potassium bromide solution (Averty, 19841, and back extracted into a buffered (pH 4.9) aqueous phase (Bloom, 1989). CH,Hg(II) associated with the solid phase was

C. Gagnon et al./Marine Chemistry 59 (1997) 159-176 163

extracted with 10 ml of a KOH/methanol solution (Bloom, 1989, 1992) from 0.5 g of wet sediments. The extracted CH,Hg(II) was analyzed following derivitization using sodium tetraethylborate. The volatile alkylmercury derivatives were trapped on a Tenax@ -TA column (Ma, 1994), separated by gas chromatography (GC), thermally decomposed and measured as Hg” by CVAF (Bloom, 1989; Bloom and Fitzgerald, 1988).

The amount of Hg, Fe, and Mn associated with ‘reactive’ phases were determined following reaction of freeze-dried sediments with 1 N HCl for 24 h at room temperature cl:50 solid/solution) according to the method of Leventhal and Taylor (1990). The reduction of Hg(I1) to volatile elemental Hg prevented us from using other common operationally-defined extraction methods (e.g., citrate-dithionate-bicarbonate). The 1 N HCl-extractable Hg (Hg,,,) was analyzed by CV-AAS as described above. The Fe and Mn concentrations in the 1 N HCl-extracts and acidified porewater samples were determined by atomic absorption spectrophotometry (AAS) using an air-acetylene flame. Reproducibility was better than 5% and detection limits were, respectively, 0.6 and 0.06 mg 1-l for Fe and Mn.

Determination of sediment particle size distributions was performed with a LS 100 Coulter counter. They were used to identify the position of landslide or debris flow layers when not readily visible during subsampling.

4. Results and correlations

4.1. Distribution of total and dissolved mercury

The vertical distribution of total mercury in the sediment cores collected on both cruises are presented in Fig. 2. The historical record of mercury discharge from the chlor-alkali plant is clearly illustrated by the shape of

Hg (ng g*’ dry wt)

SAG-54

Landslide layer

10

m: _

30.

40-

SAG-IS-91

. . . . . . . . Landslide layer . . . .

CL-l-91

llwo zoo0 , , /

SAG-6-92 SAG-15-92

. . . Landslide layer

m

CL-l-92

Fig. 2. Depth distribution of total mercury in sediments collected in 1991 and 1992.

164 C. Gagnon et al. /Marine Chemistry 59 (1997) 159-l 76

these profiles. Despite the closure of the plant in 1976, mercury concentrations in surface sediments (O-l cm) remain high (N 500 ng gg’) when compared to the pre-industrial level ( < 100 ng g-’ ; Barbeau et al., 1981a; Louchouam and Lucotte, 1997) observed at the bottom of core SAG-1592 (40-45 ng g- ’ ). Assuming a sedimentation rate of about 0.3 cm y-’ in the landward basin (Barbeau et al., 1981b) these cores represent a 100 year record of deposition. The vertical distributions of total Hg in cores SAG-1591 and SAG-15-92 are noticeably different from each other and from other profiles reported in previous studies (Gobeil and Cossa, 1984; Gagnon et al., 1993). These discrepancies reflect the spatial heterogeneity of the sediments resulting from frequent submarine mass flow events on the landward slope of the inner basin (Pelletier and Locat, 1993). The presence of high Hg concentrations within the slump layers at SAG-15-92 probably resulted from the transport of contaminated sediments which had accumulated upstream. The highest Hg concentrations (up to 10,000 ng g- ’ ) were measured immediately below the 1971 landslide deposit layer in cores sampled near the head of the fjord (SAG-5 and SAG-6). The landslide occurred soon after the period of peak productivity and contaminated effluent release at the chlor-alkali plant. The Hg-laden sediments were rapidly buried by the Hg-poor clay layer thus preventing its rapid remobilization by oxic degradation of labile organic matter to which it was mostly associated (Smith and Loring, 1981) and which had accumulated at the former sediment-water interface. In comparison, Hg concentrations measured in sediments recovered from the remote St. Lawrence trough station (CL-l) were generally lower (average 370 ng gg ’ >, and displayed a rather smooth profile (Fig. 2). Mercury accumulating in the St. Lawrence Estuary originate from various sources including the Saguenay fjord, the St. Lawrence river, as well as direct inputs from the atmosphere (Gobeil and Cossa, 1993).

Dissolved Hg concentrations in the porewaters (Fig. 3) were highly variable, ranging from 17 to 500 ng 1-l.

SAG-S-91

SAG-b-92

SAG-IS-91

SAG-15-92

. . Debris flow

Hg (ns L-‘)

Mn (mg L-l)

SAG-30-92 Cl_-l-92

Fig. 3. Depth distribution of dissolved mercury and manganese in sediment porewaters collected in 1991 and 1992

C. Gagnon et al. /Marine Chemistry 59 (1997) 159-l 76 165

They are, however, consistent with values previously reported for the fjord (Gobeil and Cossa, 1984). Dissolved Hg accounts for less than 0.01% of the integrated bulk sediment Hg content. Its vertical distribution in the sedimentary column is also uncorrelated to that of the bulk sediment (e.g., r2 = 0.066 at SAG-591 and r’ = 0.012 at SAG-15-91). The lack of covariance between Hg in the dissolved and solid phases has already been noted in sediments from other areas (Bothner et al., 1980). The observations suggest that porewater Hg concentrations are not simply controlled by an exchange equilibrium between these two phases. Alternatively, the depth distribution of dissolved Hg may depend on the affinity of specific components of the sediments for Hg rather than on total solid Hg content (Bono, 1997). Distributions of dissolved Hg at SAG-5 and SAG-6 are characterized by higher values just above and below the landslide layers. Under the clay layer, high porewater concentrations could originate from the degradation of Hg-labile organic matter buried by the landslide.

Dissolved manganese (Mnniss) (Fig. 3) and extractable Mn and Fe (Fe,,, and Mnu,,) profiles (see also Fig. 4) generally serve as good indicators of redox conditions prevailing in the sediments. Despite the slow oxidation of Mn(I1) and its precipitation to oxides, the distribution of Mn u,-., provides a more easily discernable marker of the redox change than Fe,,, because the latter also includes a silicate component as well as products of AVS oxidation. The very low dissolved Mn concentrations observed near the sediment-water interface and their

HgHCL : HgToT Ratio

I 400

/ 20 40 60 80

40. 0 20 40 60 80

0.2 0.4 0.6 0.8 1.0

j ,qy ,

0 20 40 60 80

1.0 I

Mn.c,andFe~c~

(w g-' drywt)

Fig. 4. Depth distribution of 1 N HCl-extractable mercury, iron, and manganese in the solid phase in sediments collected in 1991.

166 C. Gagnon et al. / Marine Chemistry 59 (1997) 159-176

rapid increase within the first l-2 cm reflect the thinness of the oxygen penetration depth where insoluble detrital and authigenic Mn and Fe oxihydroxides accumulate. Upon burial and under the suboxic conditions established below, these oxihydroxides are readily reduced, releasing adsorbed or coprecipitated trace metals, including Hg, to the porewaters. Gobeil and Cossa (1993) demonstrated that Hg is strongly associated with a hydroxylamine/acetic acid extractable phase (i.e., Mn and/or Fe oxihydroxides) in the oxic layer of the Laurentian Trough sediments. Dissolved Hg peaks are observed at/or near the redox boundary in some, but not all the cores (Fig. 3). Their presence or absence seems to reflect the complexity of the sedimentation regime and the heterogeneity of sampled sediments. Peaks are also resolved at the sediment-water interface (e.g., SAG-591 and SAG-15-91) where Hg may be released as a result of organic matter degradation (Matty and Long, 1995), a behaviour akin to that of Cd (Gobeil et al., 1987). Dissolved Hg peaks were also observed at depth in the sulphidic sediments ( > 10 cm) of some cores (i.e., SAG-5-91, SAG-30-92, CL-1-91, CL-l-92).

4.2. Inorganic mercury and methylmercury partitioning

The distribution and the behaviour of methylmercury in sediment cores presented in this study have been discussed previously (Gagnon et al., 1996b). Briefly, dissolved CH,Hg(II) profiles display a subsurface maximum (up to 10 ng Hg 1-l > and highly variable concentrations in the anoxic layers, but it was barely detectable in both the porewaters and solids recovered from the surficial oxic layers. Porewater CH,Hg(II) accounts for less than 1% of the total CH,Hg(II) in the bulk sediment. Whereas solid inorganic Hg dominates in all cores, methylmercury associated with the solid phase represents 0.4% of the total Hg in the bulk sediment. This value is similar to results obtained in other estuarine sediments (e.g., 0.5%; Craig and Moreton, 1986). Despite the relatively low CH,Hg(II) concentrations in both the solid and dissolved phases, its environmental impact may be important because this compound is much more toxic than inorganic Hg and is readily bioaccumulated by aquatic organisms (e.g., Wood, 1984; Bloom, 1992).

Vertical concentration profiles of HCl-extractable iron (Fe,,,) and manganese (Mn,,,) are presented in Fig. 4. Depth distributions of 1N HCl extractable mercury (Hg,,,) are presented in the same figure as ratios of Hg.,,:total Hg (Hg,,, >. These are assumed to reflect an upper limit to the remobilizable fraction of the sedimentary Hg. The ratio is highest in the surface sediment (I 0.71, within the landslide layer at SAG-5-91 ( < 0.751, and at depth in most cores ( < 0.85). The strong correlation between the Hgnc,:HgTOT ratio and Fe,,, as well as with MnHc, concentrations in the oxidized, surface sediments provides more supporting evidence for the role played by authigenic and detrital iron and manganese oxides on the diagenetic remobilization of mercury as already documented by Gobeil and Cossa (1993).

5. Discussion

5.1. Speciation, association and mobility

Many factors may affect the chemical association and the remobilization of dissolved mercury in sediment porewaters and some of these factors can tentatively be identified and evaluated for their potential role in the Saguenay fjord. In spite of the sulphidic character of these sediments, relatively high porewater Hg concentra- tions were measured. Using a porewater CH,S concentration of 0.2-1.0 PM (Gagnon et al., 1996a), a pH of 7.5 (Mucci and Edenbom, 1992), and assuming equilibrium with HgS (typically Ksp = (Hg*‘XHS-)/(H+) = 10-38.9 in natural waters; Dyrssen and Kremling, 1990) calculations indicate that Hg concentrations should not exceed 30 ng 1-l (given an activity coefficient of 0.3; Whitfield (1975); Kremling (1983)) in these porewaters. Thus, the high concentrations of porewater Hg would suggest that HgS precipitation is not limiting its solubility. The formation of solid HgS from marine porewaters may be uncommon because the activity coefficient of

Hg ‘+ is most likely depressed by the presence of dissolved organic matter (DOM). Direct measurements of the activity of Hg in marine porewaters are still lacking to support or disprove this hypothesis.

C. Gagnon et al. / Marine Chemistry 59 (1997) 159-l 76 167

The formation of polysulphide as well as organic complexes has been suggested to explain the presence of porewater metal concentrations largely exceeding values predicted from the thermodynamic solubility of metal sulphides such as HgS (Lindberg and Harriss, 1974; Stunun and Morgan, 1981). Mercury can form stable, soluble complexes with sulphide and polysulphide species (HgSz-, HgHS;, Hg(HS),, etc.) which may dominate in sulphidic seawater (Dyrssen, 1985; Paquette and Helz, 1995). In fact, stability constants of mercury sulphide complexes, log K,,, = 42.0 and log KHgHs+ = 30.1, are the highest of all trace metals (Dyrssen, 1988). The predominance of the dissolved HgSi- species (98%) in reducing seawater has been inferred by Lu and Chen (1977) based on log p = 0.57 at pH 7.5 for the reaction:

HgS,,, + S2- @ HgS;- (1)

Although polysulphides concentrations are relatively low (CSi- + CH,S < 6 PM; Gagnon et al., 1996a) in Saguenay Fjord sediments, the formation of polysulphide complexes with Hg may still be quantitatively important given the magnitude of their formation constant. The formation of soluble complexes with other sulphur species (e.g., RS, S,Oi-) and DOM is also known to contribute to the solubilization of numerous metals in sulphidic environments (Boulbgue et al., 1982; Emerson et al., 1983). Few thermodynamic data, however, are available on the stability of mercury-DOM complexes in natural waters (Mantoura et al., 1978; Dyrssen and Wedborg, 1991; Gilmour and Henry, 1991) but Hg complexes with organic ligands such as EDTA are also among the most stable of all trace metals (Sillen and Martell, 1964, 1971; Martell and Smith, 1974; Andreae, 1986; Smith and Martell, 1989). Finally, recent observations indicate that a large fraction of the dissolved Hg may be associated with colloidal DOM (Guentzel et al., 1996; Stordal et al., 1996).

On the other hand, dissolved Hg can be adsorbed by several components of the solid sediment. Among these, sulphide minerals have been reported to be excellent scavengers for heavy metals, including Hg (Jean and Bancroft, 1986; Hyland et al., 1990) and significant pyritization of trace metals has been reported by Huerta-Diaz and Morse (1992) in sulphidic sediments. The anoxic sediments of the Saguenay fjord, however, are relatively poor in pyrite (FeS,) but anomalously rich in acid volatile sulphides (AVS: e.g., FeS) (Mucci and Edenbom, 1992; Gagnon et al., 1995). The adsorption and coprecipitation of trace metals with the synthetic iron monosulphide mackinawite in seawater have been investigated (Morse and Arakaki, 1993) and results suggest that this mineral may serve as a temporary sink for many trace metals during early diagenesis in anoxic sediments. In fact, the association of Hg with particulate organic matter and with AVS may represent the principal Hg sinks in the anoxic sediments of the Saguenay fjord (Bono and Mucci, 1995; Bono, 1997).

Mercury associated to authigenic sulphides, however, can be potentially remobilized following their stepwise oxidation (Moore et al., 1988; Huerta-Diaz and Morse, 1992) and Hg adsorbed to or coprecipitated with iron sulphides is a potential source of secondary contamination (Moore et al., 1988; Morse, 1994). The dominant solid sulphide phases in Saguenay fjord sediments, the metastable AVS, are very reactive towards oxygen (Morse and Comwell, 1987) and thus provide a less refractory sink for Hg than pyrite. In addition, Hg associated with iron sulphides may be released to porewaters through complexation with partially reduced sulphur species present in porewaters. It has been demonstrated experimentally that the second layer of Hg adsorbed on the iron sulphide group can be easily released using a weak complexing ligand (Jean and Bancroft, 1986) and the strongly bound Hg in the primary monolayer can also be removed by complexes such as thiosulphates (Hyland et al., 1990).

Detrital and authigenic Fe-Mn oxihydroxides which accumulate in oxic sediment layers can also be important scavengers of Hg (Jenne, 1968; Far& and Pickering, 1978). A significant fraction (up to 70%; Fig. 4) of the total Hg in the oxic layer of the fjord sediments is associated with a 1 N HCl extractable phase (i.e., Fe-oxihydroxides). The correlation between the distributions of reactive iron and manganese (Fe,,, and Mn,,,) and the ratio of HCl-extractable Hg: total Hg in surface sediments illustrates the probable importance of this scavenging process (Fig. 4). This association is sensitive to changes of redox conditions. Mercury will be released to the porewaters when iron and manganese oxides are buried and undergo reductive dissolution as a consequence of microbial degradation of organic matter (Gobeil and Cossa, 1993). The concentration profiles

168 C. Gagnon et al. /Marine Chemistry 59 (1997) 159-I76

and the mobility of Hg in sediment porewaters is therefore partially controlled by the remobilization (dissolution and precipitation) of iron and manganese oxides as the redox boundary migrates with sediment accumulation (Gobeil and Cossa, 1993).

The association of mercury with particulate organic matter is well documented. A strong correlation has been reported between the distribution of organic matter and Hg in numerous types of sediments (Lindberg and Harriss, 1974; Baldi and Bargagli, 1984) and for the Saguenay fjord (r = 0.9, Loring, 1975; Schafer et al., 1980; Pelletier and Canuel, 1988; Mucci and Edenbom, 1992). It has been suggested that terrigenous organic matter is the main scavenging agent and carrier of Hg to the sediments in the Saguenay fjord (Loring et al., 1983) and that it may also be bound to lignin, a wood pulp residue which is highly resistant to oxidation and biodegradation processes (Schafer et al., 1980). Recent results (Bono and Mucci, 1995; Bono, 1997) from sequential extractions, however, indicate that most (typically 50-85%) of the mercury found in Saguenay fjord sediments is associated with organic matter extractable by a 1 N NaOH solution. Similarly, Louchouarn and Lucotte (1997) demonstrated that there is no apparent correlation between lignin and Hg concentrations in these sediments.

In summary, adsorption and coprecipitation onto AVS and pyrite may limit the dissolved concentration of Hg in anoxic porewaters, but their oxidation could release it back into solution. Mercury in oxidized sediments would be mainly associated with fresh particulate organic matter as well as Fe and Mn oxihydroxides, effective scavengers of soluble Hg in oxic surface sediments.

5.2. Methylmercury speciation

The methylation of Hg can increase its concentration in sediment porewaters (Gagnon et al., 1996b). Whereas dissolved and particulate CH s Hg(I1) are nearly absent ( < 0.1 ng g- ’ as Hg) in the oxic surface layers of all cores, their concentrations increase rapidly in the sulphidic sediments. Our observations (Gagnon et al., 1996b) suggest that CH,Hg(II) accumulating in anoxic sediments cannot diffuse freely through the oxidized sediment-water interface because of bacterial and/or catalytic demethylation in this oxic zone (Jackson, 1989; Oremland et al., 1991, 1995). On the other hand, the anoxic and organic-rich sediments of the Saguenay fjord appear to provide conditions which are conducive to the methylation of Hg. Up to 29% of total dissolved Hg was found in the form of CH,Hg(II). In addition, the partitioning of CH,Hg(II) between porewaters and solid sediments may be controlled by selective adsorption onto organic matter or AVS (Gagnon et al., 1996b) as well as complexation by dissolved and colloidal compounds (Guentzel et al., 1996).

The speciation of CH,Hg(II) in Saguenay fjord anoxic porewaters can be assessed using the dissolved sulphur speciation data obtained from the same sediment cores (Gagnon et al., 1996a). Dissolved CH,Hg(II) forms predominantly chloro-complexes in saline oxic waters (Stumm and Morgan, 1981). In anoxic porewaters, soluble reduced sulphur species such as sulphides and thiols form strong complexes with the CH,Hg(II) cation (Dyrssen and Wedborg, 1991):

CH,Hg++ RS-+ CH,HgSR log K, = 16.12 (2)

CH,Hg++ HS-P CH,HgSH log K, = 14.5 (3)

In their study of CH,Hg(II) speciation in seawater, Dyrssen and Wedborg (1991) calculated that over 96% of CH,Hg(II) (1 pM or 0.2 ng 1~ ’ Hg total concentration) would be in the form of CH,HgSR or CH,HgSH at sulphide or thiol concentrations as small as 10 nM. Dissolved sulphide and organic sulphide concentrations measured in Saguenay fjord porewaters are approximately 1 and 10 PM, respectively. These concentrations are two to three orders of magnitude higher than required to convert nearly all of the dissolved CH,Hg(II) into CH,HgSR and CH,HgSH species.

Volatile dimethylmercury, (CH,),Hg, was not determined in this study, but traces ( < 20 fM = 4 pg I-’ as Hg) have been detected in porewaters from highly sulphidic sediments at the head of the fjord in a previous

C. Gagnon et al. /Marine Chemistry 59 (I 997) 159-I 76 169

study (Cossa et al., 1996). This latter observation is expected on the basis of the stoichiometry of one of the

proposed mechanism of (CH,),Hg formation (i.e., the abiotic transmutation of CH,Hg(II) into (CH,),Hg; Bartlett and Craig, 1981):

2CH,Hg + H,S + (CH,Hg),S + 2H++ (CH,),Hg + HgS + 2H+ (4)

The paucity of dissolved H,S in these fjord sediments, therefore, limits the formation of (CH,),Hg.

5.3. &fSects of bioirrigation on mercury distribution

Benthic macrofauna may alter the chemical and physical characteristics of sediment by moving seawater and

particles during feeding, burrowing, and tube construction (Aller, 1982; Aller and Rude, 1988). For example, transport of dissolved oxygen by burrowing organisms is known to cause a local oxidation of reduced inorganic

compounds (Oenema et al., 1988). The presence of partially oxidized sulphur species in the reducing sediments

of the Saguenay fjord is most likely the result of bioturbation and bioirrigation. Peak concentrations of

intermediate reduced sulphur species, such as thiosulphates, S,O:-, were measured in porewaters at depths

(> 15 cm) corresponding to that of maximum bioirrigation activity by polychaetes worms (Gagnon et al., 1996a). Non-local mixing as a result of bioturbation in marine sediments has only recently been clearly

documented and modeled, particularly with respect to *“Pb profiles (e.g., Soetaert et al., 1996). The non-local

mixing affects the diagenetic behaviour of sulphur as well as a number of trace metals. High dissolved Hg peaks

(over 100 ng l- ’ ) were also recorded in sediments below 15 cm and even 25 cm depth (in Fig. 3; see examples SAG-5 and SAG-30). As the oxidative release of mercury associated with iron sulphide minerals can be an important remobilization process (Morse, 1994), peaks located in anoxic sediment layers may be due to

localized bioturbation leading to the partial oxidation of sulphides and the release of Hg to the porewaters as conceptually illustrated by the following scheme:

where S,O;? represents partially and fully oxidized sulphur species. The formation of stable, soluble mercury

complexes with products of the partial oxidation of sulphides and iron sulphides (e.g., polysulphides, thiosulphate) at these depths could explain anomalies in dissolved Hg profiles. Oxidation of iron sulphides may be an important process since the most abundant reduced sulphur species in Saguenay Fjord sediments are

oxygen-reactive metastable iron monosulphides (Gagnon et al., 1995). Consequently, these AVS-rich sediments may impart an uncharacteristically high reactivity to the sedimentary Hg, since it would be susceptible to

remobilization upon exposure to an oxic environment (e.g., resuspension, dredging, bioturbation, etc.). On the other hand, diffusion of mercury through the sedimentary column is strongly inhibited by its high affinity for

particulate organic matter. The irrigation of worm tubes and physical disruptions of the sediment-water interface, however, could accelerate the transport of dissolved mercury through the oxic layer to the water

column.

5.4. Transport to the sediment-water inter$ace

The sediment porewaters from the landward basin of the fjord are about 12 times (sometimes over 100 times) richer in mercury than the overlying bottom seawater concentrations (I 4 ng 1-l) reported by Gobeil et al.

170 C. Gagnon et al./Marine Chemistry 59 (1997) 159-176

Table 1 Calculated fluxes of dissolved Hg at the sediment-water interface

Core Fhtx a (ng cm-* yr-‘)

SAG-5-91 b 36.0

SAG-6-92 20.2

SAG-6A-92 26.5

Mean 21+7

SAG-15-91 18.0

SAG-15-92 13.9

SAG-15A-92 24.9

Mean 19+5

SAG-30-91 10.1

SAG-30-92 8.5

Mean 9+1

CL-l-91 18.9

CL-l-92 3.8 Mean 11+8

’ The diffusion fluxes were calculated from the gradient concentration following Eq. (5). The concentration gradients were evaluated from

total concentrations in the porewaters extracted from the first sediment interval (i.e., O-O.5 cm, thus az = 0.5 cm) and in the overlying

bottom seawater ( I 4 ng 1-l : Gobeil et al. 1984). The precision of our flux estimates could be on the order of Ifr 20%.

b The core numbers indicate the location and the year of the sampling. In 1992, two cores were analyzed at SAG-6 and SAG-I 5.

(1984). The total Hg diffusive flux, J, at the sediment-water interface was estimated according to Fick’s first

law (Berner, 1980):

J = - @D,( ac/az) (5)

where @ is the porosity, 4, is the bulk sediment diffusion coefficient, and (Z/az) is the concentration

gradient across the sediment-water interface. The value of 0, was assumed to be equal to a2Do (Ullman and Aller, 1982) where Do is the molecular diffusion coefficient. In accordance to Gobeil and Cossa (1984) and as recommended by Bothner et al. (1980), we adopted a value of Do for mercury in seawater equal to 5 X 10e6

cms2 SK’. The concentration gradient was calculated from the total dissolved Hg concentrations in the bottom waters (Gobeil et al., 1984) and porewaters collected from the first sediment sampling interval (i.e., O-O.5 cm,

thus AZ. = 0.25 cm). Flux values calculated from Eq. (5) ranged from 3.8 to 36.0 ng cme2 yr- ’ (Table 1).

These values are similar to fluxes (l- 19 ng cme2 yr- ‘) estimated by Gobeil and Cossa (1984, 1993) in the Laurentian Trough and the landward basin of the fjord. Nevertheless, these values are likely overestimated given

our poor vertical resolution of the profile and the affinity of Hg for fresh organic matter and authigenic Fe oxides accumulating near the sediment-water interface. Fluxes estimated from porewater profiles measured on

sediment cores collected at the same station at one year intervals or even the same year are quite variable (e.g., CL- l-91 and CL-l-92). This variability can be attributed to spatial heterogeneity of the sediment as well as the

density and activity of benthic organisms.

Despite intra-station variations, flux estimates were observed to decrease ( _ 70%) progressively downstream from the head of the fjord towards the St. Lawrence Estuary. We propose that the barrier (i.e., oxic sediments) to the diffusion would be least efficient upstream because sediments are more reducing and the 0, penetration depth is shallower. This interpretation is consistent with in situ flux measurements (Bothner et al., 1980) which demonstrated that only sediments exhibiting strongly reducing conditions at the sediment-water interface can release detectable amounts of mercury to the water column. The estimated Hg flux values in the Saguenay fjord

were much lower than those (379 to 884 ng cme2 yr- ‘) measured above contaminated sediments of Bellingham Bay (Bothner et al., 1980) where low oxygen or anoxic conditions prevail in bottom waters of the bay. The presence of oxidized surface sediments, therefore, inhibits transfer of Hg to the overlying bottom waters. The simultaneous HCl-extraction of high levels of mercury (Hgn,,), iron (Fe,,,) and manganese

C. Gagnon et al/Marine Chemistry 59 (1997) 159-176 171

WATER COLUMN

Diffision Burial 200 kg/yr Diffusion

I CO.03 kg/yr

?

I i

OXIC SEDIMENT I -----------~ 1 max. 50 kg/yr \ \OJ kg/yr

r--------- I-_C_____________L~___________

I Associated - Hg(II) - Dissolved

FeS - Hg SW Hg Fe% - Hg S203- Hg RS- Hg RS-Hg

2x 104kg ++ Hg’

L

I ANOXIC SEDIMENT I

Fig. 5. Reservoirs and fluxes of mercury in Saguenay fjord sediments.

(Mn,,,) in the first cm of the sediments (Fig. 4) indicates that iron and manganese oxides probably scavenge

mercury diffusing to the sediment surface, thereby limiting its release to the overlying bottom waters. We estimated the integrated flux of Hg for the entire fjord by multiplying the mean flux values by the

approximative surface areas of the fjord head (34 km’) and landward basin (102 km2>. According to these

calculations, the flux of methylmercury is < 0.03 kg yr-’ and represents an insignificant fraction (0.2%) of the

total Hg diffusing across the interface (14 kg yr-’ ; Fig. 5). Despite a strong gradient within the sediment

column, little CH,Hg(II) escapes by molecular diffusion to the overlying water column (see Fig. 5) because oxidized sediments appear to act as an efficient barrier. Bioturbation may, however, decrease the efficiency of

this barrier for both Hg(I1) and CH,Hg(II). Tubicolous worms and even their relict burrows (Ray and Aller, 1985) may significantly increase the water-sediment exchange. Fluxes at the sediment-water interface may be

2- 10 times larger than those calculated on the basis of molecular diffusion due to the activity of macrobenthos

(Rutgers van der Loeff et al., 1984). Assuming maximum transport enhancement by benthic activity, the amount of Hg escaping from the sediments over the length of the fjord would be as much as 140 kg yr-’ or about 70%

of the Hg accumulation rate (200 kg yr ’ : see Fig. 51, estimated by multiplying the mercury concentration in

the surficial sediment (* 500 ng g-l> by the sedimentation rate (N 0.3 g cme2 yr- ‘) and the surface area of the landward basin. The extent of flux enhancement cannot, however, be evaluated precisely due to the lack of

172 C. Gagnon et al. / Marine Chemistry 59 (1997) 159- 176

information on the abundance and distribution of benthic species in the fjord sediments. On the other hand,

dissolved mercury concentrations in the deeper saline water column cannot justify the strength of these

biologically enhanced fluxes even though some of the remobilized Hg could have been scavenged by the

suspended particulate matter (SPM) for which it has a strong affinity. Finally, it should be mentioned that mercury could also be lost from sediments as gaseous elementary Hg’.

Both methylmercury and inorganic mercuric ions may be reduced to volatile Hg” through biological and

chemical reactions (Gilmour and Henry, 1991; Mason et al., 1993). Although the loss of Hg” from the water column to the atmosphere may be faster than methylation/demethylation rates, direct volatilization from

sediments was reported to be negligible relative to methylation and demethylation rates (Steffan et al., 1988).

5.5. Contribution to the sediment sur$ace

The accumulation of Hg-rich particles and remobilization of Hg from deeper sediment layers may explain the

persistence of anomalously high Hg concentrations in surface sediments in the Saguenay fjord. We believe, however, that the deposition of Hg-rich particulate matter originating from the resuspension of contaminated

sediments upstream may be the dominant source. Preliminary analyses of total Hg in SPM collected at 12

stations along the main axis of the fjord gave a range of concentrations between 0.25 and 7 pg g- ’ (average 3.7

pg g-l, n = 75). These values appeared to be independent of depth and sampling location (A. Mucci, unpubl.

data). Given that mercury concentrations in surface sediments are u 0.5 pg g - ‘, settling and accumulation of

this SPM would more than account for the observed sediment concentrations; a significant fraction of Hg

associated with this SPM would be released back to the water column following the degradation of labile organic matter at the sediment-water interface. It is interesting to note that the Hg content of SPM has often

been observed to be much larger than in the underlying sediment (e.g., Cranston and Buckley, 1972). The most common explanation is that small particles which remain suspended in the water column have a larger surface to volume ratio than the sedimentary material and are capable of sorbing more Hg.

The diffusive flux of Hg within the sediment column to the surficial sediments was estimated in the

Laurentian Trough by Gobeil and Cossa (1993) at 0.95 ng cmd2 yr- I, representing only 2.6% of the present deposition rate. This flux was calculated from the maximum porewater Hg concentration gradient located at the

depth of Fe-Mn oxide reduction (see Fig. 3). Unfortunately, with one exception, our data do not allow us to

evaluate the flux of Hg from the anoxic to the surficial oxic sediments because no obvious concentration

gradient could be distinguished. A strong gradient was observed at SAG-30-92 from which we estimated a

maximum flux at this station. The gradient used in the calculation (Eq. (5)) was obtained from the first cm interval (AZ = 1 cm> below the depth of Fe-Mn oxide reduction (i.e., z = 2-3 cm>. The maximum flux was

estimated to be on the order of 37 ng cm-* yr-’ or 50 kg yr- ‘, if we extrapolate this flux over the entire fjord. This maximum value would signify that 25% of the Hg found in the surficial sediments at this station could

have a diagenetic origin. Thus, diagenetic remobilization of Hg does not appear to be significant in these coastal marine sediments. In contrast, significant remobilization of Hg was reported in Great Lakes sediments with diffusive fluxes representing a high proportion (36-96%) of or even exceeding sedimentation fluxes (Matty and

Long, 19951. Finally, even though methylation has been hypothesized to increase the mobility of mercury in aquatic

systems (Parks et al., 19891, it does not contribute significantly to its remobilization in the Saguenay fjord sediments. The low CH,Hg(II) flux (0.1 kg yr-‘) to the surface sediments calculated for the entire fjord from concentration gradients across the redox boundary represents only 0.05% of the present accumulation rate of mercury at the sediment surface.

6. Geochemical box model for the Saguenay fjord sediments

This work provides new information about processes contributing to the relative mobility of Hg in coastal sediments. Fig. 5 summarizes the fate of Hg buried in Saguenay fjord sediments and provides a scenario of its

C. Gagnon et al. /Marine Chemistry 59 (I 9971 I59- 176 173

diagenetic behaviour. Total mercury and methylmercury inventories were estimated for both the oxidized

surface and recent anoxic sediments from average concentration values over our sampling depth and the surface area of the study area. The total anthropogenic mercury inventory was estimated at 2 X lo4 kg f 20% whereas

CH,Hg only accounts for 0.2% of this amount. Our estimate (20 t) is about three times lower than the one reported in an earlier study (60 t> by Loring and Bewers (1978). This discrepancy arises because box cores

(N 40 cm> recovered at the head of the fjord where the sedimentation rate is highest did not sample the whole record of Hg discharge. Most of the anthropogenic mercury at SAG-5/6 is found beyond the sampling depth of

the box corer (see Fig. 2). Exchanges of total Hg and CH,Hg between the various reservoirs of the system (i.e., overlying water, anoxic

and oxic sediments) were also evaluated from estimated fluxes for the entire study area (Fig. 5). They reveal weak diffusion fluxes relative to particulate Hg deposition. Inorganic Hg is released to the sediment porewaters

by processes such as the degradation of organic matter, the dissolution of authigenic and detrital Fe-Mn oxides,

as well as the oxidation of Hg-laden iron sulphides. Once in solution, Hg may be readsorbed to residual organic

matter or iron sulphides (FeS and FeS,). The solubility of Hg in anoxic sediments may be enhanced by the

formation of complexes with intermediate reduced sulphur species (organic sulphides, polysulphides, and thiosulphates) as well as with dissolved organic matter (e.g., humic compounds). If a concentration gradient is

established, dissolved Hg in the anoxic sediments may diffuse to the sediment-water interface. The presence of an oxidized sediment layer, however, limits its release to the water column due to adsorption onto Fe-Mn oxides and fresh organic matter. Dissolved methylmercury species which accumulate in the anoxic zone also

diffuse to the oxidized sediment layer but, in contrast to inorganic Hg(II), appear to be demethylated rather than adsorbed onto solid components of the sediment. The diffusion of methylmercury, however, does not contribute

to the surface sediment Hg enrichment since its flux from the anoxic sediments is negligible.

7. Conclusion

Results of this study show that, despite Hg enrichment in the interstitial waters, diagenetic remobilization

alone cannot account for the abnormally high Hg concentrations in the surticial sediments of the Saguenay fjord.

Furthermore, molecular diffusion of inorganic and metbylmercury to the overlying waters appears to be inhibited by the presence of an oxic layer near the sediment-water interface. Nevertheless, transport to the

overlying waters may be enhanced by benthic activity. The abundance of macrobenthos can be high in coastal sediments and may be a major factor controlling both the water-sediment exchange and transfer to pelagic organisms. More data are required on the abundance, distribution and Hg levels of organisms inhabiting the

contaminated sediments in order to determine their role on the transfer of Hg to higher organisms. Finally, the persistence of high Hg concentrations in surficial sediments in specific areas of the fjord may be explained by

resuspension of contaminated sediments upstream (erosion of contaminated sediments during spring runoff and submarine mass flows) and settling in the fjord as well as by biological sediment mixing.

Acknowledgements

We thank A. Bono and the crew of the research vessel Alcide C. Horth for their assistance during sample collection. We are grateful to J. Noel and C. Guignard for their technical assistance. The authors also thank D.

Cossa and B. Sundby for their helpful comments. This article also benefitted from the reviews of B. Lasorsa, J.W. Murray and one anonymous referee. This study was funded by St. Lawrence Centre/NSERC and DFO/NSERC Joint Subvention Program grants as well as NSERC Operating grants to A.M. and E.P.C. Gagnon also acknowledges the financial support of the FCAR fund. This publication is a contribution of the Oceanographic Centre of Rimouski, a partnership of INRS (Institut National de la Recherche Scientifique) and UQAR (UrGersite du Q&bee h Rimouski) operating under the auspices of the University of Quebec.

174 C. Gagnon et al./Marine Chemistry 59 (19971159-176

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